Organic Micropollutants in the Aquatic Environment: Proceedings of the Fifth European Symposium,...
457
ORGANIC MICROPOLLUTANTS IN THE AQUATIC ENVIRONMENT
Transcript of Organic Micropollutants in the Aquatic Environment: Proceedings of the Fifth European Symposium,...
ORGANIC MICROPOLLUTANTS
IN THE AQUATIC ENVIRONMENT
This symposium was jointly organised by
and
the Commission of the European Communities, Directorate-General for Science, Research and Development, Brussels (Belgium)
the "Consiglio Nazionale delle ricerche, Istituto di Ricerca sulle Acque, Roma (Italy)
ACKNOWLEDGMENT Special acknowledgment is due to Mr Antonio Ruberti, Ministro della Ricerca Scientifica and to Montedison, ESSO and ACEA for their support.
This is report 4 in the series "Water Pollution Research Reports" issued by the Commission of the European Communitie's Directorate-General XII, Division E-l 200, rue de la Loi B - 1049 Brussels (Belgium)
Commission of the European Communities
ORGANIC MICROPOLLUTANTS
IN THE AQUATIC ENVIRONMENT
PROCEEDINGS OF THE FIFTH EUROPEAN SYMPOSIUM, HELD IN ROME, ITALY, OCTOBER 20-22, 1987
Edited by
G. ANGELETTI Directorate-General for Science, Research and Development,
Commission of the European Communities, Brussels, Belgium
and
A. BJ0RSETH SCATEC, Slependen, Norway
KLUWER ACADEMIC PUBLISHERS DORDRECHT I BOSTON I LONDON
Library of Congress Cataloging in Publication Data
Organic .tcropollutants in the aquatic envtronment , proceedtngs of the fifth European sYlposiu. held in Rooe. Italy. October 20-22. 1987 I edited by G. AngeleTti and A. Bjlrseth.
p. CI.
At head of title, Co •• ission of the European Co •• unitles. ·Contains oral papers and posters presented at the 'Flfth European
Symposium on Organic Micropollutants in the Aquatic Environlent'"-Foreword.
Jointly organised by the Co •• ission of the European Co •• unities. Directorate-General for Science. Research. and Development. and the Consiglio na2ionale delle ricerche. IstltuTo dt rtcerca sulle acque.
Includes biblIographIEs and tndex. ISBN-13: 978-94-010-7843-6 .-ISBN-13: 978-94-009-2989-0 DOI: 10.1007/ 978-94-009-2989-0
1. Organic .ater pollutants--Congresses. 2. Water--Purification-Congresses. 3. Aquatic ecology--Congresses. ~. Water che.istry--Congresses. I. Angeletti. G .• 19~3- II. Bjorseth. Alf. III. Co.mission of the European Co.muniTies. DireCTorate-General for Science. Research. and Develop.ent. IV. Istituto di ricerca sulle acque (Italy) V. European Symposium on Organic MicropolluTants in thE Aquatic Environment (5th 19B7 ROle. ltaly) T~27.0707~ 19BB 62B.l·6B--dc19
ISBN-13: 978-94-010-7843-6
Publication arrangements by Commission of the European Communities
B8-6606 CIP
Directorate-General Telecommunications, Information Industries and Innovation, Luxembourg
EUR 11350 © 1988 ECSC, EEC, EAEC, Brussels and Luxembourg Softcover reprint of the hardcover 1st edition 1988
LEGAL NOTICE Neither the Commission of the European Communities nor any person acting on behalf of the Commission is responsible for the use wliich might be made of the following information.
Published by Kluwer Academic Publishers. P.O. Box 17, 3300 AA Dordrecht, The Netherlands.
Kluwer Academic Publishers incorporates the publishing programmes of D. Reidel, Martinus Nijhoff, Dr W. Junk and MTP Press.
Sold and distributed in the U.S.A. and Canada by Kluwer Academic Publishers, 101 Philip Drive, Norwell, MA 02061, U.S.A.
In all other countries, sold and distributed by Kluwer Academic Publishers Group, P.O. Box 322,3300 AH Dordrecht, The Netherlands.
All Rights Reserved No part of the material protected by this copyright notice may be reproduced or utilized in any form or by any means, electronic or mechanical, including photocopying, recording or by any information storage and retrieval system, without written permission from the copyright owner.
FOREWORD
This volume contains oral papers and posters presented at the "Fifth European Symposium on Organic Micropollutants in the Aquatic Environment" held in Rome (Italy) from 20 to 22 October 1987. The Symposium was organised within the framework of the Concerted Action COST 641*) which is included in the Fourth R&D Programme on the Environment of the Commission of the European Communities.
As for the previous symposia, the aim was to review current studies and technical progress in the area of organic micropollutants in the aquatic environment, particularly since the last symposium held in Vienna in October 1985. The programme consisted of review papers and posters related to analytical methodologies, transport and transformation of organic micropollutants in water, water treatment processes and mathematical modelling. Special sessions were devoted to laboratory data treatment and environmental scenarios.
These proceedings provide a good overview of the activities in this fieiu in Europe and constitute a valuable contribution to the understanding and solution of the problems posed by organic micropollutants in the aquatic environment.
The Commission of the European Communities wishes to express sincere thanks to the co-organizers of the Symosium, in particular to Mr A. Liberatori and T. La Noce of the Consiglio Nazionale delle Ricerche, Istituto di Ricerca sulle Acque, Rome.
Brussels/Oslo, November 1987
G. ANGELETTI A. BJ0RSETH
*) COST 641: Scientific and Technical Cooperation among European Community Member Countries and the Non-Member Countries Finland, Norway, Sweden and Switzerland, in the field of "Organic Micropollutants in the Aquatic Environment"
-v-
CON TEN T S
Foreword
SESSION I ANALYTICAL METHODOLOGIES
The use of low cost mass spectrometers for the analysis of organic micropollutants in water
S.P. SCOTT, R.L. KEELING, H. JAMES, A. WAGGOTT, and P. WHITTLE, Thames Water, Water Research Centres and North West Water
Coupling micro-LC capillary GC as a tool in environmental analysis
D. DUQUET, RSL-Alltech Europe, Eke; C. DEWAELE, Laboratory for Organic Chemistry, State University of Gent
Standardized analytical methods for EC priority pollutants S. SCHMIDT, BUro Dr. Schmidt, Leverkusen
The application of bonded silica extraction columns in sample preparation prior to the analysis of organic micropollutants in water
R.A. CALVERLEY, Analytichem International, Cambridge
POSTER SESSION I ANALYTICAL METHODOLOGIES
Evaluation of steam distillation-extraction procedure for the recovery of phenols in water
M.T. GALCERAN and F.J. SANTOS, Department of Analytical Chemistry, University of Barcelona
Isomer-specific determination of PCDD/PCDF in water leachate of a waste landfill
C. FORST, L. STIEGLITZ and G. ZWICK, Kernforschungszentrum Karlsruhe
Pesticide micropollutants in Lake Albufera (Spain) J .M. CARRASCO, Departamento de Biotecnologia, Universidad Politecnica de Valencia
- vii-
2
14
22
31
46
52
59
Techniques for the improvement of gasoline analysis P. SLINGERLAND and R.C.C. WEGMAN, Laboratory for Organic-analytical Chemistry, National Institute of Public Health and Environmental Protection, Bilthoven 62
An improved extraction method for the quanti tati ve analysis of pesticides in water
H.F. SCHOELER and J. BRODESSER, Hygiene-Institute University of Bonn
Hyphenated methods (TSP LC-MS, DLI LC-MS, LC-TID) for analyzing organophosphorus priority pollutants
D. BARCELO and J. ALBAIGES, Environmental Chemistry
69
Department, CID (CSIC), Barcelona 75
PCB's and organochlorine pesticides in eel and flounder in the Tagus estuary
M.J. BENOLIEL and M.L. SHIRLEY, Instituto Hidrografico, Lisboa
HPLC/Fluorescence spectrometry in analyses of pulp mill wastes in recipients
P. MIKKELSON, J. PAASIVIRTA and J. KNUUTINEN, Department of
83
Chemistry, University of Jyvaskyla 88
Chlorophenol compounds in snow R. PAUKKU, Institute for Environmental Research, University of Jyvaskyla; J. PAASIVIRTA and M. KNUUTILA, Department of Chemistry, Uni versi ty of Jyvaskyla; S. HERVE, Water and Environment District of Central Finland, Jyvaskyla 91
The analysis of odorous sulphur compounds by gas chromatography after thermal desorption from tenax
LW. DAVIES and J. YATES, SAC (Chromatography) and Water Research Centre
Determination of organic chemicals in sediments taken from three unpolluted estuaries in South West England
B.J. HARLAND, ICI Brixham Laboratory; R.W. GOWLING, Trent
97
Polytechnic, Nottingham 103
Organic phosphates in surface, ground and drinking water S. GALASSI, Water Research Institute, CNR, Milano; L. GUZZELLA, Biology Department, University of Milan 108
Evaluation of degree of pollution of Tiber and Aniene rivers by nitrilotriacetic acid
L. ZOCCOLILLO, G.P. CARTON I , M. RONCHETTI and A. DE LOGU, Dipartimento di Chimica, Universita di Roma "La Sapienza" 116
- viii-
Hydrocarbons in East Mediterranean sea determination and occurrence in the sediment of considered polluted and unpolluted areas of coastal environment
M. PSATHAKI, M. ZOURARI and E. STEPHANOU, Laboratory for Environmental Chemistry, Department of Chemistry, University of Crete, Iraklion 121
SESSION II TRANSPORT OF ORGANIC MICROPOLLUTANTS IN THE AQUATIC ENVIRONMENT
The Sandoz accident B. HURNI, Amt fUr Umweltschutz und Energie, Kantons Basel-Landschaft, Liestal 128
Monitoring of the River Rhein Experience gathered from accidental events in 1986
H. FRIEGE, Landesamt fUr Wasser und Abfall 132
Predicting transport behaviour of organic pollutants using simple mathematical models
P. S. GRIFFIOEN, Institute for Inland Water Management and Waste Water Treatment, Lelystadj D. VAN DE MEENT, National Institute of Public Health and Environmental Hygiene, Bilthoven 144
Fate and transport of organic compounds in rivers C.D. WATTS and K. MOORE, WRc Environment, Medmenham Laboratory 154
POSTER SESSION II TRANSPORT
Sediment-water partition coefficients of hydrophobic chemicals in the presence of third phase material
S.M. SCHRAP and A. OPPERHUIZEN, Laboratory of Environmental and Toxicological Chemistry, University of Amsterdam 170
Environmental fate of organosilicon chemicals A. OPPERHUIZEN, G.M. ASYEE and J.R. PARSONS, Laboratory of Environmental and Toxicological Chemistry, University of Amsterdam 176
Pollution of Saronicos Gulf (Athens, Greece) by fossil fuel hydrocarbons
A. MYLONA and N. MIMICOS, E. STEPHANOU, Laboratory for University of Crete, Iraklion
-ix-
NRC Democri tos, Athens j Environmental Chemistry,
184
The Sandoz/Rhine accident : The environmental fate and transport of twenty-one pesticides introduced to the Rhine River
P.D. CAPEL and W. GIGER, Swiss Federal Institute for Water Resources and Water Pollution Control (EAWAG), DUbendorf 189
Occurrence and leaching of pesticides in waters draining from agricultural land
S. REKOLAINEN, National Board of Waters and Environment 195
Polychlorinated biphenyls in the Kupa river, Croatia, Yugoslavia Z. SMIT and M. KODRIC SMIT, Medical Centre Sisak, Department of Sanitary Chemistry and Ecology, Sisak; V. DREVENKAR, Institute for Medical Research and Occupational Health, University of Zagreb
SESSION III - TRANSFORMATION OF ORGANIC MICROPOLLUTANTS IN WATER
Biodegradation of chlorinated aromatic chemicals in continuous cultures
J.R. PARSONS, D.T.H.M. SIJM and M.C. STORMS, Laboratory of Environmental and Toxicological Chemistry, University of
198
Amsterdam 206
Anaerobic degradation, processes and test methods G. SCHRAA, Department of Microbiology, University, Wageningen
The fate of organic compounds in the environment
Agricultural
A.H. NEILSON, A.-S. ALLARD, C. LINDGREN and M. REMBERGER,
215
Swedish Environmental Research Institute 228
POSTER SESSION III TRANSFORMATION
Levels of chlorophenols in the river, ground and drinking water in the Zagreb area
S. FINGLER and V. DREVENKAR, Institute for Medical Research and Occupational Health, University of Zagreb
SESSION IV - WATER TREATMENT
Biological-chemical characterization of effluents for the evaluation of the potential impact on the aquatic environment
O. SVANBERG and L. RENBERG; National Environmental Protection Board, The Emission and Product Control
238
Laboratory, Solna 244
-x-
Test methods and strategies for environmental management purposes - environmental fate testing of chemicals and effluents
N. NYHOLM, Water Quality Institute 256
Mass fluxes of linear alkylbenzenesulphonates, nonylphenol, nonylphenol mono- and diethoxylate through a sewage treatment plant
A. MARCOMINI, Department of Environmental Sciences, Uni versi ty of Venice; S. CAPRI, Water Research Institute (IRSA), Roma; P.H. BRUNNER and W. GIGER, Swiss Federal Insti tue for Water Resources and Water Pollution Control (EAWAG), DUbendorf 266
Mutagenic compounds in chlorinated waters B. HOLMBOM, Abo Akademi, Laboratory of Forest Products Chemistry; L. KRONBERG, Abo Akademi, Department of Organic Chemistry 278
The formation and removal of chemical mutagens during drinking water treatment
M. FIELDING and H. HORTH, WRc Environment, Medmenham Laboratory
Application of the ozone-hydrogen peroxide combination for the removal of toxic compounds from a groundwater
J . P . DUGUET, C. ANSELME, P. MAZOUNIE and J. MALLEVIALLE,
284
Centre de Recherche Lyonnaise des Eaux-Degremont, Le Pecq 299
POSTER SESSION IV WATER TREATMENT
Possibilities and limitations of the combined use of ozone and hydrogen peroxide in drinking water preparation from surface water
F. VAN HOOF, J. JANSSENS and E. PLUYS, Study Center for Water, Antwerp 312
Presence of polycyclic aromatic hydrocarbons in surface waters used for the production of drinking water
F. VAN HOOF and S. AERTS, Studiecentrum voor Water, Antwerp 318
Research and behaviour of organic micropollutants from waste distillery wine in anaerobic treatment
S. SANCHEZ CRESPO and J. PRADA ALVAREZ-BUYLLA, Confederacion Hidrografica del Guadiana 323
Mass spectrometric identification of halogenated surfactants in Barcelona's water treatment plant
J. RIVERA, J. CAIXACH, I. ESPADALER, A. FIGUERAS and M. DE TORRES; Lab. Espectrometria de Masses, CID - CSIC, Barcelona; F. VENTURA, AigUes de Barcelona; D. FRAISSE, Service Central d'Analyse, CNRS, Vernaison 329
- xi-
Effects of chlorine dioxide preoxidation on organic halide formation potentials
H. BEN AMOR, J. DE LAAT and M. DORE, Laboratoire de Chimie de l'Eau et des Nuisances, Poi tiers 338
NMR study of kraft pulp mill waste and natural humic substances J. VIRKKI, J. KNUUTINEN, P. MANNILA and J. PAASIVIRTA, Department of Chemistry, University of Jyvaskyla 344
Identification of bioaccumulable compounds in kraft bleaching effluents
G. E • CARLBERG, H. DRANGSHOLT, N. GJ¢S and L. H. LANDMARK, Center for Industrial Research, Oslo
Influence of humic water substances on the degradation of PAH during water chlorination
G. BECHER, National Institute of Public Health, Oslo;
347
S. JOHNSEN, Center for Industrial Research, Oslo 353
Influence of waste water disinfection treatments on some genotoxic chemical micropollutants
A. SAVINO, R. PASQUINI and R. CONTI, Dipartimento di Igiene, Universita degli Studi di Perugia; C. MELCHIORRI, A. 01 CARO, L. SEBASTIANI, A. GRELLA and S. BONACCI, Istituto di Igiene, Universita degli Studi "La Sapienza", Roma
SESSION V - MATHEMATICAL MODELLING
Evaluation of some chemical fate and transport models - A case study on the pollution of the Norrsundet Bay (Sweden)
K. KOLSET, B.F. ASCHJEM, N. CHRISTOPHERSEN, A. HEIBERG and B. VIGERUST, Center for Industrial Research, Blindern
Modelling of groundwater transport of microorganic pollutants State-of-the-art
F. DE SMEDT, Laboratory of Hydrology, Vrij e Uni versi tei t
357
372
Brussel 387
POSTER SESSION V - MODELLING
Modelling of surfactants in the Comunidad de Madrid as subbasin of Tagus River
F. CUBILLO, Models and Technology Division, Comunidad de Madrid 402
- xii-
Modelling of anthropogenic substances in aquatic systems : MASAS - A personel computer approach
R.P. SCHWARZENBACH, J. WETZEL, J. HELDSTAB and D.M. IMBODEN, Swiss Federal Institute for Water Resources and Water Pollution Control (EAWAG), DUbendorf
SESSION VI LABORATORY DATA TREATMENT
Chemometrics in environmental analytical chemistry H.A. VAN 'T KLOOSTER, National Institute of Public Health and Environmental Protection (RIVM) , Laboratory of Organic-Analytical Chemistry, Bilthoven
SESSION VII - ENVIRONMENTAL SCENARIO
Future environmental problems F. BRO-RASMUSSEN, Laboratory of Environmental Science and Ecology, Technical University of Denmark, Lyngby
LIST OF PARTICIPANTS
INDEX OF AUTHORS
- xiii-
408
416
432
441
451
SESSION I
ANALYTICAL METHODOLOGIES
Chairmen D. QUAGHEBEUR and B. CRATHORNE
The use of low cost mass spectrometers for the analysis of organic micropollutants in water
Coupling micro-LC capillary GC as a tool in environmental analysis
Standardized analytical methods for EC priority pollutants
The application of bonded silica extraction columns in sample preparation prior to the analysis of organic micropollutants in water
THE USE OF :r.ru COST MASS SPECTRCl1ETERS FOR THE ANALYSIS OF ORGANIC MICROPOILUTANTS IN WATER
S • P • scarr, R. L. KEELING, H. JAMES, A. WAQXYIT, P • WHI'ITLE Thames Water, water Research Centre and North West Water, UK
Surrrnary
Mass Spectrometry in combination with gas chromatography (GC-MS) has been applied to the analysis of organic compounds in water for over a decade. Initially, because of the high costs involved, only a few central research laboratories possessed such equipment. However, in the last few years, mass spectrometer manufacturers have produced smaller bench-top nass spectrometers at lower prices so that GC-MS facilities have become more widely available. The UK water undertakings have, collectively, considerable experience of the two most commonly available bench-top GC-MS systems - the Finnigan Ion Trap Detector (lTD) and the Hewlett-Packard Mass Selective Detector (MSD). Typical applications of these two systems to the analysis and monitoring of various organics in water will be described, and factors such as ease of use, reliability and the qt21ity of the data produced by the lTD and MSD will be discussed.
1. INTRODUCTION --------Mass Spectrometers have been in use in the UK water industry since
the late 1970's, generally limited to one per authority. Their major use has been in identifying organic micropollutants encountered in survey analysis, pollution incidents or taste and odour operational problems; any qt2ntitative work being limited to a few taste and odour compounds. Approximately two years ago low cost bench-top mass spectrometers became available. Since then the need for more qt2ntitative organic analysis conillined with the recognition that a mass spectrometer can be a highly specific as well as sensitive detector and, more importantly, the low cost has produced a minor explosion within the industry. Some six Finnigan Ion Trap Detectors (ITD) and five Hewlett-Packard Mass Selective Detectors (MSD) are now in use.
2. USES Over the last two years the perfornance of these small bench-top
Ihlchines has been shown to be perfectly adequate for routine qt2litative broad scan analysis of complex extracts from sewage sludges, groundwaters, surface water and rain. Of particular importance has been the speedy identificatlOn of pollutants or the confirnation of pesticide findings obtained using ECD/FPD/PlD.
Their prime application however is the quantative analysis of specific compounds.
2.1 GEOSMIN At~D 2-MEl'HYLlSOBORJ.\JEOL One of the first methods to be evaluated on the bench-top
instruments has been the determination of geosmin and 2-methylisoborneol. With the odour threshold in the order of 10 n9/1
-2-
the need for a fast, accurate and sensitive measurement is essential if the high cost of activated carbon treatment at the water treatment plant is to be kept to a minimum.
1-Chloro-n-alkanes are added, as an internal standard, to 1 litre of sample after which 200g of sodium chloride is added. The sample is then purged for one hour using a closed loop stripping apparatus. The carbon filter being extracted with dichloromethane. (1) + (2)
Reproducibi~ity test: lTD, A. Waggott, WRc Processes Amount injected: 2 ul standards containing 750 pg/ul
Absolute TIC response (peak area) Relative Response
1-chloro-n-decane 2-methylisoborneol geosmin MIB/NCO G/NCO
99582 67711 41016 0.68 0.41 166939 101970 91841 0.61 0.48 105910 69565 54712 0.66 0.52 168373 102822 66229 0.61 0.39 197614 120968 88060 0.61 0.45 184497 97483 74875 0.53 0.41
Even with the use of three internal standards (C8HJ7Cl, C10H21Cl, C 2H2~Cl) the selectivity of the mass spectrometer can De hinderea by t~e c6-elution of interfering ions, especially when the ions used are, by necessity, in the low mass range. This problem has recently been overcome by the use of synthesised deuterated geosmin and 2-methylisoborneol as internal standards.
2.2 PH&~OIS To 100ml of sample is added 100ng of 2,6-difluorophenol after which
2ml of 5N sodium hydroxide and 30ml of 1N sodium bicarbonate are added. This is then followed by 20ul of neat pentafluorobenzoyl chloride and 2ml of isoctane. The sample is sh~{en for 5 minutes after which sodium hydroxide is added to remove excess reagent. The sample is allowed to stand for 1 hour before the isoctane is ren~ved. (3) No initial cleanup is done and any sample producing an emulsion will require the extract being fBSSed through anhydrous sodium sulphate. calibration of the method is done by spiking Milli-Q water with standards in the range of 50-1000 ng!l. The ion used for guantification for all phenols is 195 (the pentafluorobenzoyl ion) with the molecular ion as a qualifier.
The similarity of the spectrum of different phenolic isomers has enabled, their presence to be reported even though they may not have been included in the original calibration. Because of its ubiquitous nature it has been difficult to obtain a phenol blank below approximately 200 ng!l, while pentachlorophenol is poorly derivitized and has a detection limit of 100 ng/l. Lower levels could be achieved if larger sample volumes were used.
2.3 PESTICIDES 2 litres of samples are spiked with 40 ng of deuterated (06)
lindane. After the sample has been shaken for 5 minutes with 50ml of hexane, the extract is passed through anhydrous sodium sulphate and evaporated to O.2ml. No clean-up procedure has been found to be necessary for surface and finished waters encountered so far.
-3-
Calibration is obtained by spiking laboratory tap water with standards ranging from 1-50 ng/l. Two ions are employed for each compound, one for quantification and the other for qualification.
The use of the bench-top mass spectrometers has enabled the measurement of compounds which, depending upon the type and manufacturer of capillary column, will co-elute and give significant ECD responses.
E.g. Thames Water : f-HCH, Hexachlorobenzene, Simazine, Atrazine,Propazine,~HCH and trischloroethyl-phosphate
North West Water: lr-HCH, Diazinon and Propet.arrq;>hos South West Water: ~HCH and tetrachloromethylthiobenzene (a
possible degradation product or impurity in Tecnazene).
The above three methods provide examples of the three types of internal standard currently being employed. Both the MSD and lTD are ideally suited for operation with internal standards and in particular deuterated or carbon-13 internal standards. Because of the improvements in ease of operation, precision and accuracy of results inherent in this approach it is recommended that where practically possible all determinands of environmental concern be analysed by this method. The high cost of synthesis of these labelled compounds means great care must be taken to ensure the ions used are sufficiently resolved from those of the unlabelled species, especially when a high degree of halogenation is present.
3. ADVANTAGES The relatively low cost has enabled the buying of either one, where
none would have been considered, or several bench-top machines, where previously ORe medium sized mass spectrometer would have been purchased. This has led to a distribution of resources, talent and ability within the industry leading to a more rapid and better response to problems, whether pollution incidents, increased routine analyses or method developments - all this with less staff.
The fact that they are bench-top instruments, requiring little more room than two gas chromatographs, has meant that there are no special requirements regarding their siting, e.g. no special flooring, no air conditioning. Only a 240V power supply and a helium gas supply are required.
The versatility of the lTD and MSD will bring about, as methods are developed, an overall reduction in the total number of Ge's needed within a laboratory and therefore a saving in time, money and space. They also allow the larger nBSS spectrometers that may be present in a group to be used more gainfully, e.g. accurate mass measurerrent and linked scans.
The combination of a 100 position autosampler, MSD and sequencing software has led to a complete change in working practices. If not involved in instrumental method development, the analyst now spends most of the day extracting and preparing samples for overnight or weekend running. This advantage is of course present with any nooern fully automated Ge system but -the bench-top machine is not dependant upon solvent type (the most suitable extraction solvent is no longer exchanged for one more suitable for the detector), nor does it need a range of different detectors to analyse all classes of compounds. The same sample vial can be run under 24 different methods (if necessary),
e.g. the same hexane extract could be sequenced £0>"'
-4-
(i) (ii) (iii) (iv) (v)
This chc'C~ks at run.
organochlorine pesticides P.C.B. 's phosphorus pesticides triazines geosmin and 2-methylisoborneol
same methodology also incorporates sensitivity and performance the start and end of the sequence and calibrations within the
Many of the present analytical procedures use external standards but these do not correct for variation in extraction efficiency caused by matrix effects or errors caused by instrument bias and variations in volun~s. The use of close isomers or hornologues as internal standards go some way to overcome these problenls but a better alternative is the use of isotopically labelled compounds which considerable improve the precision and accuracy of anlysis. As mentioned previously the bench-top instrument is especially suitable in its ability to use deuterated and carbon-13 labelled comp:mnds and many of the pesticides now in use are commercially available as labelled species. Complete chromatographic separation is no longer a necessity, although this is no excuse for poor chromatography. In the case of surface and finished waters, this specificity has enabled many of the clean-up procedures, which are often the most t~ consuming steps, to be eliminated. This advantage is partially offset by the possible reduction in the column lifetime.
Of particular importance with both the lTD and MSD has been the ease of use that has been encountered by all analysts who have used the data syst~s. The staff have been actively encouraged to use the instruments and develop their own methods with the minimum of supervision. This has been facilitated by the use of simple softkey operations, which can be edited to suit particular requir~nts, and the ability to build more complex routines using an easy to follow command language. These methods are absolutely crucial in automatically acquiring, editing, measuring and reporting of the data that has accumulated after a weekend run of 30-50 complex samples. Of course once these methods have been developed they can be copied onto ren10vable discs and distributed to other laboratories.
Both the SySten1S have access to the EPA-NIH 42,000 compound library and any user created library. Both the lTD and MSD can automatically search through a long run and produce reasonable fits, if the original compound is present in the database accessed. The lTD when coupled to an IBM-AT computer is particularly impressive in its speed when library searching, putting many larger much lOClre expensive data SySten1S to shame.
Selective ion lOClnitoring can be achieved with less than 10 pg of component but the lTD, because of its revolutionary design, has been able to record adequate (recognisable) mass spectra with less than 100 pg of an average compound in full scan (50-450 daltons) node. In certain circumstances the lTD full scan data is better than MID data, although this might not be the case if the spectral background levels are high and care must be taken if matrix effects are a possible problem.
4. DISADVANTAGES Even though the bench-top instruments are sensitive they are not as
sensitive for certain compounds as an ECD and this has meant that in many cases the sample volume has had to be increased and the final volumes decreased.
-5-
e.g. for organo chlorine pesticides, sample volume increased from 1 to 2 litres, final extract volume decreased from lml to 0.2ml
High investment in time is initially required to train staff and develop the n~thods. It is absolutely essential to use the correct ions for quantification and qualification. The ratio of the qualifying and quantifying ions must lie within certain limits before the compound is reported as present. These ions must be interference free and this can only be checked by running the n~thod with as many different sample types as possible. Before the method is accepted it is often necessary to run comparisons between the existing GC and the new MS methods.
For good peak area measurements at least ten determinations must be acquired across the- sample peak. Modern capillary columns often have peaks only 3-5 seconds wide and with dwell times of 5Oms, peak switching can really only accarnnndate about six different ions at anyone time. The MSD is capable of ten groups of twenty ions, thus in a complex analysis it may be necessary to run the sample more than once, incorporating different ions in each run.
A further consideration is that the laboratory can become totally dependant on just one instrument to do the majority of its organic analysis. Therefore it is essential that the equipn~nt is very reliable and is regular I y serviced and maintained.
5. RELIABILITY: USER'S EXPERIENCE Both the MSD and lTD have similar service costs and maintenance
agreements. Currently the cost is approximately £3000 per annum.
5.1 WRc PROCESSES It has been found that the lTD is not stable enough in terms of
same day reproducibility to operate with external standard techniques although it has been shown to be more than adequate when internal standards are employed. Possible reasons for this instability are:
(i)
(ii)
(iii)
Zero adustment controlling acceptable background noise needs constant attention from run to run - maybe connected with lack of air conditioning although the condition is apparent even when day-time temperature remains constant.
There are a large number of parameters which can be used to optimise instrument performance, i.e. emission current, segment tuning values, B values, scan speed and zero setting - small electronic instabilities may therefore accumulate.
The instrument is particularly sensitive to water and even small leaks can badly affect all aspects of performance.
The instrument has been out of action for 4.6% of the operating time. Problems have been associated with burnt out transfer line heaters (twice), electron multiplier replacements (twice), shorting out of high voltage connections to source housing, and air leaks (many).
5.2 WRc ENVIRONMENT Many problems were encountered during the first year of operation,
when the percentage downtime on the lTD was about 50%, but recently, following a cCft'Plete overhaul by.Finnigan (necessitated. by a rather catastrophic fault on the transfer line) it has been much more reliable.
-6-
The major problems were short lifetimes of electron multipliers, the turbo molecular pump and transfer line. The ceramic insulators, which isolate the various components in the ion trap itself, also required regular cleaning. While this is a relatively straightforward operation, exposure of the electron multipliers which Finnigan initially fitted, to air (an unavoidable consequence when the trap is dismantled) contributed to their short lifetime. Finnigan now use a multilplier from a different manufacturer, which seems to be Lmaffected by exposure to air, and this is no longer a problau.
5.3 NORTH WEST WATER AUTHORI'ry In the first year the ITO has been down for 8% of the operating
time, many of the problems mirroring those encountered by the WRc laboratories, but following the identification of ceramics as the cause of RF leakage and the change in multiplier manufacturer it has worked well.
It is felt that many of the problaus encountered with the 11TI were those associated with a new design of instrument. These teething problaus have been rectified by Finnigan and it is now proving to be very reliable - a fact supported by the accounts from new users.
5.4 THAME~ WATER AUTHORITY Downtime on the MSO instrument has been 5%, most of this was
encountered in the first 6 months of operation. The majority of this tin~ has involved not the MSO but the data system with the graphics card, monitor and winchester disc drive having to be replaced. The last item being particularly salutary and all important information is now backed-up immediately. It would sean that we received a rogue data system as the expected failure rate is once in seven years. It should be stated that callout response was excellent (same day). The turbo molecular pump has been replaced under contract.
Several software bugs reporting non existant hardware faults caused a few days of wasted effort in board replacement as did incorrect internal plumbing of the open split interface.
Source cleaning is simple and easy, initial performance checks can be achieved within 15 minutes of startup, although it needs overnight pumping to remove the last traces cf water vapour.
6. THE FUTURE The lTD is capable of MS-MS and chemical ionisation. The
awlication of both these techniques should increase the scope of quantitative mass spectral analysis by providing spectra containing single isolated ions more specific to the determinand - the phenol method has already shown that less fragmentation neans a lower limit of detection - and allowing the use of more easily synthesised and cheaper deuterated and el3 containing internal standards with fewer labelled atoms incorporated.
The MSO has been coupled in tandom with an infra-red detector to give GC-IR-MS which will prove extremely useful in solving unknowns or monitoring specific compound classes. The MSO can also be converted for use as a thenrospray HPLC-MS system; a factor of some i.np:>rtance as many of the organics now needing analysis are far more amenable to HPLC separation.
Foreground-background data system operation would be a very useful option to have but as yet none is available although the lTD IBM computers have been networked.
-7-
Preparations are now in hand, co-ordinated by the UK Water Industry Mass Spectrometer Users Group, for the creation and transfer (in IBM PC format) of a computerised data base containing both known and unknown spectra. Thus the experience of long term MS users will be rapidly disseminated amongst the growing number of lTD and MSD owners.
Because of the success and 'low' cost of these instruments and the more stringent requirements being demanded by legislation, the future seems bright for bench-top mass spectrometry. More and more of the water industry laboratories will consider GC-MS as a routine quantitative technique and not something limited to solving the more difficult or sometimes impossible problems.
REFERENCES
(1) Stripping of trace organic substances from water: Equipment and Procedure K. GROB and F. ZURCHER Journal of Chromatography 117, 1976, p285-294
(2) Advances in the identification and analysis of organic pollutants in water: Volume 2 Ed. by L. H. Keith Published by Ann Arbor Science CHAPTER 38: Development of a closed-loop stripping technique for the analysis of taste-and odour-causing substances in drinking water. S. W. KRASNER, C. J. HWANG and M. J. McGUIRE p. 689-710.
(3) Gas chromatographic deternrination of phenol compounds in water as their pentafluorobenzoyl derivatives L. RENBERG Chemisphere 10, 1981, p767-773.
-8-
EXTR
ACT
ION
TEC
HN
IQU
E IN
TER
NA
L ST
AN
DA
RD
STA
ND
ARD
RE
AL
TYPE
O
F SO
LUTI
ON
SA
MPL
E H
ATE
R W
RC
PRO
CES
SES
(pg
) (n
g/l
)
GEO
SMIN
CL
SA
l-CH
LO
RO
-n-A
LK
AN
ES
25
5 F
inis
hed
2-
MET
HY
LISO
BO
RN
EOL
20
4
0' -
HCH
Liq
uid
/
Liq
uid
(P
ENTA
NE)
20
10
0 T
rad
e eff
luen
t M
ETHY
L 4-
TOLU
ENES
ULP
HO
NA
TE
Liq
uid
/
Liq
uid
(D
CM
) M
ETH
YL
BEN
ZEN
ESU
LPH
ON
ATE
10
5
Sew
age
eff
luen
t
THAM
ES
WAT
ER
I G
EOSM
IN
CLSA
l-C
HL
OR
O-n
-AL
KA
NE
S 20
2
-10
0
Riv
er
&
'-0
2-
MET
HY
LISO
BO
RN
EOL
20
5-1
00
F
inis
hed
I
TECN
AZE
NE
20
1-5
0
0(
, ~
, "0
, ~
HCH
20
1-5
0
HEX
ACH
LORO
BEN
ZEN
E 20
1
-50
H
EPTA
CHLO
R 20
1
-50
A
LDR
IN
20
1-5
0
HEP
TACH
LOR
EPO
XID
E 20
1
-50
R
iver
CH
LORD
AN
E L
iqu
id
/ L
iqu
id
(HEX
AN
E)
DEU
TERA
TED
?r
-HC
H
20
5-5
0
&
DIE
LDR
IN
20
2-5
0
Fin
ish
ed
EN
DR
IN
20
1-5
0
pp-D
DE
20
1
-50
p
p-
& o
p-T
DE
20
1
-50
p
p-
& o
p-D
DT
20
1
-50
M
ETH
OX
YCH
LOR
20
2-5
0
MA
LATH
ION
20
1
-50
PA
RA
THIO
N
20
1-5
0
EXTR
ACT
ION
TE
CHN
IQU
E IN
TER
NA
L ST
AN
DA
RD
STA
ND
ARD
RE
AL
TYPE
O
F SO
LUTI
ON
SA
MPL
E W
ATER
TH
AMES
W
ATER
(p
g)
(ng
/l)
TRIB
UTY
L T
IN
SOD
IUM
BO
ROH
YD
RID
E &
TR
IPR
OPY
L T
IN
40
5-2
50
R
iver
L
iqu
id
/ L
iqu
id
(DCM
)
PHEN
OL
20
20
0-1
00
0
2-
&
4-C
HLO
RO
PHEN
OL
5 5
0-1
00
0
DIM
ETH
YLP
HEN
OL
PEN
TAFL
UO
ROBE
NZO
YL
CHLO
RID
E 2,
6-D
IFL
UO
RO
PHE
NO
L
5 5
0-1
00
0
Riv
er
o,m
,p-C
RE
SOL
&
L
iqu
id
/ L
iqu
id
(ISO
-OC
TA
NE
) 5
50
-10
00
&
CH
LORO
MET
HY
LPH
ENO
L 5
50
-10
00
F
inis
hed
-D
ICH
LORO
PHEN
OL
5 5
0-1
00
0
0 2
,4,6
-&
2,4
,5-T
RIC
HL
OR
OPH
EN
OL
5
50
-10
00
I
PEN
TACH
LORO
PHEN
OL
10
10
0-1
00
0
BEN
ZEN
E 10
00
10
00
-20
,00
0
HEP
TAN
E 10
00
10
00
-20
,00
0
TOLU
ENE
1000
1
00
0-2
0,0
00
O
CTA
NE
Liq
uid
/
Liq
uid
(P
ENTA
NE)
D
EUTE
RATE
D
OCT
AN
E ,
1000
1
00
0-2
0,0
00
ET
HY
L BE
NZE
NE
DEU
TERA
TED
N
APH
THA
LEN
E ,
1000
1
00
0-2
0,0
00
R
iver
IS
O-P
RO
PYL
BEN
ZEN
E D
EUTE
RATE
D
PHEN
AN
THRE
NE
1000
1
00
0-2
0,0
00
&
DE
CANE
10
00
10
00
-20
,00
0 F
inis
hed
NA~HTHALENE
1000
1
00
0-2
0,0
00
TE
TRA
DEC
AN
E 10
00
10
00
-20
,00
0
PHEN
AN
THRE
NE
1000
1
00
0-2
0,0
00
.... ....
lTD
(W
RC
PRO
CE
SSE
S)
Geo
smin
cali
bra
tio
n u
sin
g d
eu
tera
ted
g
eosm
in a
s in
tern
al
stan
dard
.
Cal
ibra
tion
Plot
(In
t St
ds)
Filen
aMe:
GaM
Cor
rela
tion
Coe
ff: 9
,997
ge
osM
in
COMP
ound
: 4
of 4
St
anda
rd D
evia
tion:
9,g
e9
<Area
of
SaM
ple/A
rea
of S
tand
ard)
us
(AM
ount
of S
aMPle
Inj
ecte
d)
(Lin
ILin
)
e,se
e -H
IH r
atio
112
/114
9,Sg
e
9,4e
9
e,3e
e
e,29a
9,20
0 9,
499
a,se
0 0,
800
1,00
0 ug
ll
Ion 185.88 .- . f", 1It [lATA,J'B12RalA . D
1.8OO 5 i I 5
6
J :::::1 ' \V!! 'f,
: ::: j ".:~~J~lli ~lJ",UI~ill il..",,,.,,,N l Ul''''B''''''=:d~'''''''b!2=':''''~':::::::;::::::~2:::2====::~~~~2~'' Ion 19'5.00 .. av.
9~1ZI £ 41 is B&2fE:~-::
7 ·~e:~1 • G .. 0e::4~ . , i I
5&r2:E4~
4_l£)E: ~"~
3 . e e::4 ~
frollt DATAI.TB12A08A.D
8
."0111 CRTFhJ'B12A20R.D
is 1 I tOI'l 195. Be a",... .
1 • se:s j 1 .4 E: 5 "t
1.2 r;::!S~ 1.eJE:B-::
e.~ e:: ~ ~
s.eE:4 -j
: : :::~l ___ 't J U..J.!LUl\.,.' ~~&",I" ~1 :~'~11~5 :::;::1 B l~~",:::; .. ~~",:;;.J=_=-_=_:;: .. Tt ftlotl (",In.)
MSD (Thames Water)
Examples of the determination of phenols by pentafluorobenzoylation
Top : 1000 ng/l Middle : 50 ng/l Bottom : Unknown
1 5 8
Phenol , 2 = Cresols , Chloromethylphenol , 6 Pentachlorophenol , is
3 = Chlorophenol 4 = Dimethylphenol Dichlorophenol , 7 = Trichlorophenol Internal Standard (Difluorophenol)
- 12-
~ (
o Q. R
• It:
9 (
o Q. M
• It:
II (
o Q.
• ~ It:
2S
20
1 !5
'0
gamma HCH
2 . 8
1 • S
1 ...
, .2
1 • 0
0. 0~~--------------------------------------------------~ 0.0 2 .9
Arnt Ra.-t;.10
pp - DD"T"
2.'"
MSD (Thames Water)
Examples of the calibrations obtained by spiking two litres of laboratory tap water with 1 , 2 , 5 , 10 , 20 and 50 ng/l standards. Deuterated O-HCH was used as an internal standard.
-13 -
SUMMARY
COUPLING MICR()'LC CAPILLARY GC AS A TOOL IN ENVIRONMENTAL ANALYSIS.
p. PUQUIT and c. DEWAELE··
• RSL-A1hech Europe, Begoniastraat 5, 8-9731 Eke, Belgium .. Laboratory for Organic Chemistry, State University of Gent,
Krijgslaan 281 (S4) 8-9000 Gent, Belgium
A muhidimensional chromatography system with a packed fused silica Micro-LC column connected on-line with a capillary GC column is described. •
The Micro-LC column can act as a sample clean-up orland a chemical class separation column prior to introduction of the sample into the capillary GC column. This technique was used for analyzing complex mixtures and/or trace compounds in complex matrixes.
1 INTROPUCTION
The aim of any form of chromatography is to separate compounds of interest in the shortest time possible. Therefore research has been focussed on the development of very selective and highly efficient LC and CGC columns. Notwithstanding important improvements in column technology, the performance of orie single column is quite often not enough to separate complex mixtures. Better resolution of complex mixtures can be obtained by multidimensional column chromatography (1). In this approach, fractions of one chromatographic column are transferred to a secondary column, with different selectivity, for further separation.
An attractive multidimensional chromatography system is obtained by coupling a LC and a CGC system. The LC column can act as a preseparation column or as a highly efficient and selective sample clean-up column prior to CGC analysis. This coupling can be carried out off-line (2-4) or on-line (5-6). On-line coupling has to be preferred because it provides better quantitation (no loss of sample during transfer) and it allows easy automatation.
The use of conventional. size LC columns, however, has the disadvantage that only a small fraction of the eluted peak can be introduced into the capillary GC column. Even with microbore LC columns (internal diameter of 1 mm) the peak volume that has to be transferred to GC is in the range of 10 to 100 Jll, thus largely exceeding the injection volumes in capillary GC. Grob et al. used an oncolumn concentration technique with very long retention gaps (30-50 m) before the column (7). Such retention gaps allow to transfer up to 1 ml of the eluent, but have the drawback that the solvent evaporation takes a long time (up to 1 hour)(8). In order to reduce the peak volume, Raglione and Hartwich (9) have developed a bundled capillary stream splitter which is placed between the LC detector and the GC.
The introduction of packed fused silica capillary LC columns opens new perspectives for on-line LC-GC coupling (10-12). Due to the small internal column volume these columns allow entire peak volumes to be introduced into the capillary gas chromatograph. This paper describes an on-line multidimensional system coupling packed fused silica Micro-LC columns to capillary GC columns.
-14-
2 ADVANTAGES OF MICBO-LC COLUMNS
The heart of the LC system is a packed fused silica Micro-LC column. These columns are made from capillary fused silica tubing (0.32 mm 10) and are slurry packed under high pressure. The capillary Micro-LC columns offer the same efficiency and the same column stability as conventional columns (13). The advantages of packed fused silica Micro-LC columns are :
- better permeability : under comparable conditions, packed fused silica Micro-LC columns have greater permeability than conventional LC columns (14-15). In consequence smaller particles can be used in Micro-LC columns than in conventional columns or the column length with equal particle size can be longer (16-17).
- high mass sensitivitY : the use of Micro-LC columns in combination with concentration sensitive detectors (such as UV-, fluorescence-, refractive index- and electrochemical detector) results in an increased sensitivity. The sensitivity for concentration sensitive detectors depends on the concentration at the peak maximum ( Cmax ) and is given in equation 1.
Vinj x Cinj Cmax = --------------------------- (equation 1 )
eAs(1 +k)'J2lt"lL.H
where Vinj: injected volume Cinj : injected concentration e : column porosity As : surface area of the column k : capacity ratio L : column length H : plate height
The sensitivity is increased with a factor 206 when switching from a conventional (4.6 mm 10) to a MicroLC column (0.32 mm 10), provided that the same mass of sample is injected. This is illustrated in Figure 1 where the same amount of a test mixture is injected on a conventional and a Micro-LC column. The concentration of the peaks eluted from the conventional column is too low to be detected. Verzele reported a detection limit of 0.7 picogram for pyrene with Micro-LC fluorescence detection (16). The gain in mass sensitivity is most important when only small sample amounts are available. This is the case with many pharmaceutical, biological and environmental samples.
A B
Figure 1 : Difference in mass sensitivity between a conventional and a Micro-LC column. A. 150mm x 0.32mm fused silica column packed with 5 J.llTl BoSiL C 18; mobile phase 75/25
CH3CNlH20; flow rate 3JlVmin; detection UV at 254nm; sample polyaromatic hydrocarbons B. 150mm x 4.6mm column packed with 5 Jlm BoSiL C18; mobile phase 75/25 CH3CNlH20;
flow rate 1 mVmin; detection UV at 254nm; same sample
- 15-
- new detection prospects: the smaller flowrate of Micro-LC columns allows the use of flame and plasma based detectors(18). Laser based detectors are also very interesting as Micro-LC detector and they show detection limtts in the low ferntogram range (19-20). Micro-LC columns can also be used in combination wtth chemiluminescence detection (21).
- chemical inertness : the capillary Micro-LC column is totally metal free which resuHs in an improved recovery and peak shape for metal sensitive compounds.
- better guantitation : the accuracy and reproducibiltty of quantttative resuHs on Micro-LC columns are better due to the improved peak symmetry and the chemical inertness (16).
- low solyent consumption : the optimum flow rate through a Micro-LC column of 320 J.lITlID is about 3 to 5111 per minute. This allows the use of "exotic" (read expensive) mobile phases.
- easy coupling LC-CGC and LC-MS : Micro-LC columns are very suitable for direct on-line LC-CGC due to the small internal column volume. The typical peak volumes, from 1 to Sill, allow transfer of the entire peak into the capillary GC system (see section 3) . Coupling Micro-LC-MS also is more easy. Direct introduction of the column effluent in the source of the mass spectrometer is possible but an interface which controls the evaporation of the solvent is preferred. Figure 2 shows the Micro-LC-MS analysis of PCB's. Micro-LC-FAB-MS (continuous flow fast atom bombardement) can give very good resuHs as discussed recently (22).
100 10
Figure 2: Direct coupling Micro-LC-MS.
200 20
300 Scan
150mm x 0.32mm fused silica column packed with 5 J.lITl RoSiL C18; mobile phase 95/5 CH30H/H20; flow rate 1.8I1Vmin; mass spectrometer Finnigan 4000; sample Arochlor mixture
3. COUPLING MICRO-LC AND CAPILLARY GC (MICRO-LC-CGQ)
3 1 INSTRUMENWION
A block diagram of the on-line Micro-LC-CGC system is shown in Figure 3. There are three important parts in the system namely the Micro-LC eqUipment, the interface and the capillary GC system.
Because of the extremely small volumes, Micro-LC columns must be used with appropriate instrumentation.
- 16-
2
9 11
3
13 ,
8 I
[TId '~
10 1- JJL ____________ 1
15
Figure 3: Block diagram of the on-line micro LC-CGC system: 1) pump, 2) three-way splitter device, 3) analytical HPLC column, 4) waste, 5) injection valve, 6) packed micro LC column, 7) detector, 8) recorder, 9) fused silica connection with 50 1JITl1.0., 10) GC, 11) on-column injector, 12) retention gap, 13) low dead vlume connector, 14) analytical capillary column and 15) recorder and data acquisition system.
- J;l!J.!III2 : as already mentioned, flow rates for Micro-LC columns are in the range of 1 to 10 III per minute. These flow rates can be delivered in two different ways : direct constant-flow and split-flow technique.
The flow has to be precise, accurate and pulseless. Reciprocating pumps are not very appropriate for these low flow rates. Microliter syringe pumps (Brownlee, Carlo Erba, Isco ... ) can generate these low flow rates. An easy alternative to generate the necessary low flow rates is the use of a conventional HPLC pump with the split flow technique(23). This system uses a zero dead volume tee installed between the pump and the injector. One leg of the tee is connected to the injection valve, the second to the restrictor and the third to the pump. The connection tube between the tee and the injection valve must have the lowest dead volume possible, in order to avoid a delay in gradient elution. The restrictor can be a conventional column. The total flow delivered from the pump is split between the Micro-LC column and the restrictor column according to the ratio of the resistance in both COlumns.
-~ : due to the size of a Micro-LC column, injection volumes have to. be very small. With an internal loop valve (Valco) sample volumes from 0.06 to 1 III can be injected precisely and with minimum variance. The Micro-LC column is directly installed into the valve body to minimize extra column effects.
-~ : UV- and fluorescence detection are the most popular LC detectors. In order to meet the high requirements of Micro-LC the detector cells have to be modified. Modification of a fluorescence detector can just be done by inserting fused silica tubing (100, 250 or 320 Ilm 10) into the original cell. The modification of a UV-detector cell is more difficult (24), but cassette-like micro-cells for several UVdetectors are commercialy available (RSL-Alltech Europe).
The connection between Micro-LC and capillary GC system can be carried out in two ways: via an on-column injector or via a switching valve. The procedure of the peak transfer will be discussed in detail in the next paragraph.
The capillary GC system basically consists of a capillary column of which the stationary phase is immobilized in order to prevent phase stripping. A retention gap is installed before the capillary column to handle the large solvent volumes.
The retention gap, which is a piece of uncoated, deactivated fused silica tubing, accepts the band broadening caused by the injection but reconcentrates the solute bands at the beginning of the coated capillary column (25). If a 2m X 0.32mm precolumn (retention gap) is used, up to 30 III solvent can be introduced without sample liquid running into the separation coumn. The system can be used in combination with any GC detector (FlO, ECO, NPO, MS ... ).
-17-
3 2. PEAK TRANSFER
The interface between the Micro-LC and capillary GC system can be either an on-column injector or a switching valve. The latter is preferred because the injection procedure is easier and allows automation. The following procedure is adapted for the transfer via an on-column injector. When the desired LC fraction appears on the recorder, the LC injection valve is placed in the middle between the fill and inject position (stop-flow). By doing so, the split flow system stays under pressure and the solvent flow does not pass through the column. A delay between the appearance of the peak and the switch to stop-flOW has been taken into account to correct the dead volume of the detector cell and the connection tubing. The Micro-LC detector outlet, made of a 0.5 m x 50 j.UTl fused silica capillary, is inserted into the capillary GC column via an on-column injector. Peak transfer is started by placing the LC valve back in the injection position. After complete peak transfer, the LC injector is switched to stopflow position again, the fused silica connection tubing is drawn back, the valve of the on-column injector is closed and the GC temperature programming is started. This transfer procedure was evaluated with a test probe which consisted of polycyclic aromatic hydrocarbons. The relative standard deviation was between 3% and 7% on the measured peak areas (12).
The second way to carry out the Micro-LC-CGC connection is via a four port switching valve which is placed directly after the LC detector. This valve is mounted outside the GC oven. In the normal position the LC effluent goes from the Micro-LC column to waste while the carrier gas enters the capillary GC column. In transfer positions, the LC effluent is directed into the capillary GC column. If the desired LC fraction is transfered into the capillary GC column, the switching valve is put back in the begin position and the CGC analysis is started. This interface allows reproducible peak transfers without loss in resolution. An example of polyaromatic hydrocarbon analysis is shown in figure 4.
4 2
4MIN
40'C IOO'C 200' C 12MIN
figure 4: Micro-LC chromatogram of polycyclic aromatic hydrocarbons.
Column: 150 x 0.32 mm packed with 5 j.UTl RoSiL-PONA; mobile phase: hexane - methylene chloride 75125; flow rate: 5 ~min; detection UV at 254nm. chromatogram of the first transfer from the Micro-LC system. Peak transfer is made by a 4 port Valco switching valve. Column 25m x 0.25mm RSL-300 (df=0.2j.UTl); retentiongap: 5m x O.25mm; tefll)8rature program: 35°C-70°C/10°C min ·11 260°C. Peaks: 1 .anthracene 2.triphenylene 3.benzothiophene 4.dibenzothiophene 5.carbazole
-18-
3.3 INFLUENCE OF THE SOLVENT
At present, most of the reported applications on coupled LC-CGC, are obtained in the straight phase LC mode. Only Cortes reported an example in reversed phase HPLC mode (10). The solvent, however, was pure acetonitrile. Water is a very difficult solvent in GC especially because it forms large volumes of vapor per unit volume of liquid (26) . On the other hand direct injection of 1 to 2 J.1I water is known in CGC for instance for the analysis of volatile halocarbons in drinking water (27). Water also does not damage silicone and polyethylene glycol stationary phases. Recently, Takayama e.a. reported that for alkanes and alcohols in methanol-water mixtures no peak distortion was observed (28). Peaks begin only to distort when large volumes are injected. It is obvious that Micro-LC columns due to their small internal volume are more suitable for reversed phase LC-CGC coupling. More specific research in this area is required. In any case, reversed phase Micro-LC-CGC coupling is restricted to eluents free of involatile additives such as buffer salts or ion pair reagents.
3 4 APPLICATIONS
The proposed multidimensional chromatographic system is very suitable for the analysis of complex mixtures. An example for this is the analysis of a light gasoline fraction (12). First the gasoline fraction is preseparated into different classes on a Micro-LC column packed with 5 ~ RoSiL-PONA, a spherical high surface silica gel. Afterwards, each group (parafins, olefins, naphthenes and aromatics) is separately transfered to the capillary GC system for further analysis. By doing so, less complicated CGC chromatograms are obtained for the different classes.
A second example is the analysis of additives in PVC. The standard procedure for this analysis is a time consuming solvent extraction followed by GC analysis. The additives are separated from the polymer matrix by size exclusion. A Micro-LC column packed with 5 11m Rogel (a polystyrenedivinylbenzene packing material of Alltech-RSL with a pore diameter of 7 nm) was used. The total analysis of phthalates in PVC was carried out in less than 15 minutes (Figure 5). The above is an example of coupling Micro-SEC to CGC. This combination has great general potential.
• 8
2 3
I I o 5 10min
Figure 5: Micro-LC-CGC analysis of additives in PVC
A. Micro-LC system
I
!: il Ii
~t
A
"---. I I o 6 """
Column: 250 x 0.32mm packed with 511m RoGel ( pore diameter 7nm); mobile phase tetrahydrofuran; flow rate 3 J.1IImin; detection UVat 220nm
B. CGG system Peak transfer is made by a 4 port Valco switching valve. Column 10m x 0.10mm RSL-150 (df 0.2~); retention gap: 2m x 0.1 Omm; temperature program: 60°C/1 DOC min -11 26DoC. Peaks: 1. dimethylphthalate 2. diethy~hthalate 3. dioctylphthalate
- 19-
r\ I \ I \
0 5 15 25 tR(mln ) -
,..,
r
I ;! .. .. .,.
~ '?' c 0 C-ol <I> L
U w A
45 50
tR(min)_
Figure 6: Analysis of a sediment extract from the river Meuse.
(A) LC-GC-ECD chromatogram after soxhlet extraction and desulphuration. During the chromatogram the attenuation was changed. (B) LC-ECD qhromatogram of the same sample used in Figure 6A (Reproduced with permission from ref. 34.)
Coupling LC-GC can also be used for the analysis of minor compounds in complex mixtures. Gianeselio et al reported the analysis of broxaterol in human plasma and urine by coupled LC-GC (29). By "picking" the traces of interest out from the complex matrix the detection limits are seriously decreased. Another example is the analysis of heroin metabol~es in urine (30).
Coupling Micro-LC-CGC is also very useful in environmental analysis. Cortes et al described the analysis of polychlorinated biphenyls (PCBs) in coal tar (10). The on-line multidimensional system reduces not only sample preparation time but the system is also superior to either LC or GC alone. Polychlorinated biphenyls were also isolated from fish using coupled LC-GC (31).
The determination of polychlorinated biphenyls (PCBs) in sediment by CGC, requires a rather laborious and time-<:onsuming pre-treatment (32-33). Two separate adsorption columns have to be used and intermediate concentration, drying, etc. steps introduce a high risk of contamination and loss of analytes. Figure 6A shows the analysis of a sediment sample from the river Meuse by on-line coupling LC-CGC (34). The LC-ECD chromatogram is shown in Figure 6B, the section which is transfered to the capillary column is indicated by dotted lines. The multidimensional system makes sample clean-up easy and fast.
- 20-
4 CONCLUSIONS
The proposed multidimensional system has been developed by coupling of a packed fused silica Micro-LC column with a capillary GC column. The typical low flow rates of the Micro-LC system allow loading of the entire LC peak volume into the capillary GC column. The use of a retention gap allows the injection of several microliters without loss in resolution or bad peak shape of the components of interest. The Micro-LC system is very selective and can be used in the straight phase, reversed phase or size exclusion mode. The Micro-LC column can act as a concentration column, a sample clean-up column or as a group separation column prior to CGC analysis. For many applications, sample preparation times are reduced and detection limits are decreased. The proposed technique is a promising tool in the analysis of complex mixtures.
ACKNOWLEpGEMENT
We thank Prof. Verzele for the stimulating discussions. This research is supported by the IWONL (Instituut ter Aanmoediging van Wetenschappelijk Onderzoek in Landbouw en Nijverheicl).
REFERENCES
(1) R. Majors, J. Chromatog. Sci., 18 (1980) 571 (2) F. Desanzo, P. Uden and S. Sigga, Anal Chem. 52 (1980) 906. (3) P. Williams, K. Bartle, G. Andrews ans D. Mills, HRC & CC9 (1986) 39. (4) S. Morimoto, F. Morishita and T. Kojima, J. Chromatog. 368 (1986) 291. (5) B. Majors, Pittsburgh Conference on Analytical and Applied Spectroscopy (1979) paper 116 (6) J. Apffel and H. McNair, J. Chromatog. 297 (1983) 139. (7) K. GrobJr, D. Frolich, B. Schilling, H. Neukomand P. Nageli, J. Chromatog. 295 (1984) 55. (8) F. Munari, A. Trisciani, G. Mapelli, S. Trestianu, K. Grob Jr and J. Colin, HRC & CC 8 (1985) 601. (9) T. Raglione and B. Hartwich, Anal. Chem. 58 (1986) 2680.
(10) H. Cortes, C. Pfeiffer and B. Richter, HRC & CC 8 (1985) 469. (11) H. Cortes, B. Richter, C. PfeHfer and D. Jensen, J. Chromatog. 349 (1985) 55. (12) D. Duquet, C. Dewaele and M. Verzele, Proc. 8th Int. Symp. Capillary Chromatography,
Riva, May 19,1987,1263. (13) M. De Weerdt, C. Dewaele and M. Verzele, submitted to J. Chromatog. (14) M. Novotny in "Microcolumn Separations" Eds. M. Novotny and D. Ishii, Elsevier,
Amsterdam (1984) 20. (15) M. Verzele, M. Deweerdt, C. Dewaele, G. de Jong, N. Lammers and F. Spruit, LC/GC Magazine
(1986)1162 (16) M.Verzele and C. Dewaele, HRC &CC, 10 (1987) 280. (17) A. Hirose, D. Wiesler and M. Novotny, Chromatographia, 18 (1984) 239 (18) S. Folestad, L. Johnson and B. Josefsson, Anal. Chem. 54 (1982) 925 (19) E. Guthrie, J. Jorgenson and P. Dluzneski, J. Chromatog. Sci. 22 (1984) 171 (20) V. McGuffin and R. Zare, Appl. Spectrosc. 39 (1985) 847 (21) G. DeJong, N. Lammers, J. Spruit, C. Dewaele and M. Verzele, Anal. Chem. 59 (1987) 1458 (22) P. Boulanguer, Y. Leroy, G. Ricart, J. Alonso, C. Colbert, D. Duquet and C. Dewaele, submitted
to Carbohydr. Res. (23) SJ. van derWal and F. Yang, HRC &CC,8 (1983) 216. (24) J. Vindevogel, G. Schuddinck, C. Dewaele and M. Verzele, Proc. 8th Int. Symp. Capillary
Chromatography, Riva, May 19,1987,1175. (25) K. Grob Jr, J. Chromatog. 237 (1982) 15. (26) K. Grob and B. Schilling, HRC & CC 8 (1985) 726. (27) K. Grob and G. Grob, HRC& CC 6 (1983) 133. (28) Y. Takayama, E. Sumiya and S. Kawai, HRC & CC, 10 (1987) 201. (29) V. Gianesello, L. Bolzani, E. Brenn and A. Gazzaniga, Proc. 8th Int. Symp. Capillary
Chromatography, Riva, May 19,1987,1216. (30) F. Munari and K. Grob, Proc. 8th Int. Symp. Capillary Chromatography, Riva, May 19, 1987, 1226.
(31) K.Grob,HRC&CC, 10(1987)416. (32) A. Holden and K. Marsden, J. Chromatog., 44 (1969) 481. (33) D. Wells, A. Cowan and A. Christie, J. Chromatog., 328 (1985) 372 (34) F. Maris, E. Noroozian, R. Otten, B. Van Dijck G. De Jong and U. Brinkman, Proc. 8th Int. Symp.
Capillary Chromatography, Riva, May 19,1987,1236.
- 21-
STANDARDIZED ANALYTICAL METHODS FOR EC PRIORITY POLLUTANTS
Summary
S. Schmidt
BUro Dr. Schmidt, H.-T.-v.-Bottinger-Str.8 0-5090 Leverkusen 1
Some examples are given for the progress being made in the analysis of the 129 EC priority pollutants (Easily volatile halogenated hydrocarbons, benzene and derivatives, chloroethene, phenolic compounds, polycyclic aromatic hydrocarbons). The results of interlaboratory trials, as far as available, are stated and interpreted. Preliminary results on test methods are offered. Some reasons for the slow progress is given.
1 Introduction
The following contribution only deals with standardized methods or standards in preparation for the determination of some of the 129 priority pollutants. Standardized methods are always conventions and neither represent the latest investigations nor contain the most advanced technique. The methods must be suitable for routine control purposes and the analytical expenditure must be related to the facilities of laboratories in industries, universities and state authorities. A well trained team of analysts in a well equipped laboratory dealing with analyses on a more or less known matrix will easily deliver "better" results; at least results in a lower application range.
In Germany F.R., the Federal States are in charge of the water monitoring programmes. We would have to face a great number of problems if each of our states would apply different methods for the determination of one parameter or a group of parameters because our rivers often pass several states on their way to the sea.
Besides, as a country in the centre of Europe we have to face serious problems with transfrontier water pollution which can only be solved by joint measurements. Once having agreed on a method, we can more easily carry out interlaboratory trials which enable us to give clear statements on the reliability of the result. In regard to long-term monitoring of water quality, the reliability and the comparability of results is given highest priority.
- 22-
2 STANDARDIZED METHODS
2.1 EASILY VOLATILE HALOGENATED HYDROCARBONS
According to the following method, 22 of the 129 EEC priority pollutants may be analyzed. This includes fluorinated, chlorinated, brominated, iodinated, mainly non-aromatic hydrocarbons having 1 to 6 carbon atoms. Their boiling temperature lies in the range of 20 to 180 0C.
Samples are collected in amber glass bottles with conical shoulder. Prior to use, the bottles are heated for at least 1 h at 1000C in a ventilated oven.
When sampling, the bottles are completely filled, normally by immersion. Losses of volatile halogenated hydrocarbons will occur when composite samples are mixed from individual samples. Therefore, the individual samples are extracted and their extracts are mixed. If necessary, a reducing agent is added. The extraction must be carried out within 48 h after sampling. From the brim-filled and cooled sample water is discarded and the residual volume is determined by weighing. Ice-cold pentane is added.
Instead 0 f pentane, hexane or dime thy Ibenzene may be used. The choice depends on the retention time and the determinand in question. For drinking water and surface water, 2 to 10 ml of extracting agent will be sufficient, in the case of waste water 50 ml will be appropriate. Under ice-cooling, the sample and the pentane are mixed vigorously using a magnetic stirrer. The extracting agent must be finely dispersed in the sample in order to obtain a reproducible recovery. The extract may be removed directly or via a microseparator. An aliquot of the clear sample is injected into the gas chromatograph. Usually, capillary columns will be chosen. At least two separation columns of different stationary phases and polarity are required to ensure quantitative results. An electron capture detector (ECD) is suitable for most purposes. In case of chloroethene, chloroethane, other monochlorohydrocarbons as well as difluorohydrocarbons a flame ionization detector (FID) will be more appropriate.
Special care must be taken in the calibration step. Several approaches are necessary. At first, a calibration of the gas chomatographic system should be made. Subsequently, the extraction procedure may be included in the calibration of the entire procedure. By this stepwise procedure, it is possible to calculate the extraction recovery of the individual substances. The joint calibration corresponds to the evaluation procedure using an external standard. The use of an internal standard is especially recommended if pentane serves as extracting agent. The step-by-step approach is advisable for laboratories lacking sufficient experience wIth the ECD, or in the case where unexplainable results require trouble-shooting. The single-stage calibration over the entire procedure may be applied as a routine calibration prior to or after each series of analyses, or if further
- 23-
substances are included in the work programme. Interferences are likely to occur in the sampling step
by losses of the determinand. Often, interfering peaks may occur at the beginning of the chromatogram. These are often due to a contamination of the sample e.g. as a result of sprays being used or solvents being stored in the laboratory or by refrigerants from chillers.
The phase separation may be disturbed by emulsions being formed or by suspended matter being present in the sample. This may be overcome by the use of a microseparator.
The linear dynamic range of the ECD is limited, strongly depending on the respective compound. Besides halogen compounds, the ECD detects halogenated ethers, ketones, nitro compounds, sulfur compounds such as carbon disulfide, and many others.
Two interlaboratory trials have been carried out with about 18 participating laboratories. The results are given in table 1. Two matrices have been tested, spiked drinking water and waste water.
The results in the interlaboratory trial were evaluated by calibration over the total procedure without internal standard. All participants had received the same calibration stock solution. The statistical recovery rate is the average recovery of the true value from all laboratories.
Except for dichloromethane, the results are fairly satisfying. Only a few number of laboratories have tried to analyze dichloromethane; most of them reported far too high values. The consequence is the statement that this method is not suitable for the determination of dichloromethane in these low concentrations. A separate method using head space gas chromatography will be elaborated.
2.2 BENZENE AND DERIVATIVES
12 Substances out of the 129 list may be analyzed according to this procedure. The standard method in preparation contains both an extraction procedure and a head space method.
Concerning the sampling and the sampling procedure the same demands can be formulated as for the easily volatile halogenated hydrocarbons.
The extraction step is carried out using pentane. Special care must be taken to ensure a sufficient purity of pentane. A distillation over an effective column may help to clean contaminated pentane, but with some batches even a repetition of this time-consuming procedure will not lead to success. The relation of extracting agent to water sample depends on the application range. For a range of 1 to 10 1l9/l, a ratio of organic layer: water sample of 1: 10 is appropriate, for concentrations between 1 and 10 mg/l a ratio of 100:10 is suitable. For gas chromatography 2 columns of different stationary phases and different polarity should be used. The installation of both columns on
- 24-
Easil
y
vo
lati
le
halo
gen
ate
d
hy
dro
carb
on
s
-R
esu
lts
fro
m
an
in
terla
bo
rato
ry
trial
carrie
d
ou
t in
N
ov
em
ber
19
86
Matr
ix
Su
bsta
nce
L N
N
AP
x so
lI
x W
FR
SR
VR
51
V
I
A
CH
C1
3 1
8
70
0
14
,7
16
,1
11
0
3,0
1
8,6
1
,6
9,9
B
18
7
0
0 7
9,2
1
8,6
2
3,6
4
,2
5,3
A
CC
14
18
7
0
0 1
5,9
1
6,8
1
06
4
,8
28
,6
2,0
1
2,0
B
18
7
0
0 7
9,3
7
6,2
9
6
18
,3
23
,9
7,2
9
,4
A
CC
13
CH
3 1
7
63
1
0
13
,1
13
,7
10
5
2,1
1
4,9
1
,0
7,0
B
18
7
0
0 6
5,7
7
1,2
1
08
1
2,3
1
7,2
6
,3
8,9
I tv
A
C
C1
2C
HC
l 1
6
60
1
4
14
,6
15
,2
10
4
2,6
1
7,0
1
,0
6,8
V
>
I
B
17
6
6
6 7
3,0
7
4,7
1
02
1
4,8
1
9,8
7
,3
9,7
A
CC
12
CC
12
17
6
5
7 1
6,1
1
5,9
9
9
3,1
1
9,1
1
,2
7,7
B
18
7
0
0 8
0,5
8
1,3
1
01
1
4,7
1
8,1
6
,4
7,8
A
CH
Br
3 1
7
65
7
14
,3
13
,2
92
2
,2
16
,3
1,2
8
,7
B
CH
2C
l2
9 3
3
13
6
5,8
1
01
1
53
4
6,5
4
6,2
1
1,9
1
1,8
A
Dri
nk
ing
w
ate
r x
mea
n
valu
e
in
\J-g
/l
B
Wast
e
wate
r W
FR
Reco
very
rate
in
0
' ,0
L N
um
ber
o
f la
bo
rato
rie
s
SR
Rep
ro
du
cib
ilit
y
sta
nd
ard
d
ev
iati
on
, in
\J
-g/l
N
N
um
ber
o
f v
alu
es
VR
Rep
ro
du
cib
ilit
y
varia
tio
n
co
effic
ien
t in
0
' ,0
NA
P O
utl
ier
in
0'
51
R
ep
eata
bil
ity
sta
nd
ard
d
ev
iati
on
in
\J
-g/l
,0
x so
lI
tru
e
valu
e
in
\J-g
/l
VI
Rep
eata
bil
ity
v
aria
tio
n
co
effic
ien
t,
in
0'
,0
one injection is highly recommended to allow a simultaneous procedure. For detection, aFlame Ionic Detector (F ID) is suitable. In some cases, a Photo Ionic Detector (PID) or a mass selective detector may be prefereable. Peak overlapping cannot be excluded. An indication for such an overlapping is a deviating result on the two columns.
The same calibration procedure as described for the easily volatile halogenated hydrocarbons can be applied. As internal standard, toluene d8 has been proved suitable.
For· head space chromatography potassium carbonate is added to the sample. The amount depends on the volume of the sample. Some solid should remain in the sample when the operation temperature is reached - a temperature of 80 DC is suitable -. Thus, the fixed ionic strength in the liquid will influence the distribution equivalent in favour of the vapour phase. Always the same amount of salt should be added in order to achieve a constant vapour phase.
It is not settled yet wether the same results will be achieved when applying the extraction procedure or the head space method. First investigations have shown a good c orrel at ion 0 f greate r than 0, ,994 for benzene, toluene, ethyl benzene and o-xylene. But further investigations using waste water samples are necessary before making a final statement.
The interferences with these methods are mostly the same as stated for the easily volatile halogenated hydrocarbons. A special problem arises with benzene because it is spread everywhere. Interferences by suspended matter or emulsifiers will be smaller when applying the head space method instead of the extraction procedure.
The expected lower limit of determination is 0,5 ~g/l . Interlaboratory trials are in preparation. It. may be possible that these low values will not be achievable in the case of waste water analysis.
2.3 CHLOROETHENE
The separate standard on the determination of chloroethene includes head space gas chromatography. Special problems arise during sampling because chloroethene escapes very easily. Preferably, sampling is carried out using head space flasks which are closed with a septum immediately after sampling. In the case of normal sampling using bottles and subsequent refilling processes it is advisable to analyze wich each set of samples the contents of an empty, clean head space flask in order to make sure that chloroethene later analyzed in the water did not derive from the air and had entered the sample during refilling. Samples must be analyzed as quickly as possible, at least within 48 hours.
If sampling was not carried out using a head space flask, the sample must be filled in these using any device to avoid underpressure. The addition of salt is not necessary. The water sample is then brought to 80 D C (at
- 26-
least for 1 h) and the gas chromatography is carried out using the following working conditions: Injector temperature: 150°C Detector temperature: 220°C Column temperature: 50°C, isothermal Helium with a gas flow of 25 ml/min serves as carrier gas. A packed column, Carbopack C, grain size 0,15 to 0.18 mm, loaded with 0,19~~ picric acid, has proved- suitable. A flame ionic detector is used.
The calibration may be carried out either by direct measurement or by a standard addition procedure.
An interlaboratory trial was carried out recently with 18 laboratories participating. The samples were not distributed. Instead, the participating members took the samples themselves at a fixed date from 6 containers. As matrix, distilled water, municipal and industrial waste water were used, spiked with chloroethene in concentrations between 15 and 500 ~g/l. All samples were analyzed the following day in the respective laboratories.
In the case of distilled water spiked with 15 ~g/l chloroethene a recovery rate of 95~~ and a repeatability standard deviation SR of 2.99 ~g/l and a reproducibility standard deviation of 0,70 ~g/l were achieved. Further values canbe seen in table 2. (Table to be included).
2.4 PHENOLIC COMPOUNDS
Various extraction and derivatization procedures have been proposed for the separation of phenolic compounds; their applicability depends on the individual problem. In principal, extraction procedures may be used for all kinds of water, the lower limit of determination, however, will not be as low as for the derivatization procedures.
According to our standard in preparation, the unfiltered sample is extracted with diethylether. The pH of the water sample should be less than 2. If the presence of oxidizing agent is expected, especially in the case of free chlorine, sodium sulfite is added. The acidified water sample is extracted with diethylether, the organic phase is treated several times with sodium hydroxide solution and extracted again. The organic phase is cleaned by running it through a column filled with silica gel. The etheral phase is subseqently concentrated by isothermal distillation at room temperature. Starting with a sample volume of 800 ml and ending at a residual volume of 200 ml, the enrichll'ent factor is 4ooo.It may be increased to approximately 10 by altering the volumes. The extract may then either be analyzed immediately or stored at -2o oe for not more than 1 week.
Gas chromatography is carried out using two capillary columns of di fferent polarity and stationary phases installed in one injector. In most cases a flame ionic detection or in the case of polychlorinated phenols an electron capture detection is used.
A lower limit of determination of 0,05 to 0,1 ~g/l is expected. First interlaboratory trials show the results
- 27-
given in table 3 (to be inserted later). Alternatively phenols are separated by derivatisation
with pentafluorobenzoyl chloride. The method delivers good results provided the procedure is followed very strictly. The acidified water sample is treated with sodium hydroxide in a separating funnel. Hexane and an internal standard, e.g. 2,4-dibromophenol, is added. The organic layer is discarded, the aquous layer treated with sodium hydrogencarbonate, hexane, and pentafluorobenzoy 1 chloride. After shaking for five min the organic layer is once again treated with hydroxide solution, dried over sodium sulfate and concentrated after addition of decane as a keeper. The phenols are analyzed by gas chromatography with an electron capture detector.
Interferences with these methods are likely to occur in presence of surfactants, emulsi fiers and higher amounts of organic solvents which will affect the extraction recovery. In case of a second liquid phase in the water sample caused by organic substances the determination of phenols must be restricted to the aquous phase and the non-aquous phase is to be investigated separately.
2.5 POLYCYCLIC AROMATIC HYDROCARBONS
A thin layer mBthod for the determination of 6 polycyclic aromatic hydrocarbons was standardized in 1981. Although the method is still widely used, especially for moni tor ing 0 f water used for food products, better and quicker result.s were expected from methods using HPLC.
In the me·antime, however, work has started again using thin layer chomatography with plates other than those coated with aluminium oxide and acetylated cellulose. First investigations involve plates coated with silica gel impregnated with various substances. As extracting agents isooctane or a mixture of dimethylsulfoxide/ water/hexane are tested. A draft standard for the 6 PAH is expected in 1988 and further investigations will include a larger number of polycyclic aromatic hydrocarbons.
The HPLC method for the determination of PAH starts with a cyclohexane extraction followed by a precleaning step using silica gel and hexane/dichloromethane (1:1). The gradient elution is carried out with methanol/tetrahyrofurane, temperature 30 0 C, flow 0,7 ml/min. A suitable column is 4 ~m Nucleosil 100-PAH (250x4). A fluorescence detection with a wave length programm allows the determination of 0, 005 ~g/l. Values from interlaboratory trials are not jet available.
3 FINAL REMARKS AND PROSPECTS
Compared with the results presented by Dr Laubereau at the fourth European Symposium on Micropollutants it seems as if only little progress has been made. To some respect unfortunately this is true and there are several reasons for it.
- 28-
The interlaboratory trials - in our opinion an integral part of each standard are very expensive and time consuming. Despite of the difficulties concerning the financial part we had to face a number of more or less severe accidents with surface water pollution during the last year. These took almost all extra capacity reserved in the laboratories for the elaboration of new methods and the participation in interlaboratory trials.
But there is still another reason. The list of the 129 priority pollutants was published in 1982. A lot of criticism turned up afterwards concerning the choice of substances included therein. Additional lists including more substances have been discussed since then. But even if new lists contain 300 or more substances, it must be kept in mind that among the 100.000 so-called "Altstoffe" -chemicals produced and registered- there are 600 organotin compounds, more than 2000 organic fluorocompounds more than 15.000 organochlorinated hydrocarbons. Besides, about 300 biocides are known.
These figures indicate that we will never be able to analyze more than a fractional number.
Of course it is most desirable to have as many single substance methods for the control of surface water especially in order to allow a quick reaction in the case of accidents.
Methods suitable for routine control measurements of water quality are urgently needed and these methods must cover as many substances as possible. Therefore in the last years more and more activity -and financial support- was given to the evaluation of sum parameters like adsorb able organic halogenated hydrocarbons (AoX), purgeable organic halogenated hydrocarbons (POX) adsorbable organic fluorine compounds (AoF) and others.
Of course it is known and accepted that sum parameters have a lot of disadvantages one of the most important that it is not known which substances are measured.
Coming back to the list of 129 priority pollutants I would like to recall that 92 out of the 129 substances are halogenated hydrocarbons. It was checked wether these substances may be determined using the AOX. With the exclusion of benzylidene chloride, chloroacetic acid, chloroethanol and trichlorotrifluorothane satisfying recovery rates were achieved.
Be sid e s , b i 0 ass a y s for to xi cit Y t est i n g r e c e i vern 0 r e and more importance. This includes not only the well-known fish test but also tests using algae, plants, and especially bacteria.
REFERENCES
German standard methods for the examination of water, waste water and sludge; jointly determinable substances (group F); determinationjof easily volatile halogenated hydrocarbons (F4) ; DIN 38 407 F 4
- 29-
--; jointly determinable substances (group F); determination of benzene and homologues (F 5) (standard in preparation)
; single components (group P); determination of chloroethene (P 2) (DIN 38 413 P 2)
--, jointly determinable substances (group F); determination of phenols by gas chromatography (F 10) (standard in preparation)
--, jointly determinable substances (group F); determination of polycyclic aromatic hydrocarbons by thin layer chromatography (F 9) (standard in preparation)
--, jointly determinable substances (group F); determination of polycyclic aromatic hydrocarbons by HPLC (F8) (standard in preparation)
-~-
THE APPLICATION OF BONDED SILICA EXTRACTION COLUMNS IN SAMPLE PREPARATION PRIOR TO THE ANALYSIS OF ORGANIC MICROPOLLUTANTS IN WATER
R.A.CALVERLEY Analytichem International, P.O.Box 234, Cambridge, U.K. CB2 IPE.
Summary
For many applications involving the extraction, concentration and purification of organic micropollutants from water samples, the use of bonded silica extraction columns can be a valuable alternative to traditional sample preparation techniques such as liquid-liquid extraction. The commercial availability of a wide range of bonded silica sorbents means that extraction and clean-up procedures can now be developed utilizing non-polar, polar and ion exchange mechanisms.
1.INTRODUCTION A traditional approach to the preparation of water samples prior to
analysis of organic micropollutants has been to utilize liquid-liquid extraction. For many analysts and applications this technique has proved cumbersome, time consuming and lacking in extraction efficiency or selectivity. These shortcomings have led to the development of column chromatographic techniques to replace the liquid-liquid extraction steps. Extraction chromatography, involving the removal of constituents from a liquid sample as it passes through a column of solid adsorbent, exploits the surface characteristics of a wide variety of materials (1). These include charcoal, polyurethane foam and a variety of micro reticular resins.
The use of both hydrophobic and ion exchange microreticular resins in the extraction of organic pollutants from water has been well documented. One practical problem associated with the use of these resins is that an exhaustive washing of the resin material is often necessary to avoid contamination of the column eluent by resin decompostion products (2) •
During the last decade great improvements have been made in both the methods for manufacturing and the quality control of a wide range of bonded silica sorbents. Depending on the nature of the functional group bonded to the surface of the silica, these materials can exhibit nonpolar, polar or ion exchange characteristics. The commercial availability of these sorbents either as bulk material or as pre-packed extraction columns (fig. 1) e.g. Bond ElutR Extraction Columns (Analytichem Int., Harbor City, CA, U.S.A.) or Sep-PakR Cartridges (Waters Division, Millipore Corp., Milford, MA, U.S.A.) has resulted in their growing use in sample preparation procedures. The use of bonded silica sorbent technology in this· manner can be viewed as a natural evolution of extraction chromatography, now probably better known as so·lid phase extraction (SPE). In this paper the abbreviation SPE refers to the use of bonded silica sorbents in the solid phase extraction columns.
The characteristics of a wide range of commercially available bonded silicas and the nature of the interactions that are possible between the sorbents and organic compounds with which they are capable of interacting are described. Details of the procedure for the preparation
- 31-
interacting are described. Details of the procedure for the preparation of the extraction columns prior to use are given and the need to maintain solvation of the bonded silica sorbent during the extraction of large volumes of water is highlighted. Some examples citing the use of a variety of different sorbents in the extraction, purification and concentration of organic pollutants from water are outlined.
POLYETHYLENE OR STAINLESS STEEL FRITTED DISKS (2O!Un)
Ail.l Example 01 a Commercially Available Ex1ractlon Column
MEDICAL GRADE POLYPROPYLENE SAMPLE RESERVOIR
SORBENTBEO (TYPICALLY 1 OO·1000mg SORBENT)
2. CHARACTERISTICS OF BONDED SILICA SORBENTS General properties. Bonded silica sorbents are prepared by reaction
of the surface hydroxyl groups (silanols) with halo- or alkoxy-silyl derivatives, resulting in the covalent bonding of a wide range of functional groups (table I). The high surface coverage achievable during the bonding process imparts an adsorptive characteristic to the bonded silica that is primarily a function of the phase covalently bonded to the silica surface. Due to steric hindrance during the bonding reaction, some silanol groups will be inaccessible to the bonding reagent. Even after a secondary "end-capping" reaction with an appropriate small silane reagent aimed at reducing residual silanol group concentrations, some unreacted silanol groups will remain on the surface of the silica. These few silanol groups are particularly significant if basic compounds are being extracted. At a pH greater than 2, the silanol groups can ionize to produce a negative charge on the surface of the silica, thereby causing a strong interaction with any molecule carrying a positive charge. This electrostatic interaction is eliminated by the use of pH control or high ionic strength buffers.
The base silicas used in the preparation of bonded silica sorbents for sample preparation generally have a high surface area, typically 350-500 m2/g. This allows maximum sample loading capacity for a given mass of sorbent. The porosity of the silica is often in the region of 60 Angstrom, and this porosity facilitates high capacities for molecules with molecular weights of less than approximately 20,000 because they have access to a large percentage of the silica surface area located within the silica pore structure. Depending on the source of the products, the average particle size of the sorbents is likely to be between 30-125 micron. These particle sizes are significantly larger than
- 32-
Type
of I
nler
acU
ons
AvaU
able
NO
N·P
OU
R
pKa
NON·
POLA
R PO
LAR
ANIO
N E
XCHA
NGE
CAT
ION
EXC
HAN
GE
Cia
Oc
tadec
yt -$
-C1ef
i3J
• 0
......
....
f
C8
Octyf
-~i-CeH17
• 0
()
C2
Ethy
t -$-~Hs
• •
, ... .... '
CH
Cy
c:Ioh
exyI
-~-()
• 0
, ....
..... '
PH
Phen
yt -*-
(6)
• 0
, ....
'", .. '
POLA
R
CN
Cy
aoop
ropy
I -~
i-CH
~H~H~N
• •
, ...... ..... ,
20H
Di
oI -S
i-CHz
CHzC
HzO
CHzC
H-CH
:z
• •
· .
• I
I .... '
OH
OH
SI
S.ilic
a -~
i-OH
•
, ....... · .
~
I NH
2 Am
inopr
opyl
-~i-CH2C~CH2N~
9.8
0 •
• · .
.....
' PS
A EI
h}1e
nedia
mine
N-pr
opyi
-~iC
HzCH
2CHz
NHC~
C~NH
2 10
.1
0 •
• , ....
... 10
.9
· . IO
N E
XCH
ANG
E
SCX
Benz
enes
ulpo
nytp
r -*
CH
zCH
z {O}~ -H
+ •
0 •
PRS ~
-$-CH~~~-~ -H
+ 0
0 •
CB
A
Carb
oxye
lhyl
-~iC
HiCH
zCOO
H 4.
8 0
0 •
I
I
• D
EA
Dielh
ylarn
inopr
opy
-Si~HzN(~H,lz
10.7
0
• f· ..
.
I ' .. '
I +
0
0 •
~ ....
SA
X Tr
imeth
ylami
nopr
opyt
-~i-C~HzN·{~CI-
......
i
.-P
rlm
ary
· O
-Sec
onda
ry
::::-
Sila
nol A
ctiv
ity
Table
L S
truct
ures
and
Pro
pet1l
es o
r Som
e Com
mer
cially
Ava
ilable
BM
led
Silic
a So
rtlen
lS
HPLC column packing materials to ensure an easy flow of the samples through the extraction columns when using a vacuum manifold sample preparation station. The silica particles are sufficiently rigid to withstand high ve+ocity flow of solvents when these sorbents are packed into small extraction columns.
Bonded silica sorbents do not shrink or swell when in contact with aqueous or organic solvents. Normally, bonded silicas are considered stable over a pH range of 2-7.5. However, these sorbents are compatible with samples and solvents outside this pH range if they are passed through the extraction column in a rapid pass manner.
3. SORBENT-ISOLATE INTERACTIONS The selectivity of the extraction procedures carried out using
bonded silica sorbents is a function of both the solvated bonded phase and the solvents used to elute interferences and isolates from the extraction columns. The wide range of sorbents illustrated in table I generally provide sufficient opportunities to optimize an extraction and purification procedure. This range of sorbents can be conveniently divided into three classes according to the nature of the primary interactions that are possible between the isolates and the sorbents i.e. non-polar, polar and ion exchange. A handbook has been published which reviews the characteristics of bonded silica sorbents and describes how to control the available sorbent-isolate interactions for sample preparation procedures (3).
3.1 Non-polar interactions Non-polar interactions are those based on van der Waal's dispersion
forces that occur between the carbonaceous component of the isolate and the sorbent functional group. The principal non-polar sorbents are CIS' CS' C2' Cyclohexyl and Phenyl phases. However, since all sorbent functional groups are bonded to the silica substrate through a carbon chain, all the s~rbents illustrated in table I are capable of exhibiting non-polar interactions to some extent (except silica).
Non-polar sorbents, in particular CIS' have been the first choice for most analysts extracting organic compounds from aqueous samples. The aqueous environment is ideal for this mode of extraction because it maximises the van der Waal interactions between the sorbent and the isolate.
For most compounds the CIS sorbent offers the strongest interactions of all the non-polar sorbents. For some isolates the choice of a CIS sorbent can be a problem since too strong an interaction between the isolate and the CIS sorbent necessitates the use of a large volume of a strong eluting solvent. The solution to this problem is to use a less retentive sorbent such as CS' Phenyl or C2. This will result in a reduction in the volume of eluting solvent required to elute the isolate from the extraction column and a more concentrated isolate solution will be produced.
The elution of isolates from non-polar sorbents is normally achieved using solvents or solvent mixtures with sufficient non-polar character to disrupt the non-polar interactions. For many weakly retained compounds, an aqueous based solvent containing a very small percentage methanol or acetonitrile is capable of disrupting the non-polar interaction. In some applications the elution of non-polar isolates is only possible when relatively non-polar solvents such as chloroform, ethyl acetate or even hexane are used. The choice of eluting solvent is normally governed by the solubility of the isolates in the solvents. The transition from an aqueous sample passing through the extraction column
-~-
to a water immiscible eluting solvent is achieved by drying the column after sample application by pulling air through the column using a vacuum manifold sample processing station or by centrifugation of the extraction column.
Almost all isolates have a potential for retention on a non-polar sorbent. Exceptions include extremely polar molecules, such as carbohydrates, which are highly water soluble.
Most compounds containing ionizable functional groups can be retained on a non-polar sorbent. The control of retention and elution can usually be achieved by optimizing the pH of the sample, the interference and isolate elution solvents.
3.2 Polar interactions Retention of polar compounds by one of the polar sorbents such as
CN, Diol, NH2 and PSA is facilitated by a relatively non-polar environment such as hexane, ethyl acetate, dichloromethane or mixtures of these solvents. Retention mechanisms include hydrogen bonding, dipoledipole, induced dipole and pi-pi interactions. The exclusion of water from the sample prior to loading the sample onto the extraction column is of paramount importance in this mode of extraction/clean-up. For this reason, the use of polar sorbents in the sample preparation step for analysing organics in an aqueous environment is normally preceded by an initial liquid-liquid extraction of the aqueous sample using a relatively non-polar solvent.
Elution of polar compounds from one of the polar sorbents is achieved using a polar solvent mixture. This might consist of a non-polar solvent to which a polar modifier such as an alcohol is added. A mixture of alcohol/water may also be appropriate.
3.3 Ionic interactions The third class of interactions available using bonded silica
sorbents is ionic. A wide choice of sorbents from the weak anion exchangers (NH2' PSA, DEA),strong anion exchanger (SAX), weak cation exchanger (CBA) and the strong cation exchangers (PRS and SCX) provide an adequate range of sorbents for extracting isolates that can carry a positive or negative charge.
Isolate retention by ion exchange is generally from an aqueous sample. However, the extraction of ionic compounds from organic solvents is also a possibility. Retention and elution of isolates and interferences can be influenced by judicious use of pH control and selection of appropriate buffer concentrations. In general, where an isolate contains an ionizable group, the extraction of the isolate using an ion exchange mechanism is often capable of producing a cleaner extract when compared with a non-polar CIB phase. There are normally more compounds present in a water sample capable of being retained by a CIB sorbent compared with an ion exchange sorbent.
4. PRACTICAL ASPECTS ON THE USE OF NON-POLAR SORBENTS The use of non-polar sorbents in the extraction and clean-up of
water samples prior to the analysis of organic pollutants has been the method of choice for the vast majority of scientists who use solid phase extraction technology. The discussion and applications which follow reflect the importance of the role that non-polar interactions have in this application area.
The development of robust and reproducable sample extraction and clean-up procedures using bonded silica extraction columns is dependent upon attention being given to a number of important aspects of their use.
-~-
These include control over parameters such as flow rate of the samples through the extraction columns, pH control of the sample, interference and isolate elution solvents.
In the extraction of water samples containing particulate matter, attention has to be given to the possibilities of a partitioning of the isolates between the water and particulate matter contained in the water samples. This consideration can of course apply to other sample preparation procedures used in the extraction of water samples. The addition of either methanol, ethanol or iso-propyl alcohol to water samples at concentrations up to 15% v/v has been used to overcome adsorption of isolates by particulate matter prior to loading the samples on the extraction column. For samples high in suspended matter concentration, a filtration step prior to loading the water sample on the column may be necessary to prevent column blockage.
In the extraction of water samples using non-polar sorbents such as C1a, Ca, C2' Cyclohexyl or Phenyl, two aspects of their use require particular attention. First, the technique used to prepare the sorbents, i.e. the solvation of the sorbent. Second, the maintenance of sorbent solvation during the course of loading water samples onto the extraction column.
4.1 Preparation of the sorbent prior to sample loading Whether obtained as a bulk material or prepackaged in a
commercially available extraction column, bonded silica sorbents are invariably found in a dry state prior to use. The significance of this fact is that in the dry state the covalently bonded functional groups will be lying flat on the surface of the silica and will not significantly interact with the isolates that pass through the sorbent bed. To activate the sorbent it is necessary to solvate or wet both the surface of the silica and the bonded functional group. Solvents capable of achieving this goal include methanol, ethanol, iso-propyl alcohol, acetonitrile and tetrahydrofuran, to name a few. Between 10 and 20 bed volumes of solvent are required to achieve a solvated surface and the excess solvent is removed from the sorbent beds by a wash with water or buffer.In this context the bed volume is defined as the volume of solvent required to fill all interstitial spaces between the sorbent particles of a given mass of sorbent packed into a cylindrical bed. For 40~, 60 Angstrom sorbents,bed volumes are approximately 120~1/100mg of sorbent. After this solvation step, the functional groups are standing up on the surface of the silica. A small amount of the wetting solvent will remain associated with the sorbent that is now in a retentive state. This trace of wetting solvent will maintain a fully solvated state provided only a small volume of aqueous sample is loaded onto the column. This volume will be influenced by the choice and mass of sorbent as well as the solvation solvent used. For a 500mg SPE column, this volume could be as low as 25ml.
4.2 Maintenance of sorbent solvation during the extraction of large volume water samples.
When sample volumes in excess of 25ml are loaded onto a 500mg extraction column, a decrease in extraction efficiency may be observed for some isolates. This is particularly noticeable with more polar molecules. The reason for this decrease in efficiency is that the water sample passing through the sorbent bed gradually removes the residue of the solvation solvent associated with the sorbent. The functional groups start to fold over towards the surface of the silica and the sorbent becomes less retentive. A successful solution to this problem is to add
-36-
between 1-10% methanol to the water sample prior to loading the sample onto the extraction column (4, 5, 6). This technique has the advantage that the addition of organic solvent to the water samples can minimize the loss of non-polar isolates by adsorption onto the sample container walls or onto suspended matter in the sample. Although methanol has generally been the solvent employed in this role, there are good reasons to believe that iso-propyl alcohol may be a better choice (7). Iso-propyl alcohol has a more hydrophobic character compared to methanol and will interact more strongly with the non-polar sorbent. For this reason there will be a greater tendency for the iso-propyl alcohol to remain associated with the sorbent. More work has to be carried out on this aspect of sorbent extraction in order to establish the ground rules for the optimum selection of solvents and concentrations to be added to large volume water samples prior to extraction.
These solvation requirements are also applicable to the ion exchange sorbents and polar sorbents such as eN and Diol when used in the non-polar mode. For polar sorbents used in a polar mode it is normally appropriate to condition the column with the same non-polar solvent in which the isolates are dissolved.
5. APPLICATIONS There are now over one hundred references in the literature citing
the use of solid phase extraction in the isolation of organics from water. The applications highlighted here have been selected to illustrate the potential of SPE columns to efficiently extract, concentrate and purify acidic, neutral and basic isolates from water samples.
5.1 Organochlorine pesticides Andrews and Good evaluated a range of SPE columns for the
extraction of organochlorine pesticides from water (4). After preparation of the various sorbents with methanol followed by a water wash, 100ml samples spiked with standards at the 1 ~g/l level were loaded onto the columns. After drying the columns, the pesticides were eluted with ethyl acetate. Table II lists the recoveries obtained using the various sorbents. The optimum choice was found to be the Cs sorbent.
This study also covered the effects of adding methanol to the water sample prior to extraction. The results listed in table III indicate the need to add an appropriate solvent to the water sample in order to maintain solvation of the sorbent for efficient extraction.
Table II. Recovery of Various Pesticides vs Bonded-Phase Chemical Functions
Recovery ("!o) Sorben! Lindane Heptachlor Aldrin Endosulfan Dieldrin DDT
C18 Heavy loading 98 80 81 100 89 86 C18 Lighlloading 93 n 87 100 94 83 C8 105 79 86 102 92 79 Diphenyl 100 72 79 96 89 73 Cyclohexyl 89 76 78 87 86 72 C4 104 76 88 100 92 70 C2 102 76 84 97 96 73 Cyano 64 73 78 94 90 65
- 37-
Table III. Recovery of Pesticides for Various Types of Water Samples
Recovery (%)
Sample Undana Heptachlor Aldrin EndosuHan Dieldrin DDT
Distilled water 93 83 81 94 89 85 Saltwater 106 85 88 89 94 87 5% methanol 98 93 85 95 94 87 10% methanol 102 91 88 102 102 90
from ANDREWS and GOOD (1982)
5.2 Neutral priority pollutants Chladek and Marano developed an SPE column method for acidic and
neutral priority pollutants in wastewater as an alternative to the U.S. EPA method 625 involving liquid-liquid extraction (5). For the neutral pollutants (table IV), they agreed with Andrews and Good on the selection of the Ca sorbent. The effect of the addition of methanol to the water samples was also evaluated with spiked pure water samples and wastewater.
Table V highlights the need for the addition of methanol to the sample. For some phthalates the addition of up to 40% methanol was necessary to achieve good recoveries when spiked pure water was used. However, when these experiments were repeated using real wastewater samples, the addition of 10% methanol was found to be adequate. It was proposed that the presence of surfactant-like compounds and/or fine particulates in the wastewater samples were keeping the less soluble pollutants in suspension. The method is outlined in fig. 2.
Table IV. Recovery of Neutral Pollutants Using Different Sorbents
Sorbent type C2 C4 C8 C18
TableV.
Percent methanol
o 10 20 30 40 50
Recovery (%) Naphthalene Dimethyl- Anthracene & Di-n-butyl-
phthalate phenanthrene phthalate 57 66 59 66 77 88 61 80 90 105 80 90 80 85 67 83
Recovery of Neutral Pollutants From Water Containing Different Amounts of Methanol
Recovery (%) from C8 Sorbent Naphthalene Dimethyl- Anthracene & Di-n-butyl-
phthalate phenanthrene phthalate
90 105 80 90 70 50 60 82 80 17 60 90 70 3 65 90 46 1 70 87 8 0 36 60
from CHLADEK and MARANO (1984)
- 38-
Butylbenzyl- Di-n-octyl-phthalate phthalate
41 9 62 8 60 15 72 6
Butyfbenzyl- Di-n-octyf phthalate phthalate
60 15 65 20 73 28 84 22 96 65 60 60
NEUTRAlS ACIDS
200 -1000mL !+- ADJUST pH OF WATER SAMPLE T04- 7
150 ml ... ADD OF WATER SAMPLE 37.5g NaCI
~ ADO 10% ADJUST pH
METHANOL TO 1.- '2 , ~ ~WRBENT-V V
PASS THROUGH CONDITIONED
oem BONDED PHASE CARTRIDGES
CVClOHEXYl SORBENT -y ELUTE THE ANAL YTES
FROM BOTH CARTRIDGES WITH 2m! OF
ETHYL ACETATE
ANALYZE BOTH ELUANTS BY GC OR GC/MS
Rg 2. Procedure for the Extraction of Neutral and Acidic Priority Pollutants from Water from CHLADEK AND MORANO (1984)
- 39-
~ ELUTE THE ANALYTES
WITH2mLOF ACETONITRILE
~ ANALYZE ELUANT BY GCIMS OR LC
5.3 Acidic priority pollutants In the study of Chladek and Marano, phenol proved to be the least
retentive of the acidic priority pollutants for the sorbents evaluated. The extraction systems were evaluated and optimized for phenol because it was the worse case compound. Table VI lists the extraction efficiency of phenol for a number of different sorbents. This data indicated that the Cyclohexyl sorbent gave the most promising results. It was found that the extraction efficiency could be improved by the addition of NaCl to the sample in concentrations up to 25% w/v. The pH of the samples was also adjusted to 1-2 in order to suppress ionisation of the phenols. The use of Teflon water containers was also recommended to reduce losses of some phenolic pollutants on glassware. Table VII lists recoveries obtained under these optimized extraction conditions (fig. 2).
Table VI Recovery of Phenol Using Various Sorbents
Sorbenl Recovery (%)* C4 <10 C8 10
C18 20 CN <10 Cyclohexyl 37
Based on extraction of lOOcm3 of water spiked at I-ppm level for phenol. No NaQ was added to the water prior to extraction. Methanol was used 8S elution solvent.
Table VII. Recovery Data for Acid Pollutants using Optimized Cyclohexyl Sorbent ExtracUon Procedure
Compound
phenol 2-chlorophenol 2-nilrophenol 2.4.<fimelhylphenol 2.4.<fichlorophenol 4-chloro-3-melhylphenol 2.4.6-lrichlorophenol 4-nilrophenol 2.4.<finitrophenol 2-melhyl-4.6.<finilrophenol pentachlorophenol
from CHLADEK and MORANO (1984)
Recovery (%)
90 94 80 86 85 87 75 92
100 100 80
5.4 Herbicides 2,4-dichlorophenoxyacetic acid and picloram. In an extraction procedure developed by Wells, two herbicides,
2,4-dichlorophenoxyacetic acid (2,4-0) and picloram were simultaneously extracted from water (8). By optimizing the pH and solvent strength of the eluting solvents, it was possible to achieve a complete separation of the two herbicides prior to their individual analysis using different HPLC methods. The extraction scheme, illustrated in fig.3, involves the initial acidification of the water sample prior to loading onto a C18 column. By suppressing the ionization of the carboxylic acid group on both herbicides, retention of these compounds by the C18 sorbent is enhanced. Selective elution of the protonated picloram using a 25% acetic
- 40-
AQUEOUS SAMPLE (200mL) ADJUST TO pH 2.0
CONDITION COLUMN
-- C18 SORBENT ~~=l
AQUEOUS FRACTION / DISCARD
ELUTION, 250/0 HOAc PIC LORAM
1 ELUTION,MeOH
2,4·0
CI
CI 0- GH2 -GOOH
Fig. 3 Extraction Scheme for the Selective Desorption of Pidoram and 2,4-0
- 41 -
CI
......... OH G
" o
acid solution is feasible because under these pH conditions 2,4-D is more hydrophobic and will retain on the CIB sorbent until it is finally eluted with methanol.
5.5 Polynuclear aromatic hydrocarbons (PAR's) The extraction of PAR's using ClB sorbents has become a routine
procedure in a number of laboratories and has proved adequate for most analysts using HPLC separation in conjunction with fluorescence detection. This approach has been modified by Benjamin in an effort to improve the purity of his PAR extracts which were subsequently analysed by capillary GC using flame ionisation detection (6). This modified procedure is illustrated in fig. 4. It involves the addition of a layer of amino propyl sorbent (NH2) above the top frit of a ClB extraction column. The function of the NH2 sorbent is to remove a significant amount of polar material that would otherwise co-elute with the PAR fraction from the extraction column. The role of NH2 sorbent in efficiently extracting humic acids from water has already been established (9). The use of the NH2 sorbent as an anion exchanger in conjunction with a nonpolar sorbent shows promise in the extraction of water samples . The possibility of being able conveniently to combine more than one sorbent by stacking commercially available extraction columns using an adapter further enhances the potential usefulness of this technique(lO).
6. CONCLUSION When faced with the analytical problem of extracting low level
organic micropollutants from water samples a number of different approaches can be taken. The use of bonded silica extraction columns offers some interesting opportunities in the combined extraction, concentration and purification of the isolates of interest. Method development using this technique is much simplified by a sound knowledge of the characteristics of the various bonded silica sorbents and when considered along with a number of properties of the isolates to be extracted, for example solubility data, pKa values and any existing chromatographic data, the task of developing an extraction/clean-up procedure using this technology is much simplified.
The development of commercially available bonded silica extraction columns was originally focused on the needs of the analyst involved in the extraction of drugs from small volumes of biological samples. However, the needs of the analyst involved in extracting nanogram to picogram/litre concentrations of organic micropollutants from water has now been recognised. The commercial availability of stainless steel frits as a replacement for the porous polyethylene frits, to minimize phthalate contamination of the sample, is an illustration of this. It is anticipated that further developments can be expected in the manufacture of ultra clean extraction devices utilizing bonded silica sorbents specifically designed for environmental sample preparation.
-~-
AQUEOUS SAMPLE (200mL) ADO 2mL METHANOL
-i 1. 50:50 MeCI2:1,1,I·TRICHLOROTRIFLUROETHANE (TCTFE) 2. TCTFE
CONDITK>N COLUMN 3. DRY COLUMN I 4. MEOH + 5. WATER
DRY COlUMN~ AQUEOUS FRACTION VACUUM MANIFOLD DISCARD
ELUTE PAH'S WITH TCTFE
Fig.4 Extraction Scheme for Polynuclear Aromatic Hydrocarbons in Water from BENJAMIN (1985)
- 43-
REFERENCES 1. BRAUN, T., GHERSINI, G. Extraction Chromatography, Journal of
Chromatography Library, Vol. 2. Elsevier Scientific Publishing Company (1975)
2. STRACHAN, M.G., JOHNS, R.B. Anal. Chem., 59, 636 (1987) 3. BLEVINS, D.O., BURKE, M.F., GOOD, T.J., HARRIS, P.A.,
VAN HORNE, K.C.,YAGO, L.S. Sorbent Extraction Technology; (ed.VAN HORNE. K.C.) Analytichem International Inc. (1985)
4. ANDREWS, J.S., GOOD, T.J. Am. Lab., 4, 70 (1982) 5. CHLADEK, E., MARANO,R.S., J. Chromatogr. Sci., 22, 313 (1984) 6. BENJAMIN, J. Proceedings of the Second International Symposium
"Sample Preparation and Isolation using Bonded Silicas", Philadelphia, 57-62 (1985)
7. BURKE, M.F. Univ. of Arizona. Personal Communication 8. WELLS. M.M.J. Proceedings of the Third Annual International
Symoposium "Sample Preparation and Isolation using Bonded Silicas", Cherry Hill, 117-135 (1986)
9. CSIKY, I., MARKO-VARGA, G., JONSSON, J.A. Anal. Chim. Acta., 178, 307 (1985)
10. OZRETICH, R.J., SCHROEDER, W.P. Anal. Chem., 58, 2041 (1986)
-44-
POSTER SESSION I ANALYTICAL METHODOLOGIES
Evaluation of steam distillation-extract"ion procedure for the recovery of phenols in water
Isomer-specific determination of PCDD/PCDF in water leachate of a waste landfill
Pesticide micropollutants in Lake Albufera (Spain)
Techniques for the improvement of gasoline analysis
An improved extraction method for the quanti tati ve analysis of pesticides in water
Hyphenated methods (TSP LC-MS, DLI LC-MS, LC-TID) for analyzing organophosphorus priority pollutants
PCB's and organochlorine pesticides in eel and flounder in the Tagus estuary
HPLC/Fluorescence spectrometry in analyses of pulp mill wastes in recipients
Chlorophenol compounds in snow
The analysis of odorous sulphur compounds by gas chromatography after thermal desorption from tenax
Determination of organic chemicals in sediments taken from three unpolluted estuaries in South West England
Organic phosphates in surface, ground and drinking water
Evaluation of degree of pollution of Tiber and Aniene rivers by nitrilotriacetic acid
Hydrocarbons in East Mediterranean sea: determination and occurrence in the sediment of considered polluted and unpolluted areas of coastal environment
EVALUATION OF STEAM DISTILLATION-EXTRACTION PROCEDURE FOR THE RECOVERY OF PHENOLS IN WATER
. M.T. GALCERAN and F.J.SANTOS Department of Analytical Chemistry. University of Barcelona,
Diagonal, 647, 08028 Barcelona (Spain)
Summary
The quantitative performance of steam distillation extraction for the isolation of nine priority pollutant phenols from water was studied, and the recoveries compared with those of the direct extraction with dichloromethane and the two step extraction using tetrabutylamonium as ionpair reagent. For a concentration of 1 mg 1-1 using acidification, strong salting of the water sample and a distillation time of 1.S h. the recoveries are similar to those obtained with extraction in acidic medium with dichloromethane. Compounds of high polarity with strong interactions with water by hydrogen bonds, as p-nitrophenol, are not extracted by stream distillatio-extraction method.
1. INTRODUCTION Environmental aspects of phenol and its derivatives have
becomeincreasin-gly important in recent years. Trace amounts of phenolic compounds can give detrimental effects on water quality. Phenols are toxic to aquatic life and mammals and can impart objectional tastes and odours to water and fish ( 1,2,3 ). As a result of the toxicity of these compounds, the U.S.E.P.A. has included eleven phenols among the list of compounds on the Priority Pollutant List (4).
Several authors (3,4) have analyzed phenols in water with detection limits in the ug 1-1 level but the analysis must be generaly made by using a preconcentration step prior to determination. Various methods, procedures and devices for trace enrichmenthave been published. Unfortunately little has been reported on evaluation and comparison of these methods and systems for different classes of compounds and concentration levels.
Recently Rijks (6) has proposed to use steam distillationextraction as a pre concentration technique for phenols. This technique was introduced by Likens and Nickerson in 1964 (7) and modified later to develop a micro-version for analytical application (8,9). It was shown that high extraction efficiencies are obtained in a relatively short time for phenol and methylphenols (6,10).
Many analytical methods for the determination of phenolic compounds by gas and liquid chromatography have been reported (12). The analysis by gas chromatography can be carried out as
- 46-
free phenols (13) or derivatizated as acetates, or using an halogen containing reagent to enhance the response of the ECD (15, 16,17). The liquid chromatographic separation of phenols have been reported by some authors (11,18,19, 20) using reverse and normal phases.
In this work the isolation of the phenols more frequently present as priority pollutants in water using continuous steam distillation-extraction is investigated. The recoveries are compared with those obtained using conventional extraction with dichloromethane and a two step extraction using tetrabutylamonium chloride as ion-pair reagent proposed by Realini (11).
2. EXPERIMENTAL
Reagents and materials The solvent dichloromethane (Carlo Erba) used for the ex
traction was redistilled in an all glass equipment to obtain sufficiently pure grade.
Phenolic standards were obtained from Merk, Fluka or Sigma and all were of reagent grade. pH adjustments were made by HC1, acetic acid and NaOH obtained from Panreac (Spain).The cationic ion-pair reagent, tetrabutylamonium chloride, was obtained from Carlo Erba.
The text mixture consisted of phenol, p-nitrophenol, o-chlQ phenol, o-nitrophenol, p-chlorophenol, 2,4-dichlorophenol, 2,4,6-trichlorophenol, 4,6-dinitro-o-cresol, pentachlorophenol. The concentration of the stock solution was about 500 mg l-lper componentin methanol. 2,4,6-tribromophenol was used as internal standard.
Mobile phases in HPLC were water (Culligan purified) and acetonitrile (Carlo Erba). All solvents were degassed and filtered, acetonitrile trough a Millipore PTFE filter, and the water trough a cellulose acetate filter, 0.45 urn pore size and 47 mm diameter.
Liquid chromatography All analysis were performed on a Kontron 620 liquid chrom~
tograph. This instrument has a two solvent system with gradient programming capability. An uv detector at 280 nm and a 150x4.6 mm 5u Rosil C18 (Altech) column were employed. The flow was 1.00 ml/min and injection volume 5 ul.
Fig.1 shows the chromatogram of a standard sample obtained using the solvent gradient proposed. Solvent A is water with 1% acetic acid and solvent B is acetonitrile with 1% acetic acid. The solvent gradient starts at 48% B and increases to 80% B in 3 min.
Procedure
Direct extraction 250 ml portion water sample was acidified with HCl to pH 2
The sample was then extracted with one 100 ml and three 50 ml portions of dichloromethane. For all the extractions the separating funnels were shaken for 10 minutes and the two phases allowed to separate for 10 min. The extracts were combined, concentrated, dryed with sodium sulfate and finally evaporated to dryness with a stream of nitrogen. The sample was redissolved
- 47-
in 1m1 of dichloromethane and injected onto the liquid chromatQ graph.
,00
00
80
10 ... ~eo
~ ~~O
40
30
20
' 0
I f--------tt----tt-'/
I.S.
0 10 ti_ lain)
1 PH-ENOL
2 p - N t UOPtI[ feal
) o· CHLOROP~EHOl
" o- HlTROPHENOl
S .-~Kl ORO · . · C R£50L
6 Z. 4 - 0 1 Ull OROPH(IIOl
7 4.6-01(l-tl ORG-o- (Il(S,OL
82. '. 6-l>ICKlOROPKE"Ol
1.S.lit lBRO~OPH(NOl
9: HH'U(.H1.0ROPH(HOL
Fig. 1 Chromatogram of the standard sample
Two step ion-pair extraction The first step is carried out as described before but the
water layer was saved and O.OlM tetrabutylamonium chloride was added to the water as a cationic ion-pair reagent. Next the pH was adjusted to 14. The sample was further extracted four times with 25 ml portions of dichloromethane. Both acidic and basic extracts were combined, concentrated, dryed with sodium sulfate and lastly evaporated to dryness with a stream of nitrogen. The sample is redissolved and injected onto the liquid chromatograph.
Stearn distillation-extraction The apparatus used for this study is shown in Fig 2. The
organic compounds are distilled from the water sample which is placed in the 250 ml flask, A. Simultaneously the extraction solvent, dichloromethane is distilled from the 10 ml flask B. The vapours are condensed by the cold finger cooled by passing the water through melting ice. The aquous and the organic phases are separated at the bottom of the central part of the apparatus C and return through their return arms D and E to the flasks. The distance of the two overflows to the bottom of the central part of the apparatus is critical for the separation of the two phases.
The boiling of the extracting solvent was started 10 to 15 minutes before the boiling of the water sample. After the required time was elapsed, 1.5 hours, the boiling of the water was stopped but the reflux of the solvent was continued for a further 10 minutes. Later the solvent in C were transfered to the flask B.
The extract in B, 1-2ml, was concentrated to 1ml and inje~ ted in the liquid chromatograph. The recoveries were calculated relatively to the internal standard. This standard was added to the solvent prior the extraction.
- 48 -
In order to improve the extraction the sample pH was adju~ ted to 1 with HCl and 60 g of NaCl were added.
o I )0 .. Ill
3. RESULTS AND DISCUSSION
Fig. 2. Micro-apparatus for continuous steam di§ tillation-extraction.
Reference mixtures were prepared by addition of a measured amount of stock solution to 2 ml of the extraction solvent" which contains a corresponding concentration of the internal standard. An identical amount of stock solution is used for the extraction after addition to 150 ml of distilled water using steam distillation extraction procedure. A third identical amount of stock solution is used for direct extraction -after addition of 250 ml of water , and a forth is used for the two step extraction using tetrabutylamonium chloride as ion-pair agent
The average recovery values obtained from samples of 1mg/l of each phenol using five different samples and the corresponding relative standard deviations are given in table I.
The recoveries obtained by extraction with dichloromethane in acidic medium are higher than the values described in the li terature, for instance 39% for phenol or 90% for pentachlorophe nol (11) in front of 56.1% and 96.8% obtained in this work.When using the ion-pair extraction method the recoveries are better but not so high as the values obtained by Realini (11).
Using steam distillation-extraction method the recoveries are similar to the values obtained with acid extraction but the standard deviations are lower. The recoveries are better than those obtained by Rijks (6) for the phenol and methylphenols and similar to the values obtained by Janda (10) using diethylether as solvent.
In the table I we can see that using steam distillation -
- 49-
extraction method the p-nitrophenol is not extracted. This fact is due to the role of the hydrogen bonding on the extend of volatilization of a phenolic compound in the steam distillationextraction process. The intermolecular hydrogen bonding between the molecules of the compounds and the water can produce chains and rings which dismish volatility.
'I'able 1. Recovery data Recovery
Compound Acid ext. Ion-pair ext. SDE
Mean RSD Mean RSD Mean RSD N: 5 N:5 N: 5
Phenol 56.1 5.8 63.9 3.8 55.1 3.8 p-Nitrophenol 68.0 7.2 78.7 6.8 o-Chlorophenol 74.4 7.1 82.4 3.0 66.2 5.9 o-Nitrophenol 80.9 8.1 84.2 6.6 80.8 1.2 p-Chloro-m-cresol 87.7 2.3 94.4 1.4 93.0 2.4 2,4-Dichlorophenol 82.3 2.3 88.8 4.5 75.4 1.6 4,6-Dinitro-o-cresol 93.5 3.6 96.4 2.7 71.1 4.4 2,4,6-Trichlorophenol 84.5 2.4 86.0 7.1 88.0 3.3 Pentachlorophenol 96.8 2.1 97.4 1.3 78.6 5.2
The steam distillation extraction method can be used for the e~ traction of phenols, but the recoveries for p-nitrophenol are low according the values obtained by Norwitz (21) for meta and para phenols in the distillation process.
4. CONCLUSIONS Considering the quantitative recoveries, the process rate
and the enrichment factor, the steam distillation-extraction method give similar results as those obtained with extraction in acidic medium. Advantatges of the SDE are the dual isolation of phenols that dismished the simultaneous extraction of interfering compounds and the relatively small amount of solvent used which prevents solvent contamination.
A disadvantatge is the low recoveries for compounds as the p-nitrophenol and other nitro and amino phenols that can be extracted by the direct extraction method.
REFERENCES
(1) Environmental Protection Agency, Quality Criteria for Water , Superintendent of Documents, US Government Printing OffkE Order NQ 005-001-01049-4 Washington,DC (1976)
(2) V.P.Kozak, G.V.Simsiman, G.Chester,D.Stansby and J. Harkin. Reviews of the environmental effects of pollutants: XI.ChlQ rophenols. Oak Ridge Nat. Lab. ORNL/EIS-128 (1979)
(3) L.Renberg in A.Br¢rseth and G.Angeletti (Ed). Proceedingstt the Second European Symposium on the Analysis of Organic Mi cropollutants in water. Ki1larny 1981. D.Reide1. London (1982) .
(4) Environmental Protection Agency, Toxic Substance Control kt USEPA, Washington,DC, (1979)
(5) K.A.Pinkerton, J.High. Resolut.Chromatog. Chromatogr.Commun 4, (1981)
- 50-
(6) J.Rijks, J.Cuvers, Th.Noy, C.Cramers. J.of Chromatog. 279 (1983) 395
(7) S.Likens, G.Nikerson. Proc.Amer.Soc.Brew.Chem. (1964), 5 (8) M.Goodefroot, P.Sandra, H.Verzele. J.Chromatog. 203 (1981)
325 (9) M.Goodefroot, M.Stechele, P.Sandra, M.Verzele, J.High Reso-
lut. Chromatog.Chromatog. Commun. 5 (1982) 75 (10) V.Janda,K.Krijt, J.of Chromatogr. 283 (1984) 309 (11) P.A.Realini, J.of Chromatogr. Sci. 19 (1981) 124 (12) E.Tesarova,V.Pakakova, Chromatographia 17 (1983) 269 (TI) C.Lenenberger, R.Coney,J.W.Craydon,E.Molnar-Kubica, W.Giger
Chimia 37 (1983)' 345 (14) R.C.C.Wegmann, A.W.M.Hofstee, Water Res. 13 (1979) 651 (15) R.T.Cotts, E.E.Hargersheiner, F.M.J.Pasutto, J. of Chroma
togr. 179 (1979) 291 (16) L.Renberg, Chemosphere 10 (1981) 767 (17) R.S.K. Buisson, P.W.W.Kirk, J.N.Lester, J.Chromatogr.Sci.
22 (1984) 339 (18) G.Blo, F.Dondi, A.Betti, G.Bighi, J. of Chromatogr.257(19~ (19) B.Shultz, J.of Chromatogr. 269 (1983) 208 (20) N.G.Buckman, J.O.Hill, R.J.Magee. J.of Chromatogr.284 (1984)
441 (21) G.Norwitz, N.Nataro,P.N.Keliher. Anal.Chem. 58(1986) 639
- 51-
SUMMARY
ISOMER-SPECIFIC DETERMINATION OF PCDD/PCDF IN WATER LEACHATE OF A WASTE LANDFILL
C. Ftlrst, L. Stieglitz, G. Zwick Kernforschungszentrum Karlsruhe
D-7500 Karlsruhe, F.R.G.
A clean-up procedure for the isomer-specific analysis of polychlorina ted dibenzodioxins (PCDDs) and dibenzofurans (PCDFs) in oil extracts from water leachate of a hazardous waste landfill is described. The samples were chromatographed on Alumina B-Super I, silica gel combined with silica gel/44 % conc. H2S04 and on Bio- Beads S-X3. Separation of 2,3,7, 8-tetra-CDD and final sample purification were performed on Alumina B-Super I (micro column). Among the 2,3,7,8-tetra- to hexa-CDDs/CDFs, 2,3,7,8-TCDD was by far the most abundant isomer, determined at a mean concentration of 70,S ppb. Industrial wastes from 2,4,5-trichlorophenol production are assumed to be the main source for this high concentration. From the isomer distribution pattern of the hepta-CDFs, pentachlorophenol can be considered as source of the higher chlorinated PCDDs/PCDFs.
INTRODUCTION Since 2.3,7,8-TCDD was detected in leachates of the landfill Georgs
werder/Hamburg (F.R.G.) in 1983 (1), the isomer-specific determination of PCDDs and PCDFs in different samples of hazardous waste landfills has been the subject of much concern in recent years: they were detected in water and oil leachates (2,3), in bottom sediments (3) and in PCB oil (4). But there are no reports on PCDD/PCDF determination in oil extracts from water leachates, nor do the reports of the leachates include any description of the sample pretreatment and clean-up. Based on a clean-up procedure for the PCDD/PCDF analysis in oil samples (5), a method for the PCDD/PCDF determination in oil extracts from water leachate of a waste landfill was developed and optimized with respect to the complexity of the sample material. From the resulting isomer profiles, clues for possible PCDD/PCDF sources can be obtained.
2. EXPERIMENTAL The oil extracts of the leachates (extraction ratio oil: water, 1:500)
were stored in the landfill in 500-liter-barrels. Samples were taken from the upper part of the barrels containing the liquid oil and from the bottom layer.
a) An amount of 50 g of the liquid oil extract was homogenized in an ultrasonic bath (stock material 1). Then 2 g of stock material 1 were dissolved in 10 ml of benzene. The following C13-1abelled PCDDs were added in concentrations of 25 to 80 ng: 2,3,7,8-tetra-CDD, l,2,3,7,8-penta-CDD, l,2,3,6,7,8-hexa-CDD, l,2,3,4.6,7,8-hepta-CDD and octa-CDD.
- 52-
b) A quantity of 200 g of the sample taken from the bottom layer (stock material 2) was homogenized with an ultrasonic probe at 60oC. 3 g of stockmaterial 2 were transferred to 300 ml of toluene. After addition of the same C13-labelled PCDDs as described above, the mixture was refluxed (24 hrs), and filtered. The filtrate was concentrated and dissolved in 10 ml of benzene.
The pretreated samples of stockmaterial 1 and 2 were cleaned up as follows: each sample was applied to a column of 20 g Alumina B-Super I and 10 g of Na 2S04 , prewashed with 400 ml of hexane. After sample application, 60 ml of benzene and 400 ml of hexane/dichloromethane (98:2) were passed through the column. The PCDDs/PCDFs were eluted with 100 ml of hexane/ dichloromethane (1:1).
The concentrated eluate was applied to a column, filled from bottom to top with 10 g of silica gel, 20 g of silica gel/44 % cone. H2S04 and 10 g of Na 2S04 (6), prewashed with 150 ml of hexane. The PCDDs/PCDFs were eluted with 150 ml of hexane.
The concentrate was chroma to graphed on a column of Bio-Beads S-X3, equilibrated with 400 ml of cyclohexane/ethylacetate (1:1). First 120 ml of cyc1ohexane/ethylacetate (1:1) were passed through the column, the PCDDs and PCDFs were eluted subsequently with 60 ml of the same solvent mixture. Sample application and elution were performed by flash chromatography (7) using nitrogen pressure (10 ml N2/min, 3 bar).
After complete removal of the solvent, the residue was redissolved in benzene and applied to a column filled with 5 g of Alumina B-super I and 3 g of Na 2So4 , prewashed with 100 ml of hexane. The PCDDs/PCDFs were eluted with 50 ml of hexane/dichloromethane (80:20), except for 2,3,7,8-TCDD, which was eluted subsequently with 20 ml of hexane/dichloromethane, 70:30 (8). Recoveries were determined by addition of 13c6 - 1 ,2,3,4 TCDD (25 ng). Then both fractions were concentrated to about 20 1 and analyzed by GC-MS.
The GC/MS analyses were carried out with a HP 5890 gas chromatograph and a HP 5970 mass-selective detector. A 40 m SP 2331 fused silica capillary column (0,25 mm i. d.) was used. Identification of the isomers was based on their retention times (9,10). Quantification was achieved via the C13-labelled internal standards.
3.RESULTS AND DISCUSSION In Tables I-III, the PCDD/PCDF contents of three samples (sample No
A,B,C) from stock material 1 are listed with their mean value x and standard deviation s. Among the tetra-CDDs 2,3,7,8-TCDD is by far the most abundant isomer at a mean concentration of 70,S ppb. The mass fragmentograms from the 2,3,7,8-TCDD fraction are shown in Fig. 1. At m/z 320, only 2,3,7,8-TCDD is detected. The 2,3,7,8-TCDD in the PCDD/PCDF fraction (m/z 320, Fig. 2) corresponds only to 3-4 % of its total amount. The isomer profiles of the PCDDs and PCDFs are shown in Fig. 2-3. From the hepta-CDFs, an isomer distribution pattern is obtained as described generally for pentachlorophenol: 1,2,3,4,6,8, 9-hepta-CDF predominates over 1,2,3,4,6,7,8-hepta-CDF, whereas in fly ash samples the 1,2,3,4,6,7,8 congener is by far the most abundant isomer of all the other hepta-CDFs.
As expected, the PCDD/PCDF amounts of the sample taken from the bottom layer (stock material 2) were significantly lower than the concentrations obtained from the liquid oil extract.
Concerning the question of possible sources for the PCDDs and PCDFs in the samples analyzed, the following might be assumed: the eminently high concentration of 2,3,7,8-TCDD is generated by industrial wastes from 2,4,5-trichlorophenol production deposited in the landfill. From the distribution
- 53-
pattern of the hepta-CDFs, pentachlorophenol is assumed to be a possible source of the higher chlorinated PCDDs/PCDFs.
REFERENCES
(1) Schumacher, E. in: Dioxine, Erich Schmidt Verlag Berlin, 1985, 81-84.
(2) Gtltz, R., Vom Wasser, 65 (1985) 215-228. (3) Gtltz, R., Chemosphere, 15 (1986) 1981-1984. (4) R. E. Adams, M. M. Thomason, D. L. Strother, R. H. James
and H. C. Miller, Chemosphere, 15 (1986) 1113-1121. (5) H. Hagenmaier and H. Brunner, Fresenius Z. Anal. Chem.
324 (1986) 23-26. (6) L. L. Lamparski and T. J. Nestrick, Anal. Chem., 52
(1980) 2045-2054. (7) W. C. Still, M. Kahn and A. Mitra, J. erg. Chem., (1978)
2923-2925. (8) H. Hagenmaier, H. Brunner, R. Haag and M. Kraft,
Fresenius Z. Anal. Chem. 323 (1986) 24-28 (9) H. R. Buser and C. Rappe, Anal. Chem. 56 (1984) 442-448
(10) C. Rappe, Environ. Sci. Technol. 18 (1984) 78A-90A.
Fig. 1:
2,3,7,8 - TCDD
m/ z 320
20 22 24 26
13 C 2,3,7,8-TCDD 2
m/z 332
20 22 24 26
Mass fragmentograms at m/z 320(1) and m/z 332(2) from the 2,3,7,8-TCDD fraction of the liquid oil extract.
- 54-
en I"-
'" ru CD tD v ru
"-en v I"- '" v CD ru ru tD
'" ru
24 26
20
en "-'"
CD
"" '"
22
Penta-COOs CD I"- m/z 354 '" ru
I"-tD
'" ru
28
Tetra-COOs m/z 320
24 26
CD tD v
'" ru en CD tD CD
'" I"-en ru tD co '" tD N v N en
l"-tD
'" CD en ru I"-l"- v tD '" v ru N
30 32
Hexa-COOs
m/ z 390
en CD "-
'" ru
I"-tD v
'" N
34 36
Hepta-CDDs Oct a-COD en co co
m/z 424 en l"-tD v
'" ru
40
Fig. 2:
I"- "-tD m/z 460 tD v v '" '" N N
45 55 60
Mass fragmentograms of the tetra- to octa-CDDs obtained from the PCDD/PCDF fraction of the liquid oil extract.
- 55-
.,. '"
Tetra-CDFs OJ m/ z 306
'" I1l '" '" .,. N .,. '" '" "'N "-
'" ... -11l '" '" lD N '" -.., CD '" "-- "- N
'" OJ
OJ
, 20 22 24
<D "- "-.,.
'" <D N .,. "-- '" Penta-CDFs '" '" \- '" 11l.,. ... " '" _ I1l m/ z 340 '" "'-'" I1l '" N ... '" "
.,. I1l
'" '" '" '" N ... M ... v N N
M '" '" " N '" OJ
'" '" "- ... .,. "-I1l ... M M '" " '" N I1l N v ... N '" '" N M "- M <D
'" N " N
, 2 4 26 28 30 32 28
'" ... "-'" '" N ...
'"
'"
26
I1l "- '" ... "-
'" '" N M OJ
'" "-... M OJ
"-'"
31'3
I1l
'" '" M N
32 34
Hexa-CDFs m/ z 374
I1l
'" "-.., N
36 38
Hepta-CDFs
m/z 408
en Q)
'" ... M
Octa-CDF m/ z 444 en
'" <D
" '" ... M N
38
I1l "-
'" .,. M N
Fig. 3:
N " '" v M N
en <D "-... .., OJ
40 42 44 46 52 54
Mass fragmentograms of the tetra- to octa-CDFs obtained from the PCDD/PCDF fraction of the liquid oil extract.
- 56-
Table I: Concentrations (ng/g) of 2,3,7,8,-substltuted PCDDs In samples A,B,C from the liquid 011 extract.
Compound A B C X 5
2,3,7,8- Tetra- CDD 71,9 69,6 69,9 70,5 1,2
1,2,3,7,8- Penta- CDD 2,9 2,6 2,8. 2,8 0,1
1,2,3,4,7,8- Hexa- CDD 2,0 2,0 2,2 2,1 0,1
1,2,3,6,7,8- Hexa- CDD 8,3 8,9 8,7 8,6 0,3
1,2,3,7,8,9- Hexa- CDD 11,4 11,2 9,6 10,7 0,9
1,2,3,4,6,7,8- Hepta- CDD 35,2 42,1 44,2 40,5 4,7
Octa- CDD 90,2 116,0 117,3 107,8 15,2
Table II: Concentrations (ng/g) of 2,j,t,tl-SuOstltutea ~~UtS In samples H,~,C from the liquid 011 extract.
Compound A B C X 5
2,3,7,8- Tetra- CDF n.d. n.d. n.d. - -
1,2,3,7,8- Penta- CDF a 0,59 0,46 0,52 0,52 0,06
2,3,4,7,8- Penta- CDF 0,35 0,28 0,46 0,36 0,09
1,2,3,4,7,8- Hexa- CDFb 3,9 3,5 3,7 3,7 0,2
1,2,3,6,7,8- Hexa- CDF 3,5 3,9 4,6 3,9 0,5
1,2,3,7,9,9- Hexa- CDF 0,67 0,72 0,64 0,69 0,04
2,3,4,6,7,8- Hexa- CDF 0,94 1,1 1,1 1,0 0,09
1,2,3,4,6,7,9- Hepta- CDF 5,2 7,3 5,3 5,9 1,1
1,2,3,4,7,9,9- Hepta- CDF 0,76 0,71 0,79 0,75 0,04
Octa- CDF 30,4 50,3 36,2 39,9 10,2
anot separated from 12349- penta- CDF bnot separated from 123479- hexa- CDF n.d. a not detected
- 57-
Table III: Total PCDD/PCDF content (ng/g) of the C14- CIB congener groups In samples A,B,C from the liquid 011 extract.
-Compoundll A B C X 5
Tetra- CDDs 75,7 71,6 73,7 73,6 2,0
Penta- CDDs 9,1 B,1 B,6 B,6 0,5
Hexa- CDDs 55,7 57,B 49,B 54,4 4,1
Hepta- CDDs 53,3 65,3 64,1 60,9 6,6
Octa- CDD 90,2 116,0 117,3 107,B 15,2
total- PCDDs 2B4,0 31B,B 313,5 305,4 1B,7
Tetra- CDFs 6,B 6,6 6,1 6,5 0,4
Penta- CDFs 6,6 5,4 5,1 5,7 O,B
Hexa- CDFs 23,0 21,7 20,0 21,6 1,5
Hepta- CDFs 14,0 15,4 14,1 14,5 O,B
Octa- CDF 30,4 50,3 36,2 38,9 10,2
total- PCDFs 80,B 99,4 B 1,5 87,2 10,5
- 58-
Summary
PESTICIDE MICROPOLLUTANTS IN LAKE ALBUFERA (SPAIN)
J.M. CARRASCO Departamento de Biotecnologia
Universidad Politecnica de Valencia
The pollution by pesticides is studied in lake Albufera of Valencia (Spain) to determine the hazards involved in the pesticide treatment of nearly rice field and forest and in urban and industrial contamination.
At present Molinate, Benthiocarb and Fenitrothion are the pesticide which have been found to reach the highest concentration in waters from the lake.
DDT and BHC were the most frequently detected organochlorinated pesticides in this ecosystem.
1. INTRODUCTION The lake "Albufera of Valencia" is situated 10 Km to the south of
the town of Valencia and it constitutes one of most important humid areas of Spain from an ornithologic, touristic and economic standpoint.
The pesticides currently used for treating the cultivations and woods surrounding the lake could be a serious danger to its fauna and flora and knowing the levels of contamination involved is necessary in order to evaluate its effects.
Our research work into the pesticide micro-contamination suffered by this ecosystem started in 1.970 as a part of the Spanish contribution towards the O.C.D.E. sponsored "Pilot study on the Mediterranean region".
The insecticide (1) and urban waste (2) levels of contamination were considered, the evolution of DDT residues with the time vas evaluated (3) and finally, a special attention to weedkiller contamination was paid (4).
This paper collets some of the results of our surveys over lake contamination levels and its effects in living organisms.
2. SAMPLING Several sampling points distributed throughout the perimeter and
the middle of the lake were selected and samples of water at 10 cm. below the lake surface were taken.
Samples of plankton, plants, fishes and birds were also taken and analysed.
3. ANALYTICAL PROCEDURES One L sample of water was filtered through paper and then through
Whatman GF/C paper. Two mL IN HCL was added to each sample, and the samples were extracted with 3 portions of 100 mL n-hexane. The extracts were dried by shaking with 10 g. anhydrous Na2S04, filtered, concentrated to 2 mI. and inyected in the gas chromatograpn.
Analyses of plankton, plants, fishes and birds were performed according to the methods recommended by U.S. Food and Drug Administration (5).
- 59-
Nitrogen-phosphorus detectors and electron-capture detectors were used and columns were glass, packed with either 10% DC-200 or 10% DC-200 and 15% QF-1 on Gas-Chrom Q 80-100 mesh.
Perkin-Elmer F-11 and Sigma 3 and Hewlett-Packard 5890A gas chromatograph were used.
4. RESULTS AND DISCUSSION Organochlorinated insecticides Over the 1.970-1.972 period, the highest organochlorinated insecti
cide contamination in the lake's central waters appertained to both BCH wich in fact reached a level of 0.23 giL in autumm and DDT wich level was of 0.15 giL in spring. The average content of DDT in the lake's central water was of 0.11 giL.
The water analyses we have carried out over the last few years have recorded no DDT residues.
The larger concentration of the organochlorinated insecticides is found on the upper trophic strata and we have found some differences in DDT and BHC biomagnification.
The levels of organochlorinated insecticides found in waters, plants fishes and birds are lower than those which might be lethal to these living organisms and even that those which might adversaly affect the reproduction of birds by decreasing the number of viable eggs.
In the last year only one of the so analysed birds actually showed substantial residues of pp'DDE, metabolite of DDT of a weaker toxicity than the insecticide concerned and this seems to point out that the residues in question came from either an old contamination or other areas since the sub jet of our analyses was migrating birds really.
DDT exceeded in a few samplings over 1. 970-1. 972 the values which may be regarded as lethal to "Daphnia Pulex", the most sensitive organism to this insecticides. These data indicate that the DDT contamination may well affect the lake's microcrustaceous by altering the composition of its population in the favour of the most resistant organisms. Nevertheless, as we have verified for these 16 years of research work, abundant rainfall brings about the reappearance of the sensitive microcrustaceous which reproduction is made possible by the contribution of clean waters which in turn potentiate the bloom of the populations at the end of the winter.
Organophosphate insecticides In studing the organophosphorous insecticide contamination in lake
Albufera along these 16 years it has been found out that the treatment induced contamination is fully seasonal in character, owing to easy hydrolisis of these compounds.
Over the 1.970-1.980 period it was found that the maximum contamination resulting from the above products at water of lake Albufera was for fenthion which in fact reached lethal levels to microcrustaceous during several days in summer. Nevertheless the repopulation of the species in question was noticed when the contribution of clean waters lowered the levels of contamination.
At present, the fenitrothion based treatment normally brings about a low contamination in the lake's water in spring and summer but in general its levels lower donw to non detectable values in the course of a few weeks only.
-60-
Herbicides Molinate and Benthiocarb are the most widely used weed killers in
rice fields surrounding the Lake A1bufera. The samples with highest levels of Molinate were collected on the days when herbicide was applied. In water of the Lake, levels of Mo1inate are always below those considered lethal to fish.
Benthiocarb was detected only in May: the highest values were 5 ppb found in 1. 983 and 15 ppb in 1. 985 in front one channel outfall. In the middle of the lake, benthiocarb was detected only on May 29 and 30, 1.985, at 0.5 and 2.6 ppb respectively.
5. CONCLUSIONS The most frequently detected organoch1orinated insecticides were DDT
and BHC. Its levels are always below those considered lethal to fish and birds.
The organophosphate insecticides contamination is seasonal in character.
The herbicides Mo1inate and Benthiocarb do not reach levels that are lethal to fish.
When pesticide levels were lowered we have verified the repopulation of the sensitive organisms.
REFERENCES (1) CARRASCO, J.M., CUNAT, P, MARTINEZ,M., MARTINEZ, R.M. and PRIMO, E.
(1.972). Contaminacion de 1a A1bufera de Valencia. I. Nive1es de contaminacion por insecticidas. Rev. Agroquim. Tecno1. Aliment. 12,583-596.
(2) PRIMO, E., CUNAT, P., CARRASCO, J.M., BLANCO, M.C. and MARTINEZ, M. (1.975). Contaminacion de 1a A1bufera de Valencia. II. Nive1es de contaminacion por residuos urbanos. Rev. Agroquim. Tecno1. Aliment. 15, 98-112.
(3) PUCHADES, R. and CARRASCO, J .M. (1. 984). Contaminacion por DDT de 1a A1bufera de Valencia. XX Reunion Biena1 de 1a Real Sociedad Espanola de Quimica. Comunicacion 33-37.
(4) CARRASCO, J.M., PLANTA, M., GOMEZ-CASALS, V. and MORAGUES, V. (1.987) Pesticide Residues in Lake A1bufera, Valencia, Spain. J. Assoc. Off. Anal. Chern. 70, 4, 752-753.
(5) FOOD AND DRUG ADMINISTRATION (1.984) . Pesticide Analytical Manual Vol. I, F.D.A. Washington, DC.
- 61-
TECHNIQUES FOR THE IMPROVEMENT OF GASOLINE ANALYSIS
P.Slingerland and R.C.C.Wegman Laboratory for Organic-analytical Chemistry
National Institute of Public Health and Environmental Protection P.O. Box I, 3720 BA Bilthoven, The Netherlands
Abstract
For the automated analysis of gasoline components in water samples a gas chromatograph equipped with a multifunction controller, an oncolumn autosampler and a data acquisition system was used. Use of retention indices improves the reliability of the identification. However, components which differ less than 0.7 index units could not be discriminated by the available acquisition software. Further improvements of retention indices with the described equipment are not possible. Improvement of the GC-equipment is less important. A more accurate retention time registration is necessary. Control runs with GCMS are recommended. Pattern recognition techniques could be helpful to avoid serious mistakes in identification and quantitation.
1. INTRODUCTION
In order to describe adsorption and biodegradation behaviour of gasoline in soil and sediment, a large number of analyses in soil/sediment and water samples had to be carried out. In addition to the total amount of gasoline, more than 20 gasoline components (Table 1) had to be analysed in a quick and reliable way. For the automated analysis a gas chromatograph equipped with a multifunction controller, an on-column autosampler and a data acquisition system were available. One of the problems in the analysis of the individual gasoline components is the identification on base of retention time. Automatic determination of a specific component in a complex mixture is not very simple. Leaching water samples from soil columns, spiked with gasoline, show a large number of compounds in a relatively short GC-traject (see gas chromatogram in Figure 1).
Another problem is the wide range of concentrations which had to be measured. Most of the problems can be solved by using internal standards such as deuterated toluene and/or deuterated ethylbenzene. In order to improve the identification, retention time measurements have to be as accurate as possible. Retention indices are much better to reproduce, because they are less vulnerable to chromatographic conditions. At this moment the ultimate accuracy has been reached by Rijks et al. (ref.l) with isothermal chromatography of hydrocarbon test mixtures on a special constructed gas chromatograph. Kovats retention indices have been obtained with a precision corresponding to a standard deviation of less then 0.03-0.04 index units. In this paper the use of an automatic system is described for the analysis of gasoline components in water samples.
-~-
TABLE 1. COMPOSITION OF TESTMIXTURE
COMPOUND FORMULA MW CAS NO. RET. TIME (in min) -----~------------------------------------------------ --------------BENZENE C6H6 78.11 71-43-2 10.87 4-ME.CYCLOHEXENE C7H12 96.17 591-47-9 14.93 3-ME.CYCLOHEXENE C7H12 96.17 591-48-0 14.97 D8-TOLUENE C7D8 100.21 2037-26-5 15.78 TOLUENE C7H8 92.14 108-88-3 15.99 1-ME.CYCLOHEXENE C7H12 96.17 591-49-1 16.56 OCTANE C8H18 114.23 111-65-9 18.43 D10-ETHYLBENZENE C8D10 116.25 25837-05-2 20.81 ETHYLBENZENE C8H10 106.17 100-41-4 21.08 M+P XYLENE C8H10 106.17 108-38-3 21.53 STYRENE C8H8 104.15 100-42-5 22.47 NONANE C9H20 128.26 111-84-2 23.75 ISOPROPYLBENZENE C9H12 120.20 98-82-8 24.45 PROPYLBENZENE C9H12 120.19 103-65-1 25.96 1,3,5-TRIME.BENZENE C9H12 120.19 108-67-8 26.72 1,2,4-TRIME.BENZENE C9H12 120.19 95-63-6 27.97 DECANE C10H22 142.28 124-18-5 28.75 INDANE C9H10 118.18 496-11-7 29.90 UNDECANE C11H24 156.31 1120-21-4 33.38 NAPHTHALENE C10H8 128.17 91-20-3 36.55 DO DE CANE C12H26 170.34 112-40-3 37.69 1-ME.NAPHTHALENE C11H10 142.20 90-12-0 41.87 BIPHENYL C12H10 154.21 92-52-4 44.29 HEXADECANE C16H34 226.45 544-76-3 52.39 HEPTADECANE C17H36 240.47 629-78-7 55.57 DOCOSANE C22H46 310.61 629-97-0 69.37
2. EXPERIMENTAL
2.1 Extraction Procedure
Thousand m1 of water were extracted with 10 m1 of CS2 in a flask with a teflon lined screw cap by thoroughly mixing for half an hour. Thirty g of soil or sediment were mixed with the same amount of anhydrous sodium sulphate and extracted in a Soxh1ett apparatus with 100 m1 of CS2 during 16 h. From the CS2 extract 1 microliter was directly injected onto the gas chromatograph. The extraction solvent CS2 was fortified with D8-to1uene and D10-ethy1benzene at a concentration of about 10 ng/~l. High quality CS2, free of interfering components (benzene!) was obtained from J.T.Baker (product no. 8395), Deventer, The Netherlands.
2.2 Equipment
The analysis was carried out with a gas chromatograph type 5300 from Carlo Erba, with a standard FID, equipped with a multifunction controller and an on-column autosampler.
The data acquisition was carried out with a Nelson interface and 2600 chromatography software Version 4.0 with the possibility to link to user programs.
- 63-
This set-up could work unattendedly for several days; raw data files were automatically stored on a hard disc of a personal computer (Olivetti M24, 640 Kb internal memory, 30 Mb hard disc) on which the Nelson software.was running.
The Ultra 1, fused silica column, length 50 m, inner diameter 0.33 mm, cross-linked methyl silicon polymer (film thickness: 0.36 ~m) was obtained from Hewlett Packard. A length of 2 m silanised empty fused silica column was coupled with the analytical column as a retention gap. The gas chromatographic conditions were as follows: on-column injection
volume: 1 ~l; temp. program: 400 C (7 min) ----> 4oC/min ----> 3000 C ----> cooling; carrier gas: Helium at a flow rate of 3 ml/min; detector
(FID): 300oC.
3. RESULTS AND DISCUSSION
The determination of retention indices in programmed temperature gas chromatography consists of two steps. First an empirical relation between the retention indices of known compounds and their retention times must be derived. Secondly this relationship is used in calculating the retention indices of other compounds. The most simple method connects the known points with straight lines and has been called the linear interpolation method or the polygon method. Janssens (ref.3) used an improved linear interpolation, Jaeschke et al. (ref.2) used a calibration curve composed of parabolic parts. In our opinion a smooth function going through the given data set should, however, lead to more precise results. From the methods mentioned in literature the cubic spline interpolation can be considered to be the most suitable method in generating smooth curves (see also ref.6 and 11). Cubic splines are functions composed of third-order polynomials pieced together in the data points. The fact that they are twice differentiable in these junction points represents their essential property; consequently adjacent polynomials have a common tangent in any data point. The cubic splines offer the following advantages: smoothness, no introduction of extraneous oscillations, easily computable and a good fit to nonlinear as well as to linear data sets. In this investigation the cubic spline interpolation module from a graphic-statistical program "RRGRAPH" developed by Gerritsen and Van den Bleek of the Technical University Delft (ref.13) was used. Further details on spline functions are given in ref.12.
The first run of a number of samples was a known mixture of all components. In this run response factors as well as retention times were automatically updated in the existing method with the aid of the autocal-program. During the chromatographic run practically all peaks elute between 8 and 42 min. In general 7 separate compounds eluted within a min. If the peaks are equally distanced, each peak needs at least a window of about 8 s. A very narrow window can better assure the identification. However the chance exits that the compound slips out of the window. Use of an internal standard as a reference compound (in this study D8-toluene) improves the procedure because it can be easily located as the largest compound in a relatively wide window. Preliminary experiments showed that correcting the expected retention times relative to the reference peak made it possible to narrow the window down to about 3 s for most of the compounds.
-64-
The distance between two succeeding normal paraffins is about 300 s, which corresponds with a retention index of one index unit. However compounds can behave more irregular which leads to a wider window.
To calculate a correct window setting a testmixture was run 8 times. The chosen window-width was more than a double of the standard deviation. In Table 2 the statistical results of retention time calculations are summarised. The third column gives the results of 8 analyses of a test mixture and the fourth column shows the more realistic results of the analysis of 11 real samples. The mean standard deviation was 2.3. The largest standard deviations were found for respectively: dlO-ethylbenzene 3.0 s, ethylbenzene 3.2 s, xylene 9.0 s, l,2,4-trime.benzene 3.8 s, and l-me.naftalene 4.6 s. At the same time it was observed that the deviation of most components in the separate runs have the same direction (+ or -) and could point to a slightly less reproducible temperature course. As a reference component for the qualitative calculations we prefer d8-toluene (SD: 1.8 s) over dlO-ethylbenzene (SD: 3.2 s).
TABLE 2. RETENTION TIME WITH STANDARD DEVIATION FOR COMPONENTS OF GASOLINE
RETENTION TIME AND SD (in s)
COMPOUND x SD SD SD test mixture real samples corrected*
n-8 n=11 --------.-------------------------------------------------------------BENZENE 652.2 1.0 1.8 0.0 4-ME.CYCLOHEXENE 895.8 2.4 0.6 3-ME.CYCLOHEXENE 898.2 2.3 0.5 D8-TOLUENE 946.8 1.5 1.8 0.0 TOLUENE 959.4 1.4 1.9 0.1 l-ME.CYCLOHEXENE 993.6 2.1 0.3 OCTANE 1105.8 0.8 2.4 0.6 D10-ETHYLBENZENE 1248.6 1.3 3.0 1.2 ETHYLBENZENE 1264.8 1.3 3.2 1.4 M+P XYLENE 1291.8 1.3 9.0 7.2 STYRENE 1348.2 1.3 NONANE 1425.0 1.2 ISOPROPYLBENZENE 1467.0 2.1 0.3 PROPYLBENZENE 1557.6 1.1 1.5 -0.3 1,3,5-TRIME.BENZENE 1603.2 0.3 2.1 0.3 l,2,4-TRIME.BENZENE 1678.2 3.8 2.0 DECANE 1725.0 0.8 INDANE 1794.0 0.8 2.1 0.3 UNDECANE 2002.8 0.6 NAPHTHALENE 2193.0 0.5 1.8 0.0 DODECANE 2261.4 0.3 1-ME.NAPHTHALENE 2512.2 0.3 4.6 2.8 BIPHENYL 2657.4 0.3 HEXADECANE 3143.4 0.2 HEPTADECANE 3334.2 0.3 DOCOSANE 4162.2 0.7 ----------------------------------------------------------------------*SD - corrected for Internal Standard shift.
- 65-
Experiments with the use of retention indices showed a substantial increase in the accuracy. The mean standard deviation was 0.3 index units which corresponds with about 0.9 s (100 IU = 300 s). This accuracy is about the same of the data acquisition equipment of which the minimum time - interval is 0.01 min = 0.6 s which corresponds with 0.2 index units. Moreover, the equipment is not capable to discriminate between compounds which elute within 2 s corresponding with 0.7 retention index units. An improvement in calculating retention indices requires a more precise registration of the retention time but would not be very helpful in the analyses of such complicated mixtures as gasoline. Regularly a misinterpretation occured as some compounds showed a retention shift that could not be smoothed out with the reference compound. In comparing the results with former runs this kind of errors are easy to correct. In fact a "manual" pattern recognition technique is used. If these actions could be translated into a computer program, the ideal of a real automatic analysis loomed.
4. CONCLUSIONS AND RECOMMENDATIONS
* It is not possible to analyse a large number of gasoline compounds automatically in a CS2 extract at a complete reliable way without use of GC-MS or pattern recognition techniques.
* Measurements had to be started at the second GC run. * Retention times had to be measured with real samples. Runs of test
mixtures give a too flattering image. * The solvent has a relatively large effect on the retention times of
early eluting compounds. * In order to identify peaks in gasoline, it is necessary to set
retention times with at least a 2 s window-width. * The reference compound had to be elute free from other compounds and
behave at the same way as gasoline compounds. * Use of retention indices improves the reliability of the
identification. For a number of compounds it is possible to obtain accurate retention indices with a SD of 0.3 index units. Yet regularly situations were encountered in which a misinterpretation arose by too large a shift in the retention time of a particular compound.
* With the described equipment it is not possible to improve retention index measurements. A more accurate retention time registration is necessary. It was shown that a step size of 0.01 min is not enough to reach a SD below 0.3 index units. Improvement of the GC-equipment (temperature course stability) is less important.
* Coeluting compounds show a more irregular shift in retention than single eluting compounds. Pattern recognition techniques and/or control by GC-MS could be helpful to avoid serious mistakes in the identification and quantitation.
* Control runs with GC-MS on a regular base are recommended.
-66-
5. REFERENCES
(1) RIJKS, A. and CRAMERS, C.A. (1974) High Precision Capillary Gas Chromatography of Hydrocarbons. Chromatographia 7 (1974) 99-106
(2) JAESCHKE, A. and ROHRSCHNEIDER, L. (1972) Automatic Standard Gas Chromatographic Analysis with Retention Index Determination. Chromatographia 5 (1972) 333-340
(3) JANSSENS, G. (1977) Calculation of Retention Indices in Temperature Programmed Gas Chromatography by Improved Linear Interpolation. Analytica Chimica Acta, 1977
(4) ZELL, M., NEU, H.J., BALLSCHMITER, K. (1977) Identifizierung der PCB-Komponenten durch Retentionsindexvergleich nach Kapillar -Gaschromatographie Chemosphere 213 (1977) 69-76
(5) BETTY, K.R. and KARASEK, F.W. (1978) Application of Automatic Calculation of Kovats Retention Indices to Environmental Analyses by Gas Chromatography - Mass Spectrometry - Calculator. J. of Chromatography 166, 1978
(6) HALANG, W.A., LANGLAIS, R. and KUGLER, E. (1978) Cubic Spline Interpolation for the Calculation of Retention Indices in Temperature-Programmed Gas-Liquid Chromatography. Analytical Chemistry 50, (1978) 1829-1832
(7) LEE, M.L. and VAS SILAROS, D.L. (1979) Retention Indices for Programmed-Temperature Capillary-Column Gas Chromatography of Polycyclic Aromatic Hydrocarbons. Analytical Chemistry 51 (1979) 768-773
(8) PERRIGO, B.J. and PEEL, H.W. (1981) The Use of Retention Indices and Temperature-Programmed Gas Chromatography in Analytical Toxicology. J. of Chromatography Science 19, May 1981
(9) TUCHMAN a.o. (1984) Capillary Gas Chromatographic Separation of Urinary Acids. Retention Indices of 101 Urinary Acids on a 5% Phenylmethyl Silicone Capillar Column J. Chromatography Science 22 (1984) 198-202
(10) ENQUIST, J. SIMILA, P., LAKKISTO, U.-M (1983) Improved precision gas chromatographic method for retention index monitoring with a calibration detector (ref.4 point to the original work of Koskinen o.a.)
(11) RIEKKOLA, M.-L., MANNINEN, P. and MOYKKYMAKI, T. (1987) Effect of non-volatile by-products and injection techniques on retention indices in capillary gas chromatography. Chromatographia (1987) 352-359
(12) SPATH, HELMUTH, Spline-Algorithmen zur Konstruktion glatter Kurven und Flachen. Reihe Datenverarbeitung
(13) STICHTING REACTOR RESEARCH (1987) RRGRAPH Version 2.1 PC-software for drawing, smoothing and interpolating of functions. Delftse Uitgevers Maatschappij b.v. Delft, The Netherlands
-~-
- 89-
"1 H ""= 0
>- -BENZENE
CYCLOHEXEN
4-ME.CYCLO
3-ME.CYCLO TOLUENE (~_,
TOLUENE
l-HE.CYCLOHEXENE
>-l t'l Ul >-l
OCTANE :3: H :>< >-l c::: ~
(") I::) ~
(") 00 H Ul
Ul >-l N
~ i'd t'l I ;<l t'l
0 '"d :>< SE t'l ~ >-l
>-l ~ :3: ~ H 0 (")
Dl0-ETHYLBENZENE (
(") Z Z .., Ul (")
~ t'l 0 ... ETHYL BENZENE
0 0 'oj 'oj Ul
0 til t""' H
! P.XYLENE t'l ~ t""' Z N Q '"d t>l t>l
2ii H ;<l Z (")
STYRENE
<;"l 0
§ Z ~ 0 H
(") .., 0 "oJ H Z
?
(")
§ t'l ~ ~
.., t'l
t'l ;<l
NONANE
AN IMPROVED EXTRACTION ME1HOD FOR TIlE QUA!,,'TITATIVE ANALYSIS OF PESTICIDES IN WATER
Sunnnary
H. F. SCHOELER and J. BRODESSER Hygiene-Institute University of Bonn/FRG
A new method for the enri~hment of pesticide residues in water by liquid/liquid extraction with n-pentane in a light phase rotation perforator according to Ludwig is described. Altogether 44 compounds (chloropesticides, urones, triazines, phosphoric acid esters and others) were analyzed at given concentrations by gas chromatography. Four phosphoric acid esters and two other compounds showed insufficient recovery rates between 50% and less th~~ 10% due to their decomposibility in aqueous solutions. All other analyzed substances showed very good average recoveries of 80% to 115%. Detection limits ranged from 10 to 100 ng/l, dependent on the detectors (ECD. PND) and enrichment factor (1000 and 2000, respectively). The applicability of this enrichment method to residue analysis of pesticides in surface water and groundwater is demonstrated and its practicability is discussed.
Introduction In 1986 almost 30.000 tons of pesticides were used in FRG
(1). About 300 active agents are commercially available in a large variety of formulations and applications. Pesticide contamination due to persistence and high mobility in soil of several active agents was reported from USA, the Netherlands and FRG (2-6), which means a hazard for drinking water production. The difficulty of pesticide analysis stems from heterogeneity of active agents and from the limit value of 0.1 ug/l required by the German drinking water regulations. For pesticide analysis enrichment techniques are necessary, e.g. liquid/liquid extraction (7), adsorption on synthetic resins (8), reversed phase material, activated charcoal etc. (9,10).
Experimental To improve the liquid/liquid extractiClll, we used a rotary
perforator (according to Ludwig) and n-pentane as extracting solvent. Fig.l shows the apparatus which was first introduced into water analysis by KuEmaul (11).
- 69-
1 extraction vessel 2 condenser 3 solvent vessel
Fig.l: Rotati~n perforator (according to Ludwig)
125 ml n-pentane is added to 1 1 of water in the rotation perforator. The solvent in the solvent vessel is heated to evaporation, enters the condenser, is liquified and flows into the the rotating distributor. From there it is dispersed into the water as minute droplets, which percolate the \Vater phase. Thus t~e organic phase is continuously exchanged. After 30 min of extraction the solvent is dried on anhydrous sodium sulphate and concentrated to 0.5 ml by careful distillation through a Vigreux- column. The concentrated extracts are stored in a refrigp.rator at -20 t until analysis by gas chromatography is carried out.
Table 1: Gas chromatography parameters
GC Carlo Erba, Mega 5300 Column SE 54 fused silica
0.22 mm TD. 0.2 urn film, 25 m Iniection splitless 1 min then split 1:20.
1 ul (hnt needle) Carrier He 1,5 ml/min. Make up (PND): He 20 ml/min Make up (ECD): Ar/CH4 (95:5) 40 ml/min Temperature 60 t/2 min isothermal,
20 t/min to 2i 0 t, 5 min isothermal, 20 t/min to 260 t, 15 min isothermal
Integrator Hewlett-Packard, 3390A
During our investigations, different classes of compounds of pesticides (see table 2) were tested whether they could be analysed by r'ltary perforation. The compoQnds were mixed according to detectability and chromatographic separation. The mixtures were measured after dosage of the c:ompounds into water at various concentratinn~. To establish the detection limit. the gas chromatographic reproducibility (plus/minus 20%) at a given concentra-
-70-
I ;:!
I
Tab
le 1
: D
etec
tio
n l
imit
s an
d re
cove
IY r
ate
s o
f p
esti
cid
es f
rom
wat
er
by r
ota
tio
n p
erfo
rati
on
acc
ordi
ng t
o L
udw
ig
com
poun
d d
etec
tio
n
reco
very
co
mpo
und
det
ecti
on
li
mi t
(ng
/l)
(%)
lim
i t(n
g/l
) re
cove
ry
(%)
----
----
----
----
----
----
----
----
----
----
----
----
----
----
----
----
----
----
----
-P
hosp
hori
c ac
id e
sters
(P
ND
) T
riaz
ine,
Uro
ne
(PN
D)
Azi
npho
s-et
hy
l 10
89
S
imaz
ine
15
83
Dic
hlor
phos
10
69
A
traz
ine
15
89
Dim
etho
ate
.::.10
P
ropa
zine
15
98
Dis
ulf
oto
n
.:50
Ter
bu
tyla
zin
e 15
94
Etr
imph
os
10
85
Seb
uty
lazi
ne
15
95
Fen
itro
thio
n
10
94
De s
met
ryne
15
85
Fen
thio
n
10
86
Pro
met
ryne
15
89
Mal
athi
on
10
91
Ter
butr
yne
15
87
Mev
inph
os
,10
D
ipro
petr
yne
15
91
Par
ath
ion
-et
hy
l 10
83
M
onur
on
dO
Par
ath
ion
-m
ethy
l 10
90
C
hlo
rto
luro
n
50
60
Thi
omet
on
':50
Lin
uron
10
0 10
2 M
etab
enzt
hiaz
uron
30
91
Ch
loro
pes
tici
des
(E
CD)
Chl
oroo
rgan
ics
and
oth
ers
(EC
D,P
ND
)
DDT
35
115
Die
ldri
n
10
105
Ben
flu
rali
n
10
95
End
osul
fan
I 25
10
1 B
utr
alin
35
95
End
osul
fan
II
25
105
Dic
hlo
ben
il
25
86
Hep
tach
lor
10
104
Din
obut
on
35
104
Lin
dane
10
10
2 H
exac
hlor
oben
zene
10
10
2
Met
hoxy
chlo
r 25
11
6 M
etaz
achl
or
35
95
Qui
ntoc
ene
35
95
Met
olac
hlor
50
94
Tec
nace
ne
10
95
Nit
roth
al-i
sop
rop
yl
35
104
Tri
flu
rali
n
10
96
----
----
----
----
----
----
----
----
----
----
----
----
----
----
----
----
----
----
----
-n
ot
dete
rmin
ed
tion was chosen. The utilized measurement results are composed of entire analysis runs and of repeated chromatographic runs; statistical dropouts were eliminated. The aim of our investigation was to analyse a pesticide-doted water with respect to recovery and detection limits within the outlined method.
Results Tab. 2 shows the classes of compounds, recoveries and de
tection limits of which were investigated. Those substances marked by an arrow could not be detected or only with insufficient results. The average recovery of the phosphoric acid esters was between 85 and 95%. The detection limits were 10 ng/l using an enrichment factor of 2000. Dichlorphos was recovered up to 70%, thiometon and disulfoton up to 50%, dimethote and mevinphos less than 10%. Easily hydrolyzable and oxidation sensitive compounds could not be successfully enriched by rotary perforation due to chemical alterations. Monuron and chloridazon could not be analysed using rotary perforation. Chlortoluron, nitrothalisopropyl and butralin were recovered at 60%, dinobuton at 40%. Reasons for the insufficient recovery of monuron, chloridazon and chlortoluron seem to be moisture sensitivity, that of nitrothalisopropyl and butralin the photochemical decomposition of these coloured compounds. In a later series of analyses these photosensitive substances were analysed under light exclusion in the same run with chloropesticides and the recoveries increased to 90%. The detection limits were between 15 ng/l for the triazines and 100 ng/l for linurone using an enrichment factor of 2000. The chloropesticides were regained at rates between 83 and 104% . The mean recoveries for chloropesticides range between 90 and 110% in two cases - DDT and methoxychlor - above 110% and dichlobenil below 90%. The detection limits were 10 ng/l for the chloropesticides and between 25 and 50 ng/l for the so-called others.
________ ~ ____ _JA __________ _
Fig.2: Chromatograms of unpolluted surface water (on the left) and of polluted groundwater (on the right)
-72-
Surface water and groundwater s les e applIcatIon 0 t e presented method is demonstrated with
samples of a minor polluted surface water and a heavily polluted groundwater. The chromatograms of fig.2 were obtained using the same capillary columns (fused silica SE 54) and an electron capturing detector (top) and a phosphorous-nitrogen- detector (bottom). On the right fig.2 shows chromatograms of a surface water sample of the Sieger land, North-Rhine-Westfalia. The recordings of both detection modes were performed with maximal sensitivity ,but only minute peaks were found representing small pesticide residues. As contrast, on the right fig.2 shows a heavily polluted groundwater of a viticulture area near Bonn. Large amounts of pesticides could be identified e.g. atrazine 2000 ng/l, simazine and metolachlor 100 ng/l and parathion- ethyl 130 ng/l. Because of the pollutant concentrations the recordings were performed with reduced sensitivity.
Discussion A critical discussion of the outlined procedure comes to
the following assessment: 1) the vast span of the recoveries at low concentrations (near
10 ng/l) can be explained by process errors, at higher concentrations (near 100 ng/l) the deviation is around plus/ minus 10% .Reference to internal standardisation is excluded when analysing real samples since peak overlapping will lead interferences,
2) the detection limits can be lowered by increasing enrichment factors,
3) analysis of polluted water can lead to a plurality of chromatographic signals which needs further detection techniques (e.g. GC/MS) or cleanup or/and separation of the complex mixture (12,13),
4) use of an electron capturing detector demands highest solvent quality (for enrichment factors of 1000 or 2000, double distillation of n-pentane via an efficinet rectification column is necessary).
Conclusion~ The main advantages are: 1) fast exhaustive extraction in 30 minutes by a nonpolar solvent, 2) without solvent changing different types of GC-detectors can be utilised, 3) the extraction can be carried out within the sampling vessel by a simple modification of the rotation perforator. Loss by adsorption on glass walls can be avoided.
References 1) Industrieverband Pflanzenschutz: Jar,resbericht 86/87,
Frankfurt 1987
2) Zaki, M. H., MOran, D. and Harris, D.: Pesticides in Goundwater. The Aldicarb Story in Suffolk Country, N.Y., Am. J. Publ. Health 72 , 1391-1395 (1982)
-TI-
3) Loch, J. P. u. Hoekstra, R.: Spuren von Pflanzenbehandlungsmitteln im Grundwasser - Konzeption und erste Ergeb nisse von Untersuchungen in Boden hoher Durchlassigkeit in den Niederlanden. Schr.-Reihe Verein WaBoLu 68 , 247-264 (1987), Fischer Verlag, Stuttgart --
4) Friesel, P.: Grundwasserqualitatsbeeintrachtigungen durch Anwendung von Pflanzenschutz- und Schadlingsbekampfungsmitteln (PSM). Bundesgesundhbl. 29, 424-427 (1986)
5) Hurle, K., GieBl, H. u. Kirchhoff, J.: Ober das Vorkommen einiger ausgewahlter Pflanzenschutzmittel im Grundwasser. Schr.-Reihe Verein WaBoLu 68 , 169-190 (1987), Fischer Verlag, Stuttgart -
6) Oehmichen, U:u. Haberer, K.: Stickstoffherbizide Im Rhein. Vom Wasser 66 , 225-241 (1986)
7) Fachgruppe Wasserchemie : Deutsche Einheitsverfahren zur wasser-, Abwasser- und Schlammuntersuchung. Gas-Chromatographische Bestimmung von schwerfltichtigen Halogenkohlenwasserstoffen und Organochlorpestiziden in Wasser (F 2) (1985).
8) Oehmi chen , U., Karrenbrock, F. u. Haberer, K.: Determination of N- Pesticides in Natural Waters. Fres. Z. anal. Chern. 327, 715-719 (1987)
9) Sturm, R., Knauth, H. D., Reinhard, K. H. u. Gandras, J.: Chlorkohlenwasserstoff- Verteilung in Sediment en und Schwebstoffen der Elbe. Vom Wasser §Z ' 23-38 (1987)
10)Deutsche Forschungsgemeinschaft : Methodensammlung zur Rtickstandsanalytik von Pflanzenschutzmitteln. Band 1,2 und 3 (1985), Verlag Chemie
11)KuBmaul, H. u. Hagazi, M.: Die Bestimmung von Organofluorverbindungen in Wasser. Vom Wasser 48 , 143-154 (1977)
12)Xu, J., Lorenz, W., Pfister, G., Bahadir, M. u. Korte, F.: Residue Analysis of Triazine Herbicides in Soil. Fres. Z. anal. Chern. 375 , 377-380 (1987)
13)Specht, W. u. Tilkes, M.: Gas-chromatographische Bestimmung von Rtickstanden an Pflanzenschutzmitteln nach Clean-up tiber Gel-Chromatographie und Mini-Kiselgel-Saulen-Chromatographie Fres. Z. anal. Chern. 322 , 443-455 (1985)
-74 -
HYPHENATED METHODS aSp LC-MS, DLI LC-MS, L C-T1 D) ANALYZING ORGANOPHOSPHORUS PRIORITY POLLUTANTS
FOR
D. BARCELO and J. ALBAIGi:S Envirorrnenlal Chemislry DeJ.)artment, C.l.D., (C.S.l.C.) cl Jorge Girona Salgado 18-26, 08034 BARC~LONA (S PAIN)
SY!I!!1mY.
The coupling of several analytical techniques via appropriate interfaces for monitoring selected organophosphorus pestICideS (OPS) included in the EEC list of priority pollutants is muslrated. By utilizing positive and negative ion (PI and NI, respectively) "filament on" thennospray liquid chromatography-mass spectromelry (TSP LC-MS) and positive and negative chemical ionization (PCI and NCt, respectively) in direct liquid introduction (DLI) narrow-bore LC-MS, with acetonitrile-water mixtures, the characterization ol' several OPS is achieved. The use of 1% chloroacetonitrile in the LC mobile phase in DLI LC-MS is also discussed. Finally, reversed-phase narrow-bore LC coupled to a thermionic detector (LC-TID) is carried out with methanol-water mixtures as eluent. Applications are reported for the determination ofOPs in aquatiC environments.
1 .-INTRODUCTION Capillary GC with thermionic (TID) (1,2) or flame photometric detection (FPD) (3)
and capillary GC-MS in the electron impact (4,5) or in the chemical ionization, positive (PI) or negative (NO, modes with methane (5,7,8,9) and arnmonia(5,6) are rrequently the methods of choice for the trace level determination of orga~hosphorus pesticides (OPS) in environmental samples. However, It is known that OPs pesticides are difficult to analyse by GC (1,3,4,10). Some are thermally labile, such as trichlorfon (1,4), or produce failing, such as parathion (3). The temperature of the GC column should be optimized in the analysis of malathion in order to minimize bleeding of the liqUid phase (10). Because of these problems, the analysis of such compounds has been carried out by LC using UY (11), reductive amperometric (12), selective thermionic (13 ) or MS (14-15) detectors. In this work, the use of hyphenated methods, such as LC-TID and LC-MS will be discussed for the determination of nine OPs pesticides included in the EEC list of priority pollutants (761464ICEE. N umbers: 5,6,43, 73,80,89,97, 100,116). Further, a comparison of t~y (TS P) and direct liquid introduction (DLI) LC-MS will be shown. As an applICation, the determination of malathion in a fish sample is reported.
2.-THERMOSPRAY LIQUID CHROMATOGRAPHY-MASS SPECTROMETRY (TSP LC-MS.l --Tne on-line LC-MS combination using a TSP interface has been aPflied succesfully by several workers in recent years (16-24). It has been reported (16 that gas-phase ionization processes need to be considered in TSP LC-MS when compounds are eluted as neutrals. These processes will be different when filament-on or filament-off with PI or NI are used .Most of the papers published have studied the PI mode (16-21) and only few ones have proposed the NI (16,22-24).
In the work described here, the applicability of TSP LC-M5 with "filament on" in acetonitrile-water mixtures containing O.1M ammonium acetate using PI or NI is demonstrated for selected OPs pesticides of environmental concern. Because of a
-75 -
lad< in the Ileratlre, special ~haSis is dedicated to the negative ion formation in TSP LC-MS. V<>yksner et al., (19 have mentioned, for carbamate pesticides, that the sensitivity in PI is 4-5 orders magnitude better than for NI. ROCently, it has been pointed out that anion atlacl'lrnent and proton abstraction were the predominant mechanisms for conyounds such us corticosteroids and analogues of diethylstilbestrol (23). Proton abstractIOn has been identified in chlorsuflron (21).
In Fig. 1, the TS P-MS spectra for cot.rnaphos and drnethoate are shown.(M + N~]+ is the base peal< for both cOOf>Ounds and also for all the OPS analyzed, sinilarly as it has been previously reported in GC-MS using ammonia as reagent gas for azinphos-rnetI:lYl and fOSll'lel: oxon (6). Voyksner and Haney (18) using a TSP LC-MS interface for anat,'sing smilar compounds, e.g., diroothoate, azinphOs-met~1 and malathion, observed 1M + H]+ as base peak and also the formation of 1M + NH4] . In comast, our results always showed that the relative abundance (')f the 1M + (CH3CN)NH.t]+ rragrnent ion is more i'nportant than the auasinolecular 1M + Hl+ ion.
"'''''1' cou WAPHO S "0 (PU
-.-,) ;-°-f'Y"1°
'- ~-<-
'"
.~--~--~~--~
c . !--
- .• ..... ,or (Nil
· c
~-· · ·
.-
".~C"":: .J"'".I· , .. DIWE THO A TE
(PI)
c....~ ,' ~ I·-·-~C:-__ ·('"
,~.
(Nil
- -w ••• /e ... , • •
Fig l.Direct flw injection TSP PI MS 8/'ld TSP NI MS spectra of cownaphos 8I1d dimeth08le. Carrier stream :acetonitrile-'w'8ter(50:50) + 0.1 M NH.tMeC02; flo'w'rate: lmLImin
In Nt the formation of different ions is compound-dependent. Three different mechanisms ¥t'ere observed. Because ammonium acetate is needed for the thermospray ionization, the stronger acidity of acetic acid (348.5 kcaUrooO over ammonia will facilmte the formation of the 1M + Iv1eccnr adduct ion. This will be the first mechanism considered of ion formation Which is calTed anion atlachment (23). This adduct ion was the base peak for vamidothion, dimethoate, ornethoate and azinphos methyl. The predominance of such rragnent will be aHmuted to the fact that the gas-phase aCidly of acetic acid is lower than the acidity of these compounds, and so, frstly the deprotonation of acetic acid occurs with the formation of the MeC02 - adducts.
The second mechanism observed is electron capture wth the formatIOn of IMf. This was the base peak for coumaphos, temephos, paraoxon-methyl and malatJiion. Such compounds presumably have a higher acidity than acetic acid, and so the IMI is predominant. There is another fact to 00 considered which has been el'rl>hasized in GC-NCI MS (8). This is the stability of the IMr formed ion. lis higher intensly in GC-NCI MS for compounds such as coumaphos and paraoxon,with relative intensitieS of 1 00% and 2SO';'; respectively, was atlriblted to the aromatic rooiely which can eslabilize the IMf by eleclron delocalization. The aforementioned different behaviollS can also be
- 76 -
observed in Fig 1, where the NI TSP MS spectra for coumaphos and dimethoate are shown.
Th~ third mechanism in NI TSP ~C-MS is the ~issociative e~ctron c.a~ure. Such mechamsm corresponds to the formatIOn of IM-Rr in all the studied pestiCides, to the formation of (-S-CSHs) for coumaphos and to the formation of the group speaic fragment (ROhPS2 for phorate. The (M-Rr fragment was the base peal< for fenitrothion, metnidathion, parathion-metftyl and parathion-ethyl and had a value between 3-200/0 for the rest of the compounds.
As a test, an homogenate tissue of red mullet (Mullus barbatus) was spiked with 0.1 ppm of malathion. in Fig 2 is shown the LC-SIM MS PI traces of the fish ~Ie alter a pretreatment procedure derived from reUS. The TSP LC-MS with PI in the single ion monitoring is seen to be sensitive and selective . The sensitivity in PI was much better than in NI and it was especially noticed for the noise level, which was at least 30 times higher in NI than in PI. Considering the amount injected (sng of malathion) and tfle signal obtained, the PI mode exhibits high selectiv~ With practically no influence Of the matrix and will be suitable for environmental analysis.
.. 0 c .. " c , D <
5.0e ..
2.0£ 4
THERMQSPRAY SIM (Pll
MuIIU$ barba.us
soik ed 0 , 1 ppm Malafhi on
IM+(CH,CN)NH, 1+ .~
t---"--------"'------------'--,,' ~ .. o~-.--._-~-~-~-~-~--
• Time (min)
,. ' 2 "
Fig 2. TSP lC-SIM PI MS chrOl1l8lolJl'8l'll of 8. fi3h 38/Tlple 3pilced 'with 0.1 ppm of malathion. njected amount,Sng. Eluent,acetoririle-O.l M~MeC02 prognmned from 50 to 80% ecetonitrilein 15m at lmllmin. Therm03pr8.y temperature:!: :!tem, 105-1 f4'C; tip, 178 'C; 'vapor, 194 'C; ion 30urce, 2&4 'C.
3.-DIRECT LIOUID INTRODUCTION LIOUID CHROMATOGRAPHY-tMSS SPECTROMETRY (DLI LC-MSl
The on-lire combination of LC-MS wlh the di'ect liquid introduction (DLI) of 10-20 IlUmin of the LC effluent into the ion source has recently been applied (15,25), providing molecular \¥eight or functional group information. As regards 'the trace level determination of OPS ~ides, it has been reported that NCI is more sensitive than PCI (26). This is ~bly due to the formation ~ more stable negative ions by the low energy electrons in the MS source. Recently, Dougherty et al (27) have successfully applied PCI, NCI and chloride-atlaclTnent NCI to di'ect probe a1is.
The applicabilily of on-lire LC-MS wth PCI and NCI to the OPs pesticide ana is ¥taS demonstrated by Parker and co-'nOO<ers (14,26). Later, this (28) and a her gro~ (15) Yr'ef'e the frst to use acetonlrile-Vtater cotiaining 1% chlor'oacetonlrile as mobile phase for obtaining chloride-atlaclTnent NCI. Such cr atlaclTnent has staTered little systematic inveStigation in LC-MS. Only ore previous exaJ1l>1e of cardiac glycosides is reported in the literatll'e by eJll)loying a moVing bel: interface (29)
With PCI the (M+Hj+ion ¥taS formed by all pesticides, and this ion usual~ ¥taS the base ~aJ<. In addition, for some of the ~icides, a rather intense (M-FGJ ¥tas obtained {where FG corresponds to (RO)2PSX), This fragment, ¥taS the baSe peak for
-77-
carbophenothion, azi~hos-methyl and azinphos-ethyl and corresponds to a resonance stabilized species prevlOSuly reported for OC-PCI MS (6). In the NCII'YlOde (without chloroacetonilrile) the mass spectra of the OPs ~e characterized by an Intense [FOr ion which generally was the base peale Exceptions to this rule are trichlorfon, that sho'HS a particular rragmertation, and the phosphorothionates that have an aromatic ring bonded to the functional gro~, such as£rathion. In this case, the thiophenoxide ion is the most abundant ion as was ak"ea cled in OC-NCI MS (1,8). FlIther, the relative intensil:j of [M-Rr (With R being met I or ethyQ was in all cases lower than 4%, its intensly sliongly decreased With decreasing the source t~ttre.
TR ICHI.ORF~ ....AM l oor HION
100 .. ,
I .... · 1 ~Ol zPOj" 100 ... [IROlzPSOr
..,c, ... c'
50
[ ... ·c'· _1 •••
".[ ... . RI ( .... - Rj"
'" M/z 150 200 250 300 350 M/z 150 200 250 300 350 1;00
100 , ..
[ ... ·C'-Rr 100
["'-R!. ,"'CI , ..... c.
~-flr
50 50
~·1!a2,.q-
r ~,cf '01 •
(iROJ2PSOj- ~N'1
'" '"~
M/Z 150 200 250 JOO 350 M/Z 150 200 250 300 350 1;00
Fig 3.Aw iiecti~ MS and CLEt-I mass spectra of trichlufon end wmidothion. Ion source tempmllre, 300 tC; CMier stretvn, acetonitrile-wer (10:30) for NCI or lICetonirie-'-'8ler1:hlorOllCetcririe (69:30: 1) forCLEt-I; ftO'w'1'lllte, 20 tAlm.
Evident changes in the NCI spectra ~e observed l4>On addlion of 1 % chloroacetonilrile to the eluent, thus the term chloride enhanced negative ionization (CLENI) is used. The CLENI process resuted in the loss of R from the ~ides to 9iYe (M-Rr ions as base peak, while also for most pesticides a relatively Intense [FOr IOn could be observed. In addlion, chloride-allactment was rather modest at a source temperature of 3000C for di'nethoate, vamidothion, trichlorfon and ronnel. The (M+Cl)"peak reached percertages beh¥een 1 and 8 %. In Fig. 3 the spectra of frichlOOon and vamidothion obtained 'Nth NCI and CLENI are given.
The LC reconstructed ion current (RIC) clTomatogams in PCI, NCI and CLENI modes for a mixture of 10 organophoSphOrus pesticKies are shown in Fig. 4; the presence of chloroacetonlrile did not artect the retention tine or peal< shape of the ana~. (l)viously, NCI is much more senslive than PCI for this type of c~unds (1,8). The senslivly in the NCI mode was 10-20 tines better than in the PCI mode for ca.rbophenothion, ronnel, azi'1>hos-rnethy1 and azirphos-ethyl and about 4 tines better for dmethoate, paraoxon-methyl, parathion-metnyl, trichforfon and cotlTlaphos. WI:h vamidothion, the senslivly in the tfC11'YlOde was l'M> tines YIa'Se than in the PCI mode. This feature has also been observed for other ~nens which have a -t-(H)CH3 gro~ (e.g. triazine herbicides, methyl-urea and carbamate pesticides) (26). . ~ c~ng the lC-HCI MS and LC-ClE:NI MS traces in Fig.4 one notices that for har of the cOOlX>unds the senslivly is si'nilar in both ionization modes and for
-78 -
azinphos-methyl, azinphos-ethyl, coumaphos, ronnel and carbophenothion there is a loss in sensitivity by a factor of 4-8 of CLENI compared to Nt!. Detection limits, in LC-NCI MS, were calculated to be SO pg (SIN=3), as in GC-NCI MS (7).whi<:h makes that technique suitable for trace analysis with signifi<:ant advantage of an easier sample clean-up.
100 4
9 6 I 7
8 I I ,
3 \ 10
U~, 4
l::! IS 0: \2
8
10 PCI
';00 500 SCAN I!>;OO 20:00 25:00 TIM E (mi")
Fig 4Reconslructed ion current chromatograms in LC-PCI MS, LCID MS and LC-CLENI MS for a rro:ture cI (1) vemidothion, (2) trichiOlfon, (3) dinethoele, (4) pnOXOI'HTletlr,.t, (5) azi'lphos-metlr,.t, (6) ~hiOl'HTletlr,.t, (7) azinphos-etlr,.t, (8) cOll1l8.phos, (9) romel and (10) carbophenothion. Mlolri of each component injected, 50 ng; eluent, acetonitrile-'w'lller (70:30) for Pel and NO or acetonitrie-'tr1ller-chloroacetonitrile (69:30: 1) forCLENI al20 ..... Imin; ion sOll'Ce temperal:lI"e, 300 <c.
4.-0N-LINE THE~ONIC DETECTOR FOR LOUID CHROMATOORAPH,( (LClTIDl The use of selective GC detectors in LC has received increasing attertion owing
to the g-owing need for high detection senstivly and selectivly in LC. Urtortunately, the cortinous loo-oduction of aqueous and organIC solveris at mllmin flow-rates, typical of convertional LC, has caused numerous technoi09ical problems. Miniattrization in LC has elminated many of the dilflCuties associated wth drect mobile phase ioo-oduction, as is very lOw fIow-rates cause only minor detection disttrbances.The ltilization of electron capttre (30,31), flame emission (32) and thermionic (13,33) detectors for microcolllTln LC has been descrt>ed previously.
In this investigation, a thermionic detector, based on a design of Kob and Bischoff (34), was ~Ioyed as an on-line detection system for narroW-bore LC. The whole system consists (j an evaporation interface, located in the detector block, and mairtlined at 300 °C. The termination of the LC colllTln is drectly connected to the
-79 -
2
A B
J
J 2
'" o d
~ a:) ..
o " • L-o > " ::>
l
Fig S.lC trace ot tomato eldract spiked 'with pesticide standards as monitored by (Aj the lC-TD system (8) UV absorbence at 254 nm. Cohllln, gIIm-lned stainless steel (200 x 0.7 mm 1.0.) packed 'wih 5.,m liChrosorb RP-18; mobile phase, melhanol-'water (80:20) at a flow rate of 40 ~Imin. Solutes: (1) trich/orfon; (2) azinphos-elh)t; (3) cOllllaphos (30-50 ng ot compound i'ljecled)
15 cm x 0 .25 rnm 1.0. stainless-steel interface and the vaporized eluent from the interface is introduced into the flame jet via a 1 Scm x 0 .12 mm J.D. fused silica capil~. A nir<>gen flow of 3 mllmin is added via a T piece just below the detector bodY. ThIS gas was pre-heated in the GC oven, in order to minimize its cooling effect. As an application, Fig. 5 shows the LC chromatograms of a tomato extract spiked with 3-5 ppm of trichlorfon, azi~hos-ethyl and coumaphos. The pesticides were separated by reversed-.ehase LC and detected with the LC-TID system (A) and UV absorba~e at 254 nm (8). The selectivity of the LC-TID system was calculated to be 1. 10 9 of carbon per gram of phosphorus. The LC-TID chromatogram shoVY'S a high selectivity for the phosphorus-containing pesticideS and a striking contrast with the complex traces obtained by the UV absOrbance, where the absOrptivity is otten low. Indeed, this selectivi~ allows simpler clean-up procedures to be used.
5 .-CONCL US IONS The relative merits of three hyphenated methods such as TSP LC-MS, DLI
LC-MS and LC-TID have been discussed with OPS ~icides as test compounds in fish, sediment and food matrices. The combination of PI and NI in TS P and the use of PCI, NCI and CLENI in DLI provided u~uivocal information abolt the OPS. For trace determinations of the OPS pesticides, PI should be preferred to NI, in TS P, and NCI and CLENI to PCI, in DLI. Pic~ levels can be detected as in GC-NCI MS but 'lYith simpler clean-up procedures. Because of the numerous cluster ions in the reagent gas spectrum, TSP LC-MS needs to start scanning from higher mlz values than in DLI (~ 50 versus 100), although this will not be a problem due to the. fact that in TS P with PI the (M+N~r is always the base ~al< for the OPS. Moroever, in DLI the base peal< with CLENI IS (M-Rr as against the functional group with NC!, so the fOlTl'ler mode has the advantage to provide more selective information and alloVY'S to distinguish ~n the different groups of pesticides.The LC-TID system sho~d detection Imits of about 40 P9 for OPS USing a flow rate of 4D}lLlmin with methanol:water mixtures. The highest eluent flow-rate that can be handled by the LC-TID system is 70 JiLlmin.
- 80 -
ACKNOWtEDGEMENTS We than< Prof. rx. R. W. Frei (Free University at Amsterdam, NL) for his advice
and encolngemert. The work ¥ras partiallY s~oo by funds from the CS IC project 6031070-40 and by a NAT.O. fellowship (b.BatcelO).
REFERENCES
! 1~! Prisnloo, SM.; De Beer, P.R. J. Assoc. Off. Anal. Chern. 1985, 68, 1100. KjillhoR:, J. J. CIYomatQ9L 1985, 325. 231. Zenon-Roland, L.; Agneessens, R.; Nanginot P.; Jacobs, H. J. Htgh Resoht. ChrornatQg[. ChromatQg[. Corrrnun. ~ L 480.
(45) Wil<ins, J.P.G.; Hill. A.R.C.; Lee, OJ. Ana~ 1 985.illL 1045. () Sil!9h, A.K.; He~n, OW.; Jordon, K.C.; Astrat M.; J. ClYomat~ 1986.
369,83. @ HOinstead, R.L.; Casida.J,E. J. Assoc. Off. Anal. Chern. 1974.57 ,J05O-1055. (7) Stan, HJ.; Kellner, G. in "Recent Advances in Food Ana~is· ;ootoo by W.
Baltes, P.B. Czodl<-Eysenberg and W. Pia nnha user ; Verlag Chernie: Weinhem, Foo, 1981 ;183-189.
8) Stan, H.J.; Kellner, G. B~. Mass S~ctrom. 1982. i 483. 9) Stout S J.; Steller, W A. Biomed. Mass ~om. 1984 .11. 207. 10j Hanilf, I.M.; Zienius, R.H. J. ClYomatQgL 1983, 264. 33. 11 Osselton, M.D.; S 00 iii ng, R.D. J. ClYomatQgL 1986. 368, 265. 12 CIar1<, G.J.; Goodin R.R.; Smiley JW. Anal. Chern. 1985. g 2223. 13 Gluckman, J.C.; BarcelO, D.; De Jong, G.J.; Frei, RW.; Maris, FA.; Brinkman, UA
Th. J. ClYomatQgL 1986. 367, 35. (14) Par1<er, C.E.; Haney, CA; Hass, J.R. J. ClYomat~ 1982, 237, 233. (15) BarcelO, D.; Maris, FA.; Geerdin<, R.B.; Frei, R.w.; De Jong, G.J.; Brinkman,
UATh. J. ClYomat~ 1987 , 394. 65. (16) B~, MM.; Par1<er, C.E.; Smth, R.W.; Gaskell, S .J. Anal. Chern. 1985,57,
2597.
1171' Alexander, A.J.; Kooarle, P. Anal. Chern. 1986, ~ 471. 18 Voyksner, R.D.; Haney, CA. Anal. Chern. 1985, 57, 991. 19 Voyksner,R.D.; BlI'SeY, J.T.; Pellizzari, E.D. Anal. Chern. 1984, 56, 1507. 20 Covey, T.R.; Cro¥tther, J.B.; De~y. E.A.; Henion, J.D. Anal. Chem. 1985. 57,
474. (21) Fenselau, C.; liberato, D.J.; Yergey, JA.; Colter, R.J.; Yergey, A.L. Anal. Chern.
1984. 1§.. 2759. (22) McFadden, WH.; Lammert SA. J. ClYornatQ.gL 1987, 385,)01. (23) Parker, C.E.; Smith, RW., Gaskell, S .J.; Bursey, MM. Anal. Chern. 1986. 58,
1661. (24) Smith, R.W.; Parker, C.E.; Johnson, DM.; Bursey, M.M. J. ClYornatQgL 1987, 394.
261.
1251 Covey, T.R.; Lee, E.D.; Bruins, A.P.; Henion, J.D. Anal Chern. 1986. ~ 1451A. 26 Parker, C.E.; Haney, CA; Harva.n, D.J.; Hass, J.R. J. ClYornatV-' 1982, 242, 77. 27 Dougherty, R.C.; Wander, J.D. Biomed. Mass s~om. 1980. ~ 401. 28 Par1<er, C.E.; YamagUChi. K.; Harvan, D.J.; Smit ,R.W.; Hass. J.R. J. ChromatQgL
1985, 319. 273. (29) Levsen, K.; Schafer, K.H.; Dobberstein, P. Biomed. Mass Spectrom. 1984. 11.
308 ..
1301 Brin<man, UA.Th.; Geerdink, R.B.; De Kok, A. J. ChromatQ9[. 1984. 291. 195. 31 Maris, ~.A.; Geerdinl<, R.B.; Brinkman, UATh. J. ClYomatQgL 1985, 328. 93. 32 McGl1fm, V.L.; Novotny, MY Anal. Chem. 1981. g 946. 33 BarcelO, D.; Maris, FA; Frei, R.W.; De Jong, G.J.; Brin<man, UATh. Intern. J.
Environ. Anal. Chern. 1987, ~ 95. (34) Kol?, B.; Bisc~, J.; J. ClYomatQg(. Sci. 1974 . .1£ 625. (35) MUir, D.C.G.; Grift, N.P.; Solomon, J. J. Assoc. Otf. Anal. Chem. 1981. §i 79.
- 81-
ANNEX
Slructll"es of the OPS ~ides cted in the text
Q2mP-Qunds Formulas R R' MoI.w..
P~hates, (RO)i)()2R'
Paraoxon-methyl CH3 C6~N02 247
P~honates. (ROhPOR' Trichlorfon CH3 C2H~CI3 2S6
P~horothionates. (RO)2PSOR' Parathion-methyl CH3 C6~N02 263 Fenilrothion CH3 C7H6N02 277 Parathion-ethyl C2HS C6~N02 291 Ronnel CH3 C6H2CI3 320 Coumaphos C2HS C10H602C1 362 Temephos CH3 C14H1403PS2 466
Pho§P-horothiolates, (ROhPOSR' omethoate CH3 C3H6NO 213 Vamidothion CH3 C6H12NOS 2S7
P~horodthioates, (RO)2PS2R' Oinethoate CH3 C3H6NO 229 Phorate C2HS C3H7S 260 Methidathion CH3 C4HSN~2S 302 Azinphos-methyl CH3 CSH6N30 317 Malathion CH3 CSH1JD4 330 Garbophenothion C2HS C7H6C1S 342 Azinphos-ethyl C2HS CSH6N30 34S
- 82-
PCB I S AND ORGANCCHIDRINE PESTICIDES ill EEL AND FIDUNDER IN THE TAGOS ESTUARY
Surrrncn:y
M.J.BENOLIEL and M.L.SHIRLEY Instituto Hidrogra:Eico,Rua das Trinas, 49
1296 LISBOA CODEX PORI'UGAL
Within the Tagus estuary, in the vicinity of an industrial zone (chloroalkaly, pesticides, fertilizers and steelworks plants) a IIDni taring programne on Organochlorine Pesticides and Polychlorinated Biphenyls has been carried out since 1983. This zone is influenced by agriculture activities as well. Flounder (Platichthys flesus) and eel (Anguilla anguilla) have been sampled b-Jice a year. lInalysis have l:een carried out both on liver and fillet. lDiI concentrations were observed which present neither risk for survival and normal development of living species nor to public health.
1. INTRODtx:TICN Determination of organochlorine pesticides and polychlorinated biphe
nyls has l:een included in different national environmental prograrrrnes once they were found to l:e toxic and persistent.
C..as chromatographic detennination of PCB I s in all kind of samples was acc~lished by means of packed colurrrns and auantification was made using the technical ARCCLOR 1254 or 1260 mixtures as standards. The p::x::>r match between the peak patterns fran the sample and the technical standard rnixb.rres has been frequently noted in literature (1, 2). The application of capillary column provided the individual congeners analysis (3) and the
need for :iJrproving and ccmplement the studies already carried out became a~t.
Hydrographic Insti tate is carrying out a monitoring prograrnne in coastal and estuarine waters which has been included in the Paris and Oslo Con ventions Joint Monitoring Prograrrrne.
The river Tagus is the most :iJrportant river of the Iberian Peninsula i it enters the Atlantic Ocean in front ~f Lisbon fonning the largest estuary of West Europe with an area of 320 Km approx:iJllately and a length of 80 Km subject to tidal influence. It is a partly mixed estuary characterized by an eno:rmous volume (2 x 109 m3), strong dynamics, a great tidal prism (750 x 103 m3) and a rather long flushing time. Upstream lies an area of i~ lets and shoals furraNed by deepe<r courses usualy with Im.1ddy bottans and surrounded by areas where a strong farming activity takes place. Inasmuch :iJrportant activities which include chemical and petrochemical ccmplexes, steelworks and shipbuilding are located in the river banks, Tagus estuary has been pinpointed as o:ne of the priority areas.
The sampling zone selected for this study is located in the riaht bank, in a channel with a mean depth of 2 meters and a lenght of 6 kilareters. Brewery,steelwork, fertilizer, pesticides and paints plants sites are close and there is a direct influence of effluents fran a chloroalkaly plant. For
- 83-
the purpose of this study eel (Anguilla anguilla) and flounder (Platichthys flesus) were selected considering their a::mrercial exploitation as well as their importance as pollution indicators and their representativeness of the area.
38'S3 '
38'~8 '
38' ~3 '
~ N
0 q,
o, , ~
_ S .... PllNG S TATIONS
"'CH LOROAl~"lY PL ANT
38' 3~~ ~2e-,---g-'~I-S -", ---g'-I~e-' -"--"",-L-g--!,""s-' L.----g~'----8'~5-5-' ---8--1, 50'
Long i tude
2. PROCEDURE
r .. .. .. c D-It
'!he sanples consisted of 25 fishes of each specie. The procedure for sanpling was devised taking into aCCOtll'lt the ICES Guidelines (4). Each fish was individually observed and prior to freezing total weight, total lenght, liver weight and sex have been established (Table 1).
MTE I N° TOTAL LENGTH (em) WEIGHT (g) LIVER WE IGHT (g) SEX I " 111-*)('-:0 " 111-*)('-:0 " 111-*)('-:0 " FEMALE
8}.11.0} I II 2 •• }-}2.0-28.8 I 150. }O-}47.80-2}().5O 0.78-07.60-}.51 n 84.0}.08 I 25 2O.0-}7 .0-29. 7 I 86.21~526.20-271.6O 0 .~18.8b-6.95 75
FlOOlllER 84.10.20 I 25 18. 5-}4. 5-2}.0 I 76. 7}-412. }5-185.10 85.0}.04 I 25 21.5-}4.5-26.0 I 96 .8}-49'J. 36-199.90 0.67-15.7}-U6 60 85. 11.29 I 25 22. 0-}2. 0-25.6 I 10}.09-426.50-205.82 0.57-10.OH.22 110 86.01.2. I 25 20. 0-29. 0-2}. 7 76.50-)1 •• 49-160.04 0.75-07.20-).07 60 86.10.2) I 25 20.2-26.6-22.6 84.01-25O.}o-1 }().09 60 84.11.21 I 25 29.0-38.0-}1.2 }2. }I-GO. 78-041.10 0.46-2. }O-O.91 85.04.10 I 25 16.0-22.0-19.2 }.50-016.06-009.67 O. IM-O. 20-0. I )
EEl 85.11 .29 I 25 17.8-27.2-22.7 9. I 5-O}().88-019. 56 0.17-0.60-0.38 86.01.2} I 2. 25.2-}5.7-}().2 27. }5-0n.I8-047. 54 0.39-1.48-0.86 86.10.2} I 2} 2).0-,1.0-26.4 17. I 0-06U I-O)}. 78 O. )}-I. ",-0.61
TABlE I - DESCRIPTICII (J' TIlE SMPlES.
-84-
Sarrples ~re obtained twice a year; one before spawning (January/March) and the bther afterwords (October/Novernber). Pooled sarrples for analysis have been taken. Liver (5.00 g) and fillet (10.00 g) of fish ~re extracted for 5 hours with n-hexane, (~rck 4371) in a soxhlet apparatus. Aliquots of the extract were cleaned up using colunns (20 mn ID) of anhydrous sodium sulphate over 15 g of alunina basic ~ck 1076) deactivated with 5% H20. The organochlorine catq:lOunds ~re eluted with 160 ml n-hexane. 'Ihe eluates have been concentrated to 2 ml and fractionated on 6 mm ID silica columns ( 2 g Si60, 25-40 pm ~~rck 9390), activated 16 h at 195 ex: before use. The first fraction was eluted with 13.5 ml n-hexane and contained PCB's, pentachlorobenzene, hexachlorobenzene and a percentage of 44' ODE. The second fraction (13.5 ml n-hexane/diethylether 85/15) contains the remainder 44' DOE, 44' DOD, 44' DIYl', (f- HCH and dieldrin.
Detennination of the chlorinated carqx>unds was perfonred by using Hewlett Packard gas chranatographs with electron capture detector (ECD).
']Wo capillary colunns ~re used: WOOT SP 2100 (25 m x 0.24 rom x 0.20 pm) : carrier gas: helium, flew 1. 2 ml/min splitter closing time: 0.60 min Terrperatures: injector .. 250 ex:
column - 1 min. .. 50 ex: 50 ex: to 215 ex: - 15 ex:/min 15 min .. 215 ex:
WOOT SE 54 (50 m x 0.20 mn x 0.33 pm): carrier gas: helium, flew 1.8 ml/rriin. splitter closing t:iJne: 0.60 min. temperatures: injector - 270 ex:
column .. 1 min. .... 50 ex: 50 ex: to 175 ex: .. 8 9CJlnin. 175 ex: to 270 ex: .. 1 9C/rnin.
No rooaningful differences were obtained in pesticides nor in PCB's contents determined by both the columns. Neverthless with SP 2100 column there is no possibility to separate congeners 28 + 31.
Blank determinations ~re perfonned for all canponents during each group of analysis. Opt:imization of gas chranatographic conaitions has been carried out (5). Recovery of the nethod is better than 85% for pesticides and 95% for PCB's.
Fat contents were establisred by neans of ~ighing an extract aliquot after evaporation.
3. RESULTS AND DISCUSSICN The quantitative results are carq:>iled in tables II an<LIII. For the PCB's quantification 13 congeners were used, their contrilJu....
tion to the total arrount of PCB's being about 40 .. 50 %. Neverthless special attention was focused on the so-called indicator congeners IUPAC N9 28, 52, 101, 138 and 153 (6). It can be estilnated that the use of these five congeners represent a decrease of 7 - 10 % with relation to the thirteen PCB's.
PCB's congeners in eel presented low values and no :inp:lrtant variations along the considered period. In eel liver the 153 congener shews a higher accumulation (42 .. 125 pg/Kg ~t ~ight) than the 138 one (13 .. 40 w/Kg ~t weight). Ratio 1381153 is about 0.30 in eel liver and about 0.70 in eel fillet. The highest PCB's values were found in flCJllll4er liver due perhaps to its high fat content. In flounder liver PCB's 101; 138 and 153 presented a decrease of five times in their concentrations during the period 1984 ":" - 1986. PCB's 28 and 52 shew no significant variations during the same period.
- 85-
I IUPAC N° I I I SNl'tE MTE " D.W. I " FAT 28 52 101 153 138 180 I 15 .... e£~·L·-1.!.n PCS\_·· .... I
84.11.21 26.0 I 5.18 1.2 I.} 6.9 86 52 -I 147 I 176 EEl 85.04.10 22.9 I 2.81 0.5} 0.94 4.6 38 }4 -I 18 I 90
FILLET 85.11.29 2}.7 I 3.03 0.41 0.40 1.9 34 28 -I 64 I n 86.01.23 24.4 5.39 4.9 3.4 7.5 123 82 41 I 221 I 286 86.10.2~ 2~.6 4.68 2.0 2.1 4.6 40 ~4 2~ I 83 124 84.11.21 5.33 1.2 0.91 " 125 40 -- I 118 199
EEl 85.04.10 1.98 0.84 0.52 2.6 70 20 -I 94 99 LIVER 85.11.29 2.56 0.60 0.50 3.5 88 27 -I 120 127
86.01.23 2.64 2.9 2.0 5.8 109 38 18 I 158 186 86.10.23 3.28 2.0 2.0 2.6 42 I~ 7.9 I 62 18 8}.11.0} 8.17 0.38 0.81 0.42 4.0 14 14 -I }O 40 84.0}.08 17.} 0.40 1.6 0.66 5.4 15 14 -I H 51
I FlOOlllERI 84.10.20 n.1 0.46 0.61 0.21 }.6 12 10 -I 26 }6 I FILLET I 85.0}.04 8.6 0.14 1.5 0.65 5.5 18 16 -I 42 52 I 185.".29 20.7 0.60 0.66 D.&} 1.7 5.4 5.0 2.0 I n.4 19.5 I I 86.01.24 20.2 0.}2 1.8 1.0 1.3 }.4 }. } 1.8 I 10.8 16.1 I I 86.10.2} 19.1 0.39 0.51 0.40 1.4 ~.4 ~.6 1.7 I 9.} I~.~ I 18}.11.0} 15.7 }.4O 5.8 }.6 54 297 268 -I 628 716 IFlOONOERI 84.0}.08 I 2}.1 8.38 }7 16 I 120 H6 }42 -I 891 10}4 I LIVER I 85.0}.04 I 28.2 5.95 21 14 II} }18 }06 -I 772 964 I I 85.11.29 I }1.9 14.9 28 17 57 215 167 891 484 701 I I 86.01.24 I 26.7 9.4} 27 16 I 2} 77 72 }9 I 215 517
• I 5 PC8's: 28. 52. 101 15}. 138. II ~I} PCB's: 28.52.49.44. 101.97.87. 151. 15}. 141. 138. 128. 180.
IUPAC N° STRUCTURE 28 2.4.4' 44 2.2' .}.5· 49 2.2' .4.5' 52 2.2' .5.5'
101 2.2' .4.5.5' 97 2.2'.r .4.5 87 2.2' .'.4.5'
151 2.2'. }.5.5' .6 15} 2.2' .4.4' .5.5' 141 2.2' .}.4.5.5' 138 2.2' .3.4.4' .5' 128 2.2' .}.}' .4.4' 180 2.2' .}.4.4' .5.5'
TABLE" - PCB's residues in eel (Angui "a anoui IIa) and flounder (Platichthys flesus), expressed in pG'Ko on product base.
44 I DDT levels were less than 10 % of the total DDT (44 I DDE + 44 I
DDD + 44' DDT) probably due to the fact that in Portugal DDT has not been used since 1970. Ratios of 44' DDTltDDT were found to be smaller in liver than in fillet both in eel and flounder. 44' DDT, 44' DDD and 44' DDE levels observed in eel fillet were about twice higher before spawning (January / /March) than afterwards. There is no evident trend as far as flounder tissues and eel are concerned, Levels observed in flounder liver were ccmparatively higher TIBinly as a result of the high fat content,
Dieldrin in eel fillet reaches its highest concentrations during ,Tanuary;March. Flounder and eel liver don't reveal any evident trend.
Lindane values were lew and similar in both species under study.
- 86-
I DATE " D.W. "FAT I 5 CBl 6 CBl 1: HCH DIELD «'DOE «'000 «'OOT I J. OOT I 24'OOT I I 84.11.21 26.0 5.18 I < I l.4 l.1 n 46 26 2.1 I 15 I < I I
EEl I 85.04.10 22.9 2.BI I 0.41 4.9· 2.B 16 49 24 4.l I 11 I 2.0 I FILLET I B5.1I.29 2l.1 l.Ol I O.~ 1.4 2.2 B.6 30 12 0.86 I 4l 2.1 I
I 86.01.2l 24.4 5.59 I 0.&2 l.1 5.1 11 &1 26 4.1 I 98 0.41 I I 86.10.2~ 25.6 4.68 I 0.~1 0.94 I.~ 5.4 l2 ~.9 1.2 I ~ <I I I 84.11.21 5.n I < I I.B < I 6.l IB 5.5 <l I 24 <I I I B5.04.10 1.98 I O.ll I.l 1.6 4.6 20 l.B < 0.6 24 < 0.5 I
EEl I 85.11.29 2.56 I O.ll I.l 2.0 5.9 15 4.2 < 0.6 19 < 0.5 I LIVER I 86.01.2l 2.64 I 0.11 1.8 4.6 9.1 18 1& < 0.6 54 2.4 I
I 86.10.2} 1.28 I 0.24 0.18 0.B1 2.1 10 2.8 < 0.6 I} I I BI.II.0l B.11 0.18 < O.ll 0.59 0.40 0.51 &.9 2.1 0.11 9.1 < 0.1 I I 84.01.08 11.1 0.40 < 0.40 0.92 0.58 2.B 1.1 4.B 1.1 15.5 < 0.1 I
flOUNDER I 84.10.20 Il.1 0.46 0/ 0.40 0.48 < 0.40 < 0.80 9.2 2.B 0.95 12.9 < 0.1 I FILlET I 85.05.04 8.6 0.14 < 0.40 0.94 < 0.40 1.1 6.1 l.O 1.4 10.1 0.56 I
I B5.11.29 20.1 0.60 o.n 0.42 0.42 1.6 4.9 2.1 1.2 8.8 0.64 I I 86.01.24 20.2 0.52 0.12 0.21 0.16 0.52 5.0 0.12 0.64 4.4 0.22 I I 86.10.25 19.1 0.19 0.094 0.24 0.2l 1.0 1.5 0.98 0.88 }.4 0.24 I I BI.II.05 15.1 5.40 < 1.0 4.1 30 9.0 91 49 < 5 146 < 5 I 84.05.08 25.1 B.18 < 1.0 l4 < 2.0 56 158 B9 14 261 24
FlOUNllER I B5.05.04 28.2 5.95 1.0 6.8 l.2 20 95 40 2.8 IlB 5.2 LIVER I B5. I 1.29 51.9 14.9 4.0 15 4.8 26 9l 66 B.1 168 10
I 86.01.24 26.1 9.45 1.9 B.1 2.4 14 61 15 1.0 85 1.0
ZOOT: "'DOE t «'DIJO t WOOT.
5 CBz - Pentachlorobenzene 6 caz - Hexach lorobenzene rHCH - lI' Hexachloro cvclohexane (I indane) DIElD - Dieldrine DOE - DichlorCHliphenvl-ethvlene 000 - DichlorCHliphenvl-ethene OOT - DichlorCHli pIlenv I-tr ichloro-ethene
TABLE 1.11 - Organochlorine pesticides residues in eel (Angui 110 angui 110) and floundar (Plotichthys flesus). expressed in pG'ICQ on product base.
4. CCNCLUSICN In the area under sttXiy lCM concentrations ~re observed for the orga
nochlorine CCIlpOunds subject to analysis. The levels present neither risksfor survival and normal development of living species nor to public health.
The values obtained are similar to those found in other european zones which are considered to have a lCM pollution degree.
ACKNavLECGEMENTS We would like to thank Mrs. I.Silva and M,C.Caleiro of our Department
for their assistance in the analytical ~k, ~ also wish to thank all our collegues contributed with treir advice and support to the elaboration of this manuscript.
REFERENCES
(1) DVINKER,J.C., HILLEBRAND, M.T.J.,PAIM)RJ<,K.H. and WIIHEIMSEN,S. (1980), Bull.Environrn. Contam. Toxicol. ,25, 956-964.
(2) DVINKER,J.C. and HILLEBRAND,M.T.J:-(19R3). Bull. Envirom, Contam. Toxi col., 31, 25-32.
(3) B1ILI.SMITER,K. and ZELL, M. (1980) .Fresenius Z.Anal. Chan., 302, 20-11. (4) ICES (1984). Appendix to Coop. Res. Rep. N9 126. -(5) TUINSTRA,L.G.M.Th., ROC6,A,H. ,GRIEPINK,B, ,WELLS,D.E. (1985). J.High
Resolution Chranatogr. and Chranatogr. Commun., e I AU3., 475 - 479 (6) KERKHOFF,M. (1985) ICES -'M:}VG 1985/8.2. -
- 87-
Summary
HPLC/FLUORESCENCE SPECTROMETRY IN ANALYSES OF PULP MILL WASTES IN RECIPIENTS
P. MIKKELSON, J. PAASIVIRTA and J. KNUUTINEN Department of chemistry, University of Jyvaskyla Kyllikinkatu 1-3, SF-40100, Jyvaskyla, Finland
This is a method for determination of lignin and humic substances in waters by HPLC with fluorimetric detection. Water samples are extracted with ethyl acetate and the carboxylic acids in these extracts are labeled with 4-Br-MMC. Results indicate that this method might be applicable to characterzation of humic waters and differentiation of humic and lignin compounds in recipient waters of pulp mills. Further experiments are needed.
1. INTRODUCTION There is an urgent need to develop a routine method for unambiguous
determination of residual lignin matter in recipient waters of pulp mills. In continuation of our previous reports (1,2) this paper presents one way to analyse lignin residues in humic waters. For this purpose aqueous reversed-phase (RP) high-performance liquid chromatography (HPLC) has been tested. Possible differences between the fluorescence properties of the 4-bromomethyl-7-methoxy-coumarin derivatives of lignin and humic matter prompted us to carry-.. out this study.
2. MATERIALS AND METHODS Studied waters were kraft pulp mill waste liquor, unpolluted natural
humic water (above the pulp mill) and polluted humic water (below the pulp mill). 1500 ml of each water sample was continuously extracted with 300 ml of ethyl acetate for 24 hours. The extract was cinsentrated and evaporated to dryness with nitrogen. The residue was derivatizised with 4-bromomethyl-7-methoxy-coumarin applying the procedure of Lam and Grushka (3). After filtration (0,2 pm filter) and removing of the precipitate formed, 5 pI of the filtrate was injected into the chromatograph. Nordic reference fulvic and humic acids were handled similarly.
The used instrument was Perkin-Elmer LS-4 fluorescence spectrometer with Perkin-Elmer 3600 luminescence data station and Perkin-Elmer series 10 liquid chromatograph. The column was Spherisorb 5 ODS.
3. RESULTS AND DISCUSSION Fluorescence chromatograms of the samples are shown in Fig.l.
It is seen very clearly that the majority of the carboxylic acids in the kraft pulp mill waste liquor (B in Fig. 1.) are different from those in natural fulvic acid (D). Characteristic peaks of the waste liquor acid components appear also in recipient water sample (C) but not in water sample taken above the mill (A).
- 88-
A D
o 5 10 o 5 10 min
o
B Ak E
6 """'" ~ 10 5 10 min
C ~ F
o 5 10 0 5 10 min
Fig. 1. HPLC traces of fluorescence derivatives of studied samples. A = unpolluted natural humic water, B = kraft pulp mill waste liquor, C = polluted water sample, D = Nordic reference fulvic acid, E = Nordic reference humic acid, F = reagents. Amounts injected 5 pI/sample, eluent 35 % aqueous methanol, flow rate 1,8 ml/min, exitation 360 nm and emission 410 nm.
- 89-
REFERENCES
(1) KNUUTINEN, J., VIRKKI, L., PASTINEN, O. and PAASIVIRTA, J. (1986) Aqueous liquid chromatographic determination of humic and lignin compounds in surface waters. International Humic Substances Society Third International Meeting, August 4-8, 1986 in Oslo Norway; Abstract p. 102.
(2) KNUUTINEN, J., VIRKKI, L., MANNILA, P., MIKKELSON, P. and PAASIVIRTA, J. (1987) Aqueous high-performance size-exclusion chromatography of dissolved organic matter in natural waters. Water Res. (submitted).
(3) LAM, S. and GRUSHKA, E. (1978) Labelling of fatty acids with 4-bromomethyl-7-methoxy-coumarin via crown ether catalyst for fluorimetric detection in high-performance liquid chromatography. J. Chromatogr. 158, 207-214.
-w-
CHLOROPHENOL ~C~O~M~P~O~U~N~D~S _I_N _S_N_O_W
R.PAUKKU,a J.PAASIVIRTA,b M.KNUUTILA band S.HERVEc
aInstitute for Environmental Research, University of Jyvaskyla Tellervonkatu 8, SF 40100 Jyvaskyla, Finland
bDepartment of Chemistry, University of Jyvaskyla Kyllikinkatu 1-3, SF-40100, Jyvaskyla, Finland
CWater and Environment District of Central Finland P.O.Box 110, SF-40101 Jyvaskyla, Finland
Summary
Polychlorinated phenols, guaiacols and catechols have been studied in snow from North Pole, North, Central and South Finland. Only Central and South Finland samples contained known compounds above limit of determination 0.5 ng/l. Airborne chlorophenol pollution appeared to be very much higher in South than in Central Finland.
1. INTRODUCTION Chlorophenol compounds have been found very frequently in
environment. They are industrial products and intentionally spread to environment in pesticide and slimicide uses. The most significant emissions of chlorophenolic compounds into the environment, however, rise from chlorination and combustion processes of organic materials (1). Therefore, methods for monitoring chlorophenols in air are of interest. Recently, we found that snow attracts chlorophenols and might be useful in following transportation of the pollutants at,cold climate areas (2).
2. SAMPLING Sampling areas in Finland are presented in Fig. 1. Snow
sampling May 1984 at North Pole has been reported together with other analysis results (3). The samples in Finland were taken by us using pre-measured plastic cylinder as device. Surface area, volume and weight of the snow sample was measured.
3. MATERIALS AND METHODS For chlorQPfienol analyses one litre of water (melted
from snow) was taken, a measured amount of 2,3,6-trichlorophenol was added to make an internal standard. Potassium carbonate was added to make the solutio~ 0.1 molar, and then shaken 5 minutes with 5 ml acetic anhydride. Then, 5 ml of hexane (Rathburn) was added and shaken again 5 min. The hexane layer was evaporated to a small volume and injected to gas chromatograph Orion Analytica MICROMAT equipped with two non-identical columns (OV-1701 and SE-54) both leading to ECD.
- 91-
..... "'"", , , HOR'IIAY r "
,\f\ ' / , I I
\ \ '" ,,' ~ '" ... ,,, ...... 1 I
'\ 'I S II E D E H .~ lAP L A H 0 ' , , \
'1 / \ I
I .11 : USSR I \ I 0 IbvNjlEIII \ \ \
N \ ,
t / • s ( '. I
\ , Fin .and I
;' \ ... .. ,
~ o KUOPIO :
OJ't"IXsKvLl / I
I
OTAIIP[R[
" I /
" I
J
I
I-'-+-i o 50 100 KII
Fig. 1. Sampling areas or places indicated as dark spots. Sampling at South Finland (MANKKAA) 1.4.83. Sampling (J1-J4) at Jyvaskyla City 3.4.83 and 25.1.84. Sampling at ten lake ices in Central Finland (1-10) 22.2.85. Sampling in Lapland (11; Rattilampi) 25.4.85.
The retention times by both columns had to match to those of our authentic reference compounds (we have all 43 theoretically possible polychlorinated phenols, catechols and guaiacols, however, only 14 compounds presented in Fig. 2 are frequently found in Finnish environment). GC run was performed using helium as the carrier gas 1 ml/min and argon/~5thane as makeug for ECD 30 ml/min. Temperature program 1000+ 4 Imin up to 250 was used.
-92-
In addition to the waters obtained from the snow samples under study, also zero water sample and zero water with added measured amounts of authentic chlorophenol reference standard compounds were treated by the above procedures. structures, abbreviations and names of the reference compounds are given in Fig. 2.
r' C~I CIt&:: ~ of Int. St.
24DCP 26DCP 236TCP 246TCP 245TCP 2,4-Dichlo- 2,6-Dichlo- 2,3,6-Tri- 2,4,6-Tri- 214,S-Tri-
rophenol rophenol chlorophenol chlorophenol ch orophenol
~ * ~ 4: * COl C I I I
2346TeCP PeCP 34DCC 345TCC TeCC 2,3,4,6-Tetra- Penta- 3,4-Dichlo- 3,4,5-Trichlo- Tetrachlo-chlorophenol chlorophenol rocatechol rocatechol rocatechol
~ *'~ *~ ",e* "~~ CI C I CI I CI I CI I I
45DCG 345TCG TeCG DMP 456TCG 4,5-Dichloro- 3,4,5-Trichlo- Tetrachlo- Trichloro- 4,5,6-Trichlo-
guaiacol roguaiacol roguaiacol 2,6-dimet- roguaiacol hoxyphencl
Fig. 2. Structures, name abbreviations and names of the chlorophenol compounds studied.
4. RESULTS AND DISCUSSION From standard runs 1t can be estimated that our limit of
determination was 0.5-1 ng/l for each tri- to pentachlorophenol. Assumin~ihat 2 months precipitation on average is less than 100 k9t:. , determination limit for fallout must be lower than 50 ng/m • Using larger samples and more sensitive method a research group from Japan has measured in Antarctic snow chlorohydrocarbon levels 0.009-0.017 ng/l for DDT residues and 0.16-1.00 ng/l for PCB (4). This must correspond for individual compounds a limit of determination around 0.003 ng/l or better. Remarkably, levels of HCH including LIND were 1.5-4.9 ng/l in Antarctic snow (4), while we find none (less than 0.5 ng/l) in Finnish and North Pole snow (2). Chlorohydrocarbon pollution is more readily measured by analyses of animals. Recently, Jensen has reported about chlorohydrocarbons in organisms of Arctic islands, where animals at Southern Svalbard contained significantly higher levels than those at Northern Svalbard (5).
Chromatograms of the chlorophenol acetates from snow samples contained conSIderably more and higher peaks than those of chlorohydrocarbons (2). For comparison, ECD chromatogram from analysis of the dilute water solution of the standard mixture is illustrated in Fig. 3 (FID appeared as a straight line).
- 93-
!I , ~ 1', ~ ~"": ::! • ~ , .., '" Chlorophenol ::; ~1 rLl.i. .... ~. 1 • • cOIIpound I' '" M I 'j ~ ~ standard.... <'! (iii " J ~ . ~"..... SE.- S4 . M (\j • ~ 0 ~~y~ " ' ! [CO 41!!v. ~ i, f ~ ~f.voJ'1 , ~~~r.l.f-"~ 1~~1 I W h.~~ nl '.V I ..J~I '\ . ; Iy ~w.w .! ! t t r '''-'-'°f'Vj'' t t \ t \
...1': \ 1.2. 4 . ' . 5. 6. 1,. 8. 12. 15. 7.9. 1,. 14 . 10.
Fig. 3. GC/ECD trace of acetylated chlorophenol standard mixture. Name abbreviations and contents in zero water, from which the model treatment was done:
1. 3. 5. 7. 9.
11 • 1 3. 15.
24DCP (5.8 ng/l) 236TCP (3.2 ng/l, Int.St.) 245TCP (5.65 ng/l) PeCP (1.00 ng/l) 345TCC (2.5 ng/l) 45DCC (6.65 ng/l) TeCG (1.15 ng/l) 456TCG (5.76 ng/l)
2. 4. 6. 8.
10. 12. 1 4.
26DCP (6.0 ng/l) 246TCP (5.45 ng/l) 2346TeCP (0.95 ng/l) 34DCC (4.8 ng/l) TeCC (0.8 ng /l ) 345TCG (2.10 ng/l) DMP (1.0 ng /1)
standard (Fig. 3), blank sample and snow sample runs showed that the limit of determination for chlorophenols in the present method was for most components of the standard lower (for TeCC higher) than 1.0 ng/l which cor~esponds for 2 months snrw precipitation (water value) 100 kg/m better than 0.1 ug/m • In North Pole snow none of the chlorophenols studied could be detected at this level. Not even unknown peaks were observed: ECD is similar to laboratory blank and FID is a straight line.
In cont~ary, snow sample from Jyv~skyl~ (Central Finland) taken ln January 1984 produced ECD chromatograms rich on peaks. The most peaks, however, were unknowns, only two chlorophenols (2346TeCP and PeCP) could be identified and quantified. This corresponds a slightly lower pollution than earlier results (1) from samples of Jyv~skyl~ April 1983 from which also 246TCP, 345TCC, 456TCG and DMP were measured. From South Finland (MANKKAA) samples April 1983 we identified and measured nine chlorophenols: 246TCP, 245TCP 2346TeCP, PeCP, 345TCC, TeCC, 456TCG, TeCG and DMP in very variable amounts depending on locality (1).
While the samples 1984-85 in Central Finland were taken after similar cold snowing period of two months, the results (Table I) are reasonably comparable and show that chlorophenol pollution of air is significantly higher at urban than at rural areas, in accordance with our previous results 1983 (1). Sample 3 was from the lake ide near a pulp bleaching plant and showed traces of chlorinated catechols and guaiacols. Thus one could state that these chlorination products are emitting to air from bleaching or from combustion of pulp mill wastes .
- 94 -
TABLE I. Fa1l2ut of chlorophenols in snow samples of Central Finland ug/m • Sampling places are indicated in Fig. 1. Statistics: X = Average, s = St. deviation, N = Number of cases.
Sample Nr
246 TCP
245 TCP
2346 TeCP
PeCP
City of Jyvaskyla January 1984
J1 J2 J3 J4
x = s = N =
0.00 0.00 0.00 0.00
0.00 0.00 0.00 0.00
4.50 8.30
17.50 9.30
9.90 5.47
4
Lake ices February 1985
1 2 3 4 5 6 7 8 9
10
0.13 0.35 0.35 0.00 0.08 0.04 0.09 0.17 0.28 0.08
X 0.157 s = 0.126 N = 10
0.00 0.00 0.00 0.00 0.00 0.11 0.41 0.00 0.12 1.27
0.19 0.40
10
1.54 0.39 0.17 1. 75 0.20 0.26 0.35 0.96 1. 76 0.77
0.81 0.65
10
7.80 15.30 15.60 17 .60
14.08 4.31
4
0.98 0.48 0.84 1.48 0.20 0.62 0.58 1.43 1. 76 0.82
0.92 0.50
10
345 TCG
0.00 0.00 0.00 0.00
0.00 0.00 0.52 0.00 0.00 0.00 0.12 0.00 0.00 0.00
0.06 0.17
10
456 TCG
0.00 0.00 0.00 0.00
0.00 0.00 0.42 0.00 0.00 0.00 0.00 0.00 0.00 0.00
0.04 0.13
10
3456 TeCG
0.00 0.00 0.00 0.00
0.00 0.00 0.04 0.00 0.00 0.00 0.00 0.00 0.00 0.00
.004 11.1
10
Sum PCP
12.30 23.60 33.10 26.90
23.98 4.90
4
2.65 1.22 2.33 3.23 0.48 1.03 1.54 2.56 3.92 2.93
2.19 1.09
10
245TCP was main component at place 10 which is located near a highway ferry. This is an additional indication of 245TCP rising from heavy winter traffic in Finland (1).
The dominating components in snow fallouts of Central Finland (Table I) were 246TCP, 2346TeCP and PeCP which are typical combustion products but also used as pesticides, slimicides and in wood preservation. Their appearence in lake water shows large seasonal variations and in snow they are most abundant near waste combustion plants (1).
All chlorinated compounds in Lapland sample were below our limit of detection. Fallout o~ each chlorophenol in this sample must be less than 0.1 ug/~.
In snow at MANKKAA, April 1983 (Table II) The highest fallouts were found near heavy traffic (places M11 and M18). In these samples high fallout of 245TCP (precursor to the highly toxic and persistent 2,3,7,8-TCDD) is most remarkable. Another type of chlorophenol pollution was most clear in sample M20 where only 246TCP, 2346TeCP and PeCP were abundant. The place is in forest 1 km northeast from a municipal waste landfill (occasionally burning) area. The above three chlorophenols are known to be the main organochlorine products of combustion of the municipal wastes (1).
- 95-
TABLE II. Fallout (precipitation) of chlorophenols ug/lfi- 2calcu-lated from snow samples in South Finland 1.4.1983 (Pl. Nr 1-20) ND = not detected. --------------------------------------------------------------Plo 246 245 2346 456 345 DMP TeCG PeCP TeCC Sum Nr TCP TCP TeCP TCG TCC PCP --------------------------------------------------------------Ml 35.8 ND 34.5 ND ND ND 10.0 74.3 ND 74.3 M2 6.93 26.2 6.05 13.0 8.97 6.78 6.20 11.2 1. 60 86.9 M3 58.5 ND 47.9 ND ND 90.9 49.8 ND 247.1 M4 21.2 ND 28.8 ND ND ND 2.92 2.85 55.8 M5 42.2 ND 63.3 ND ND ND 63.2 ND 168.8 M6 ND 11.0 ND ND ND ND 1 .96 ND 13.0 M7 3.08 ND 13.2 5.88 4.48 8.39 ND 7.13 ND 42.1 M8 5.08 20.1 3.55 12.3 9.21 9.21 2.88 9.11 0.96 72.4 M9 4.33 ND 2.39 ND ND 5.56 ND 3.75 ND 16.1 Ml0 9.59 17.5 10.9 10.3 5.63 3.34 13.6 1. 85 72.8 Mll 55.8 354. 36.6 234. 169. 84.4 28.9 136. 26.3 1125. M12 3.72 ND 8.70 ND ND 11.3 ND 2.70 ND 26.5 M13 16.9 ND 30.5 2.74 ND 16.9 ND 19.8 ND 86.8 M14 20.5 ND 45.9 ND ND ND ND 7.43 ND 73.9 M15 4.00 ND 8.35 ND ND 9.04 ND 3.55 ND 24.9 M16 74.7 ND 60.6 ND ND ND ND 7.78 1 43.1 M17 21 .1 105. 75.0 32.0 21.2 10.9 42.9 308.1 M18 52.6 58L 42.1 172 • 1 51 • 106. 35.4 68.4 18.6 1227. M19 9.15 6.63 7.66 ND ND ND ND 1.49 ND 24.9 M20 134. ND 359. ND ND ND ND 29.5 521 • 7 --------------------------------------------------------------
The results might give more comparable numbers by estimating an annual fallout (observed snow fallout divided by snowing months before sampling and multiplied by 12). By this method total annual chlorophenol fallout 1983-85 would be in urban South Finland 3500, rural South Finland 350, urban Central Finland 150 and rural Central Finland 15 ug/m2 • The PCP fallouts in Lapland and at North Pole were~on-measurable - they seem to be annually lower than 0.6 ug/m •
REFERENCES
(1) PAASIVIRTA, J. et al. (1985). Polychlorinated phenols, guaiacols and catechols in environment. Chemosphere. Vol. 14, 469-491.
(2 )
( 3 )
( 4 )
( 5 )
PAASIVIRTA, J., KNUUTILA, M. and PAUKKU, R. (1985). Study of organochlorine pollutants in snow at North Pole and comparison to the snow at North, Central and South Finland. Chemosphere. Vol.14, 1741-1748.
KARLSSON, V., HASANEN, E. and NORLUND, O.P. (1985). Major inorganic ions in North Pole snow samples. Chemosphere. Vol.14, 1127-1131.
TANABE, S., HIDAKA, H. and TATSUKAWA, R. (1983). PCBs and chlorinated hydrocarbon pesticides in Antarctic atmosphere and hydrosphere. Chemosphere. Vol.13, 277-288.
JENSEN, s. (1984). Klorerade kolvaten i limniska och marina organismer i Arktis. Communication at "Nordisk m¢te om organiske milj¢gifter, Geiranger, Norway 10-13.9. 1984.
-%-
'!HE ANALYSIS OF OOOROOS SUI.mUR exJ.1RJUN]E BY GAS ~ AFTER 'IHERofAL DESORPI'ION FRCM TENAX
I W D:lvies am J Yates SAC (CllranatclgraIily) am water Research Centre
An evaluation of an analytical methcx:l for volatile sul{ilur conpourrls arose fran the need to nonitor conpourrls knc:Mn to cause nuisance odours in atmJs{ileres near sewage treatment works. A technique using thennal deso:rption fran Tenax adsoment, cold trapping am transfer on to a capillary GC cohnnn before chranatclgraIily which has been applied to qualitative analysis of a broad rarge of organic conpourrls by gas chranatclgraIily-mass spectranetry was evaluated am subsequently IOOdified to meet the d.emarx:Is of the analysis of a rarge of reactive volatile sul{ilur conpourrls by gas chranatogra{ily using a sul{ilur-selective flame {ilotcmatric detector. Factors affecting reaJVeries fran the adsoment were identified am efficiency of transfer significantly inproved.
1. INl'ROIXTcrION As part of a progranme to identify the causes of nuisance cx:iour
problems near treatment works, a methcx:i was required for the quantitative determination of a rarge of volatile sulIilur conpourrls in atIoos{ileric samples.
Sane essentially qualitative work has previously used thennal desorption to transfer ethanethio1 am isopropanethiol am the more volatile sulIilides fran Tenax GC am Porapak Q on to a packed GC column (1), l:ut reaJVeries fran concentration traps on to capillary GC columns is often incaIplete am the factors governi.rg efficient sariple transfer poorly urrlerstood (2). '!he awlication of a thennal desorption/cold trapping technique to the followirg volatile, sul{ilur-containirg, cx:iorous conpourrls of interest was therefore investigated to identify the causes of non-quantitative transfer (3).
Boilirg point (Oc)
1. ethyl mercaptan (ethanethiol) 35 2. propyl mercaptan (propanethiol) 67 3. isq>ropyl mercaptan (2-propanethiol) 52 4. allyl mercaptan (1-propene-3-thiol) 67 5. bltyl mercaptan (b.rt:anethiol) 98 6. s-bltyl mercaptan (2-b.rt:anethiol) 85 7. i-bltyl mercaptan (2-methylpIqlal'lethiol) 87 8. t-bltyl mercaptan (2-methyl-2-propanethiol) 62 9. pentyl mercaptan (pentanethiol) 105 10. carbon disulIilide 46
-97-
11. dimethyl sulpride (2-thiapropane) 36 12. dimethyl disulpride (2,3-dithiabutane) 108 13. ethyl nethyl sulpride (2-thiabutane) 67 14. diethyl sulpride (3-thiapentane) 92 15. dipropyl sulpride (4-thiaheptane) 142 16. bis-nethyl thio methane (2,4-dithiapentane) 147 17. 1,2-ethanedithiol' 146
2. G1IS~ Before . assessin;J the themal desorption transfer am CJ:yofocusin;J
technique scm::! prelilninary GC was done to ensure separation of the COltp)UI'rls. Chranatogrcqilic resolution was ootained on a 50 m x 0.32 nun Ld. fused silica capillary cohnnn coated with a 1 J..£IlI film of i.m!oobilised polydimethylsiloxane by ten'perature progranmirg fran 200c to 1000c at 4oC/min usin;J hydrogen as carrier gas at a flow rate of 2 ml/min. Typical separations are shown in Figure 1. '!he operatin;J ten'perature am lergth of colUlTlll inserted into the sulIilur-selective flane Iilotaootric detector (FPO) used were optimised to ootain maxllnum sensitivity for the ran:Je of COltp)UI'rls analysed.
3. 'IHERMAL DESORPI'ION TEClINIspES 3.1. 'Ihennal Desorption am Cold TrappiI!:{ Via a Fused silica Transfer Line
'!he configuration of the thennal desorption heatin;J jacket, cold trap am GC analytical colUlTlll are shown in Figure 2.
Glass adsorption tubes (3 nun Ld), packed with Tenax GC (60-80 nesh), were inserted into an electrically heated jacket whose ten'perature was controlled usin;J a rl1eostat am m:mitored by a thenoocouple. '!he transfer line consisted of a lergth of about 10 em of 0.17 nun o.d. fused silica tubin;J. It was inserted through a conventional on-colUlTlll injector to approxilnately 5 em above a 15 em lergth of the analytical colUlTlll inurersed in a flask of liquid nitrogen to act as a cold trap. '!he transfer line was inserted into a region above the cold trap to prevent solidification of deso:rbed material on or at the errl of the transfer line which could cause either a blockage or losses when the line was withdrawn fran the colUlTlll.
3 . 1. 1. Method '!he adsorption tube was installed in the heatin;J jacket am connected
to the transfer line which was then lowered into the colUlTlll through the on-colUlTlll injector. '!he liquid nitrogen cold trap was installed in position am hydrogen desorption carrier gas was passed through the Tenax. '!he adsorption tube was then heated for a set tine after which the transfer line was withdrawn fran the analytical colUlTlll, the on-colUlTlll injector valve closed am the cold trap rem:wed before c.hranatograrhing the sample. '!he desorption carrier gas flow was maintained while the adsorption tube was allowed to cool.
Deso1:ption was carried out at various ten'peratures, desorption tines am carrier gas flows through both the colUlTlll am the adsorption tube. Transfer efficiency was evaluated by dosin;J the Tenax with microlitre quantities of a stamard solution containin;J between 4.0 am 12.5 ng/ILl of the sulIilur COltp)UI'rls am c::x:nprrin;J responses with on-colUlTlll injections of similar volumes of the stamard solution.
3.1.2. Results '!he followin;J conclusions were drawn. with the exception of 1,2-ethanedithiol all the COltp)UI'rls could be
deso:rbed fran Tenax GC although recoveries were rarely higher than 70% am
- 98-
often below 40%. For harologues transfer efficiency increased with volatility. Transfer of c::c.atpCX.1l'd of similar volatility decreased with chemical
reactivity. '1hus thiols were transferred less efficiently than disulprldes am disulprldes less than sulprldes.
Transfer of the less volatile CCIIp)lII'rls continued to increase usin} desorption times of up to 7 minutes.
Transfer increased with desorption tenpllature. Recoveries were inproved by between 10 am 20% :1Y increasin} the desorption tenpllature fran 1500c to 300Oc. 'l'enileratures of up to 5000c were used with no noticeable deterioration in the performarx::e of the Teriax GC even though the manufacturer's recamnerx:led uwer t:enpllature limit for the ~ is 375Oc; however, there was little gained in transfer efficiency above 300Oc.
IrM transfer efficiency was not priInarily due to irreversible adsorption or chemical reaction in the tube or transfer line. Re-heatin} adsorption tubes denonstrated that significant annmts of CCIIp)lII'rls particularly ethanethiol, diethyl sulprlde, diIrethyl disulprlde, bis-nethylthioethane am dipropyl sulprlde ca.lld still be transferred followin} the initial desorption, trawin} am chranatograIbY cycle.
Increasin} desorption carrier gas head pressure fran 0.15 to 0.25 kg an-2 increased transfer by a factor of two for the rore volatile am by three to four times for the least volatile CCIIp)lII'rls.
3.2. '!henna! Desmption Directly on to the Analytical Column '!he need to ensure that sanple is swept fran the transfer line on to
the GC column .inposes the limitation that the desorption gas flow should not exceed the carrier gas flow. However, higher desorption gas flows needed to be investigated since, even within a narrov.r rarge, increased flows led to greatly increased sanple transfer. since it was not possible to accurately nonitor the flows at the end of the transfer line am at the head of the GC column the awaratus was ll¥Xli.fied to eliminate the two iIrleperrlently controlled gas flows by cormectin} the GC column directly into the desorption tube heatin} jacket, the column bein} passed through an unused heated detector port in the top of the GC oven. '!his configuration also redu.ced. the possibility of the sanple COIll:in;J into contact with active sites.
3.2.1. Method Usin} the ll¥Xli.fied awaratus the desorption procedure was dlanged to
incorporate a preliminary ambient tenpllature purge to ensure that the Tenax was free of oxygen am to further reduce the possibility of chemical reaction durin} desorption. '!he flow was maintained at 10 ml/roin durin} purgin} am durin} the subsequent desorption/cold trawin} stage before bein} reduced to 2 ml/roin for the chranatograFhY stage.
3.2.2. Results '!he results obtained de.roc>nstrated several inprovements on the previous
procedure. with the exception of 1,2-ethanedithiol, which again was not desorbed,
the recovery of all CCIIp)lII'rls was inproved; for the IOC>st reactive CCIIp)lII'rls, such as propanethiol, to 20-30% am for the least volatile am reactive CCIIp)lII'rls to nearly 100%.
J)lrin} the ambient tenperature purge it was noted that pentane used as solvent for the solution of the sulFhur CCIIp)lII'rls was also swept away. Small losses of the IOC>st volatile sulFhur c:onp::lUlrl3 (ethanethiol, dimethyl sulprlde, cal:bon disulFhide am isopropanethiol) sanetimes occurred durin}
- 99-
the plrge but attenpts to cold trap this material resulted in poor chraI1at:ograprlc peak shapes althoogh the reasal for this effect is not clear.
IncreasinJ ·the preliminary plrge fran ale to foor minutes irx:reased the transfer durin;J the subsequent thennal desoIpt.ion fran 10 to 60% for iscpropanet:hiol an:! fran 15 to 100% for dimethyl disulpude, although lager pn-ge times of up to six minutes did not further irx::rease reocveries except for the least volatile c:atpClllI"ds (pentanethiol, bis-met:hylthiomethane an:! dipropyl sulpude). r';:here was, on the contrcuy, a trerrl for the reocveries of the mre volatile CCIIpOUl'X3s to decrease usinJ a lon;Jer than optilrum foor minute pn-ge.
'!here was no eviderx::e that significant annmts of any of the added ~ ccW.d be desoJ:Ded after the initial desozpti.avtrawinJ cycle an:! no adverse effects on dlranatc:lgraply liI'eI"e noted by continu.in:J to nm the carrier gas t:hroogh the adsoJ:bent after desol:ption.
3.3. 'Blennal Desol:ption fran Different AdsoIbent Materials In aatition to Tenax Ge, six other adsol:bents liI'eI"e investigated: silylated Tenax Ge Tenax'm silylated Tenax 'm activated camon Porapak QS cartq:lack B coated with 1.5% XE60/1% H3R>4 '!he last two of these are sold as packirg materials for packed. column
Ge analysis of volatile sulprur CCIIpOUl'X3s an:! activated camon has been suooessfully awlied to the analysis of sulprur CCIIpOOl'D; in the Grab closed loc:.p striwinJ awaratus.
None of these materials led to an lltprovement in transfer efficiency.
(1) ROE, A. B. J. Institute of water Ergineers an:! Scientists, 1982, 36(2), 118-128.
(2) GroB, K. an:! HABle, A. J. CllranatograIily, 1985, 321, 45-58. (3) water Pollution Control Federation. Manual of Practice no. 22.
washin;Jton DC, USA.
-100-
§
1(a)
Fig
ure
16
3 10
3
15
12
2 6
1311
5 71
4
II
" II
.
I I
12
5 7
~U.J
1(
b)
Ch
rom
ato
gra
ph
ic
sep
ara
tio
n o
f su
lph
ur
com
po
un
ds
by
(a)
on
-co
lum
n in
jecti
on
an
d
(b)
therm
al
deso
rpti
on
fr
om
T
enax
G
C
tube furnace
o;r cool ing
Tena. adsorb.nt
·cop i llory needle
1 iquid nitrogen
corrier ga~ inl et
II~I-I---nd~orp'ion tube
yolve le.e •
.-'Ullli---.~-=+--c arrier 90S
capillary column
, . ~ , ,. I, I I II I ' II II II
Figure 2 The1'lll8l desorption apparatus
- \02-
DETERMINATION OF ORGANIC CHEMICALS IN SEDIMENTS TAKEN FROM THREE UNPOLLUTED ESTUARIES IN SOUTH WEST ENGLAND
B.J. HARLAND* and R.W.GOWLINGt
* ICI Brixham Laboratory" Devon t Trent Polytechnic, Nottingham
Summary
Sediment samples from three relatively unpolluted estuaries in south west England have been examined. All three estuaries (viz those of the rivers Dart, Exe and Teign) drain predominantly agricultural land and receive only low volumes of primarly domestic effluents. The results of this investigation have shown that the major components in the samples are of biogenic origin. They include components of the leaf waxes of higher plants (nC27 , nC29 and nC31 alkanes), a long chain (C26 ) alcohol and sterols. Two of the latter, coprostanol and cholesterol, probably originate from sewage or animal waste discharges to the estuaries.
1. INTRODUCTION Estuarine sediments are able to accumulate persistent and hydro
phobic organic chemicals from the overlying water and characterisation of this accumulated material can provide useful information about the environmental quality of the estuary and the type of inputs to which it is subjected.
To assess the type of chemicals which might be accumulated by sediments in relatively unpolluted estuaries, sediment samples from three estuaries in south west England have been examined. All three estuaries, viz those of the rivers Dart, Exe and Teign, drain predominantly agricultural areas and receive only low volumes of primarily domestic effluents. Consequently, the results of this investigation should indicate those organic chemicals which can be expected in unpolluted estuarine sediments and should produce a useful baseline against which the results from more polluted or industrialised estuarine sediments can be compared.
2. EXPERIMENTAL Sediment samples
Sediment samples were taken from selected locations in the estuaries of the rivers Dart, Exe and Teign in south west England (see Figure 1). In each case the top 4 cm of the surface was sampled using a metal scoop and the collected sample was stored in glass bottles at 4°C until analysed.
-103 -
Analysis of samples Sediment samples were freeze-dried and then soxhlet extracted with
dichloromethane. The concentrated extract was analysed by capillary gas chromatography (FID) on a Varian 3500 gas chromatograph. A 25 m CP-Si18CB bonded phase FSOT column was used and the temperature program was 50°C (2 minutes) - 8°C/min - 300°C (10 min), the (nitrogen) carrier gas flow rate was 2 ml/min.
Mass spectrometry analysis (GC-MS) was performed on a Finnigan 8200 mass spectrometer. The chromatographic conditions employed were similar to those for GC analysis except that helium was used as the carrier gas.
3. RESULTS AND DISCUSSION In contrast to the methods of others working on the analysis of
organic chemicals in estuarine sediments (1,2), the sediment extracts obtained in this work were analysed by GC and GC-MS without prior cleanup or fractionation. Although this approach does have some disadvantages, it does allow visualisation of all the material which will elute under the gas chromatographic conditions employed.
The chromatograms of the total extracts from the Exe, Dart and Teign estuaries are shown in figures 2 (a),(b) and (c) respectively. All three sediments, although taken from separate estuaries, appear to be relatively similar in pattern and are particularly complex in the later part of their chromatograms. In each case, the greater part of the material appears to elute between the retention times of n-heptacosane (nC27) and n-tritriacontane (nC33). (An expanded version of this area for the Teign estuarine sediment is given in Figure 3.) The early part of the chromatograms,in Figure 2, ie that between the retention times of n-decane (nC IO ) and n-hexacosane (nC26 ), is comparatively devoid of peaks and this is in contrast to the chromatograms of sediments from industrialised estuaries (cf Figure 4) which are usually complex in this area, because of inputs from oil and other anthropogenic materials.
Although on first inspection the chromatograms of the sediments from the three estuaries appear to indicate a typical leaf wax pattern, ie nC27 , dominant nC 29 , nC31 , mass spectrometry analysis shows that the situation is in reality more complex. The nC 29 peak contains' another component, which may well be the major component in the sediment samples from all three estuaries. This is believed to be the long chain alcohol, hexdcosanol. The presence of its C28 homologue is also suspected at the nC 31 retention time but has not yet been confirmed.
The most likely sources of the nC27 , nCZ9 and nC31 , long chain alkanes are the leaf-waxes of higher plants (3) and this may also be the source of the long chain alcohols of which those of 26 and 28 carbon chain length are known to be the major homologues in plant waxes. However, the assignment of these alcohols of to a higher plant source cannot be automatic since they can originate from other sources (4).
Two long chain components whose structure have not yet been elucidated are also present in all of the samples and are marked X and Y on Figure 3. They appear to be homologues, two carbon units apart, and for this reason are probably of biogenic origin.
The other major components in the sample chromatograms are sterols, and of these only cholesterol and coprostanol have been identified to date. The latter is regarded as being an indicator of sewage discharges (5), since it is produced from cholesterol by bacterial reduction in the gut of higher animals. The other sterols have not been identified, and are probably from both terrestrial and marine sources.
-104 -
4. CONCLUSIONS The major components found in the sediment samples from the Dart,
Exe and Teign estuaries appear to be of biogenic origin. They include components of the leaf waxes of higher plants (nC27 , nC29 and nC31 alkanes), a long chain alcohol (hexacosanol), two unidentified long chain components, and sterols. Two of the latter, coprostanol and cholesterol, probably originate from domestic sewage or animal waste discharges to the estuary.
REFERENCES (1) READMAN, J.W., PRESTON, M.R. and MANTOURA, R.F.C. (1986) An
integrated technique to quantify, sewage oil and PAR pollution in estuarine and coastal environments. Mar. Pollut. Bull. 17, 298-308.
(2) BUCHERT, H., BIHLER, S. and BALLSCHMITER, K. (1982) Untersuchungen zur globa1en Grundbe1astung mit Umweltchemikalien.VII Hochauflosende Gas-Chromatographie persistenter Chlorkohlenwasserstoffe (CKW) und polyaromaten (AKW) in limnischen Sedimenten unterschiedlicher belastung. Fresenius Z Anal. Chem. 313 1-20.
(3) EGLINGTON, G, HAMILTON, R.J., RAPHAEL, R.A. and GONZALEZ, A.G. (1962) Hydrocarbon constituents of the wax coatings of plant leaves: a taxonomic survey. Nature 193, 739.
(4) SHAW, P.M., and JOHNS, R.B. (1986) The identification of organic input sources of sediments from the Santa Catalina Basin using factor analysis. Org. Geochem. 10, 951-958.
(5) HATCHER, P.G., KEISTER, L.E. and MCGILLIVARY, P.A. (1977) Steroids as sewage specific indicators in New York Bight sediments. Bull. Environ. Contam. & Toxicol., 17, 491-498.
-105 -
Figure 1 Map showing location of the estuaries of the Rivers Dart, Exe and Teign in south west England.
(0)
Figure 2 Chromatograms of the sediment extracts from the estuaries of (a) Exe (b) Dart and (c) Teign.
- \06-
27
Figure 3
x
29 26-0H
Expanded version of the later part of the chromatogram of the Teign estuarine seJiment given in Figure 2(c). The numbers refer to the chain length of the normal alkanes. Cholesterol and coprostanol are indicated by (Ch) and (Co) respectively. Hexacosanol is indicated by (26-0H) and X and Yare unidentified components.
Co
Ch
Figure 4 Chromatogram of a sediment extract from an industrialised estuary (run under slightly different chromatographic conditions from those in Figure 2). Cholest~rol and coprostanol are indicated by (Ch) and (Co) respectively.
- 107 -
ORGANIC PHOSPHATES IN SURFACE,GROUND AND DRINKING WATER
S.GALASSI*and L.GUZZELLA**
*Water Research Institute,CNR.20047Brugherio (Milano)-ITALY
**Biology Department,University of Milan-ITALY
Summary
During a survey undertaken in 1986 for triazine herbicide determination in surface and drinking water in Northern Italy,phosphate esters were found as ubiquitous interferring compounds,using the
GLC technique with a nitrogen/phosphorus specific detector .As these compounds occurred very frequently at levels higher than herbicides,they were identified and quantitatively determined. TBP(tri-n-butylphosphate),TIBP (tri-iso-butylphosphate) and TCEP(tris-2chloroethylphosphate) were very frequent in river Po, river Adige and Como lake waters as well as in drinking waters derived from these surface supplies.Groundwater was much less pol luted. Treatment processes could reduce this contamination: filtration on activated carbon was a very efficient system. However,reduction of this pollution at the source,at least for the most dangerous compounds,should be the most convenient policy.
1 . I NTRODUCTI ON Trialkyl and trihaloalkyl phosphates are widely used as plasti
cisers and flame-retardants. It is generally assumed that environmental pollution by these compounds is caused by inflow of industrial and dom
estic wastewaters. Indeed,they have been identified in surface waters(1-3)
and drinking waters (4). Recently,tris(2-chloroethyl)phosphate(TCEP) has
been detected in drinking water from groundwater supplies in Lombardy.
Most of the biological effects of phosphate esters is still unknown but it was found that some of them present a considerable risk for human exposure due to mutagenic activity(5) and tendency to bioconcentrate in adipose tissues(6).
In this paper preliminary results of the occurrence of tri-butyl phosphate(TBP),tri-iso-butylphosphate(TIBP)and TCEP in surface, ground and
-108-
drinking water in Northern Italy will be presented.River Po was studied
with particular attention to its initial and terminal reach,where the
water supplies of the municipalities of Turin and Ferrara are located.
Other surface and groundwater supp 1 i es were also exami ned before and after the treatment processes.The depuration efficiency for phosphate ester removal was also evaluated.
2.EXPERIMENTAL
2.1 Sampling sites Fig.1 shows the sampling site 10cation.Water treatment plant capacity
and processes are reported in Table 1.
2.2 Analytical methods Water samples were stores at 4°C in glass bottles unti1- phosphate
ester extraction,that was performed within 72 hrs.For this purpose Octadecy1cC 18 ) disposable co1umnsC1 m1-Baker) were used at a flow rate of 5-8 m1/min.Prior to extraction on C18 columns raw water samples were filtered on 0.45M;11ipore'membranes.Phosphate esters were eluted from the
column with 0.5 m1 of methanol and quantified by GLC ana1ysis.Gaschromatographic determinations were performed with a Carlo Erba instrument(4200) equipped with a nitrogen/phosphorus specific detector. A glass column C2 mX 3 mm I.D.),packed with ULTRABOND RCCarbowax 20 M)was emp10yed.GLC
conditions were as fo110ws:he1ium carrier gas,30 m1/minjhidrogen,30 m1/minj air,300 m1/minjinjector and detector temperature,240°Cjoven temperature,
160°C.A1iquot of 1-2 u1 of methanol extracts were injected. The recovery from 1 1 of phosphate ester aqueous standard CTBP and
TIBP= 50 ng/1,TCEP=500 ng/1) was about 60% for TBP and TIBP and 10~1o for TCEP. A background contamination,in the range of 1-10 ng/1,was observed in some blanks.
3. RESULTS AND DISCUSSION
The preliminary results of this investigation are shown in Table 2. Tria1ky1phosphates were always present in surface and drinking water supplied by surface water.Groundwater at Da1mine contained TBP and TIBP
levels similar to background. TCEP was detected very frequent1y:the highest level was found in river Po at Turin in April.
The present results and literature data from many areas of the world
(Table 3) show that phosphate ester contamination is widespread in surface
and drinking water in developed countries.
-109 -
To evaluate the risk for human health and aquatic life associated to
this pollution,toxicological data have to be examined.Available toxico
logical data (Table 4) show that TCEP is mutagenic on bacteria (5) and
TBP is toxic to fish (6-8),
The levels of alkylphosphates found up to now in surface waters do
not seem to be acutely toxic to aquatic fauna.On the other hand,TCEP
presence in drinking waters may represent a risk for human exposure.
Actually a reduction of phosphate ester pollution can be achieved by
treatments:drinking water was generally less polluted than corresponding
raw water. Little increases were observed sometime but they were probably
due to rubber connections in the extraction apparatus. Phosphate ester
abatemantobserved in Turin drinking water with respect to river water was
probably achieved by the process of percolation on the activated carbon
bed. The decrease observed at Ferrara was more likely due to dilution of
treated river water with well water (Table 1). However, a complete removal
of TCEP was observed only at Dalmine where activated carbon filters are
used.This treatment was adopted to remove atrazine from contaminated wells (9).
4. CONCLUSIONS
The results presented herein as well as the recent finding of TCEP pollution in groundwater supplies in Lombardy and literature data
lead to think that phosphate ester diffusion is widespread in surface and
drinking waters in all industrialized countries.
Although some reduction of this contamination could be achieved by
treatments,a complete removal requires very expensive procedures.
Therefore,to avoid risks for human health and aquatic life,the pro
duction and use of the most dangerous of these compounds should be regu-
1 ated.
Toxicological studied should be also undertaken on those compounds
whose biological effects are still unknown.
5. REFERENCES
(1) Grob,K. and G.Grob.Organic substances in potable water and in its
precursors.Part II.Application in the area of Zurigh.J.Chromatogr.
90,1974,303-313. (2) Sheldon,L.S. and R.A.Hites.Organic compounds in the Delaware river.
Environ.Sci.Technol ,,~, 1978,1186-1194.
-110 -
(3) Ishikawa,S.,Taketomi,M.,Shinoara,R.Determination of trialkyl and
triarylphosphates in environmental samples.Water Res.,1985,~,
119-125.
(4) Williams,D.T.,Nestmann,E.R.,LeBel,G.L.,Benoit,F.M. and R.Otson.
Determination of mutagenic potential and organic contamination of Great LakeSdri nki ng water. Chemosphere, 1982 ,.!...!., 263-276.
(5) Nakamura,A.,Tateno,N.,Kojima,S.,Kaniwa,M.and T.Kawamura.The mu
tagenicity of halogenated alkanols and their phosphoric acid
esters for Salmonella typhimurium.Mutat.Res., 1979,66,373-380.
(6) Sasaki,K.,Takeda,M.Uchiyama,M.Toxicity,adsorption and elimination
of phosphoric acid triesters by killifish and goldfish.Bull.
Environm.Contam.Toxicol., 1981,~,775-782.
(7) Muir,D.C.G.Phosphate esters.In:The Handbook of Environmental
Chemistry.Vol.3,Part C,O.Hutzinger (ed.),1984,41-46.
(8) NIOSH.Registry of Toxic Effects of Chemical substances.Vol.l and
2,1977.
(9) Galassi,S. and V.Leoni.The problem of atrazine in drinking water
in Italy.Proceeding of the European Conference:Impact of agri
culture on water resurces.Consequences and persectives.Berlin,
21-23 september 1987.
- 111-
.,36 :3 U ., l tlOY
- 112-
til (]) ~ .~
til
en c
.~
~
Co E m Vl
Tab
le
1.
Cq
pacit
y an
d w
ate
r tr
eatm
en
t p
rocesses o
f th
e
fiv
e
dri
nk
ing
w
ate
r p
lan
ts(F
ig.1
)
T U
RI
N
FERR
ARA
CAVA
N El
LA
CO
M 0
DA
LMIN
E T
otal
cap
acit
y (l
/s)
(150
0)
(700
) (4
00)
(500
) (1
30)
Inta
ke
from
R
iver
PO
Riv
er P
O R
iver
Adi
ge
Com
o la
ke
Gro
und
wat
er
Pre
-ox
idat
ion
wit
h C
l02
Cl0
2 C
l02
,C1
2 N
aClO
Dec
anta
tion
X
X
-P
re-o
zon
atio
n
X
w
I
Pre
-fil
trati
on
sa
nd f
ilte
r
Cla
ri-f
locc
ula
tio
n
X
X
X
X
Per
cola
tio
n o
n ac
tiv
ated
car
bon
bed
X
Fil
trati
on
thr
ough
act
ivat
ed c
arbo
n X
Fin
al d
isin
fect
ion
C1
2 C
l02
Cl0
2 N
aClO
Mix
ing
wit
h w
ell
wat
er
X
X
Sto
rage
X
X
X
X
X
.. I
Tab
le
2 P
ho
sph
ate
este
rs
in
surf
ace,
gro
un
d
and
d
rin
kin
g
wate
r in
It
aly
(
ug
/l)
du
rin
g
19
87
Sta
tio
n
TUR
IN
(riv
er
Po)
TU
RIN
(d
rin
kin
g
wate
r)
FER
RA
RA
(r
iver
Po)
FE
RR
AR
A
(well
w
ate
r)
FER
RA
RA
(d
rin
kin
g
wate
r)
TlB
P
feb
ap
ril
0.1
36
0
.03
6
0.0
80
0
.02
8
0.0
38
0
.03
3
n.s
. 0
.00
5
0.0
13
0
.01
0
CAV
AN
ELLA
(r
iver
Ad
ige)
0
.01
4
0.0
17
CA
VA
NEL
LA
(dri
nk
ing
w
ate
r)
0.0
05
0
.04
4
COM
O (d
rin
kin
g
wate
r)
DA
LMIN
E (w
ell
w
ate
r 1
) D
ALM
INE
(well
w
ate
r 1
1)
DA
LMIN
E (d
rin
kin
g
wate
r)
0.5
70
0.0
01
0
.00
1
0.0
01
n.s
.
0.0
01
0
.00
1
0.0
01
TB
P
Per
i 0
d fe
b
ap
ri I
0.1
30
0
.04
2
0.0
57
0
.02
2
0.0
31
0
.02
0
n.s
. 0
.00
2
0.0
06
0
.00
7
0.0
11
0
.01
1
0.0
07
0
.01
3
0.0
10
0.0
01
0
.00
1
0.0
01
n.s
.
0.0
01
0
.00
1
0.0
01
n.s
. =
n
ot
sam
ple
d
TCEP
feb
ap
ril
0.0
32
1
.28
4
0.0
90
0
.06
3
0.0
38
0
.01
8
n.s
. 0
.01
4
0.0
87
0
.01
6
0.0
22
0
.01
2
0.0
16
0
.00
8
0.0
40
0.0
58
0
.00
1
0.0
01
n.s
.
0.0
05
0
.00
4
0.0
03
Table 3 - Concentrations (ng/l) of phosphate esters in surface and drinking water at various
locations (Muir, 1984)
1974 Japan TBP 10 - 580 1975 Geneva TBP 2 - 82 1976 River Wall(NL) TBP 10 - 10,000 1978 Delaware river TBP 60 - 2,000 1979 Great lakes
ci ty water TBP 0.8 - 29.5 TCEP 0.1 - 12.6
1970 Canadian city water TBP 0.05 - 62 TCEP 0.05 - 52
Table 4 Biological effects of phosphate esters
TBP TCEP
ACUTE TOXICITY oral on rat on fish
LD50 LC50 (g/kg) (mg/l)
3 - 12 1.2
8.8 90
-115 -
MUTAGENICITY on Salmonella
+
EVALUATION OF DEGREE OF POLLUTION OF TIBER AND
ANIENE RIVE~S BY NITRILOTRIACETIC ACID
L. Zoccolillo, G.P. Cartoni, M. Ronchetti, A. Delogu Dipartimento di Chimica, Universita di Roma "La Sapienza" P.le A. Moro, 5 - 00185 Roma (Italy)
Summary
The degree of pollution by nitrilotriacetic acid (NTA) of Tiber and Aniene rivers has been evaluated employing a methodology developed in a previous work (1). The anali tycal procedure involves extraction from water sample, conversion to trimethylester and analysis by capillary gas chromatography. The NTA concentration increase in both rivers when these pass through the city of Rome.
INTRODUCTION
The sodium salt of nitrilotriacetic acid (NTA) has been recognized as the most suitable substitute for sodium tripolyphosphate in detergents. The NTA introduction on large scale is however much questioned for the fears regarding possible health effects and its impact on the environment. The NTA is cited as possible carcirogen (2,3). Another negative aspect is represented by possibility that NTA, being a strong complexing agent, remobilize toxic heavy metals from suspended matter and from sediments of water bodies. This effect is evidently connected with NTA concentration in water and it is therefore of great importance to monitor the NTA concentration in natural waters.
In this paper is reported the evaluation of degree of pollution by NTA of Tiber and Aniene by extraction of NTA from water sample, conversion to trimethylester and analysis by capillary gas chromatography.
EXPERIMENTAL
Materials Formic acid, methanol, chloroform, acetone, were purchased
from Carlo Erba (Milan, Italy). All reagents were of analytical
-116-
grade; the chloroform was distilled before use. Nitrilotriacetic acid and the sodium salt of NTA were obtained
from Fluka (Buchs, Switzerland). Acetyl chloride (Carlo Erba, Milan Italy). A 10% solution of
acetyl chloride in methanol was prepared. The anion exchange was Dowex 2x-8 (200-400 mesh); the resin
was used after trasforming to formate form the chloride. Internal
Switzerland) 0.1 mg/ml.
standard solution (3,4-benzoquinoline, Fluka, Buchs, was prepared in chloroform at concentration of
Apparatus Gas chromatography determination was performed on a DAN I (Mon
za, Italy) Model 6800 instrument equipped with a nitrogen specific detector.
Bonded-phase fused silica capillary columns (22 m x 0.25 mm i.d.; thickness of SE 54 cross-linked stationary phase film 0.52 ~m) were employed.
GC-MS analysis was performed on a Hewlett-Packard Model 5890 gas chromatograph connected with a Model 5970 mass spectrometer and equipped with a Hewlett-Packard data system. Capillary columns of the same type as that used in GC were employed. The column was connected directly to the ion source of the mass spectrometer. The GC conditions were as that used for the NTA determination with the FID detector, except that helium was used as carrier gas (flow rate, 2 ml/min). Spectra were obtained by electron impact at 70 eV.
Sampling The samples, collected over a one-day period, were taken in
the middle of the river. The samples were preserved wi th 1% aqueous formaldehyde and
stored at 4°C up to analysis time. The location of the sampling stations is shown in Fig. 1.
Analytical procedure For isolution of NTA from water samples has been followed the
procedure reported in literature (4,5) using an ion-exchange column.
The water sample (250 ml) were acidified to pH 2.3 with formic acid and heated at 80 0 C for 20 minutes bubbling nitrogen through the solution to purge the carbon dioxide from the sample. After cooling, pH was checked and, if necessary, readjusted. The sample was then percolated through a glass column (150 x 6 mm i.d.) packed to a height of about 5 cm, with the anion-exchange resin. A slight vacuum was applied (flow rate, 4 ml/min). NTA was eluted with
-117 -
10 ml of 16M formic acid and the eluate evaporated to dryness in a rotary evaporator. The dry sample was added of 1 ml of acetyl chloride in methanol solution (10% v/v) and kept in a water bath for 20 minutes at 90-95°C. After cooling and addition of 1 ml of distilled water, the reaction mixture was extracted three times with 1 ml of chloroform. An internal standard was added to chloroform phase and the extracts were dehydrated with anhydrous sodium sulphate. The chloroform solution was evaporated to dryness under a stream of nitrogen, dissolved in 50-100 Ml acetone and analised.
The identity of NTA was verified by capillary gas chromatography mass spectrometry. The quantitative determination were carried out by comparison with standard solutions containing known amounts of NTA and of the internal standard.
RESULTS
In order to evaluated the degree of pollution by NTA of Tiber and Aniene rivers samplings were carried out in three different periods of the year 1987.
In Fig. 2 are reported the gas chromatograms obtained from distilled water (reference blank) (A) and with NTA added (B) and the chromatogram of a river water sample (C). The NTA concentration, expressed in Mg/l, is reported in Table I.
It is observed that, as expected, the NTA amount increase strongly in both rivers when they pass through the city of Rome.
REFERENCES
1) L. Zoccolillo, M. Ronchetti, Annali di Chimica, in press. 2) A. Goyer, H.L. Falk, M. Hogan, D.O. Feldman, W. Word, J. Natl.
Cancer Inst., 66 (1980) 869. 3) Mr. Infante in, J.L. Means, T. Kucak, D.D. Crecar, Environm.
Pollut. (series B), ! (1980) 45. 4) W.A. Aue, C.R. Hastings, K.O. Gerhardt, J.O. Pierce, II, H.H.
Hill, RjFj Moseman, J. Chromatogrj, 72 (1972) 259. 5) C. Schaffner, W. Giger, J. Chromatogr., 312 (1984) 413.
-118 -
, ' 1> , .. ~ , ~
o , ~ , ~
Fig. 1 - Location of the sampling stations
TABLE I.
NTA CONCENTRATION (Mg/l) IN THE ANI ENE AND TIBER RIVERS IN DIFFERENT PERIODS OF THE YEAR 1987.
Sampling place * 9/6/1987 12/9/1987 26/9/1987
ANIENE P. Raccordo Anu1are (A) 0.2 0.5 0.5 P. Tiburtino (B) 0.6 10.4 1.8 P. Nomentano (C) 1.1 6.1 n.d. P. Sa1ario (D) 0.9 5.9 4.6
TIBER P. del Grillo (E) n.d. n.d. 0.1 P. Race. Anul. Nord (F) " 0.8 0.4 P. Olimpica (G) " 2.1 1.2 P. Race. Anul. Sud (H) " 3.4 2.2
(*) The location of the sampling station is shown in Fig. 1. n.d. = not determined
-119-
--, _. ~
! ~ I I i
~----~--·I--i ! l ' i]: - - - i
.d ; - .
! rf; , ! I i I - i
-*J.l .. :. 1 :
! I , . I I : -e~ ,.-~.- .. , ...
!
I I ,
NTA-
8 - .. - ~ ..
I.S.
" i NTA
- - -i
- -- A
- - M
--+-~S,-. ---'-I'~~I-'-' -----'-l:----: i
~ . : -,
, , . ~ ~ . ;-'
. H
.. , - -_.. .. . -.. _. , .. J : \ L
!) ' I ... i _ '-_.Ji-..... _-_.....J_ .... .::-...:::, ___ r-__ ~i __ :_,: _.Li..l( .. min 13 o 13 0 13 0
Fig. 2 - Gas chromatograms of distilled water sample (reference blank) (A) and with NTA added (B) and gas chromatogram of a river water sample (C). Fused silica capillary column (22m x 0.25 mm i.d.) with SE 54 bonded phase. Column temperature: 1600C; injector temperature: 180°C; detector temperature: 180°C. Carrier gas: nitrogen (flow rate: 5 ml/min). Splitting: 1: 8. Detector: NPD. Internal standar: 3,4-benzoquinoline (5 pg/l).
- 120-
HYDROCARBONS IN EAST MEDITERRANEAN SEA: DETERMINATION AND OCCURRENCE IN THE SEDIMENT OF CONSIDERED POLLUTED AND UNPOLLUTED AREAS OF
COASTAL ENVIRONMENT
Summary
M. PSATHAKI, M. ZOURARI and E. STEPHANOU Laboratory for Environmental Chemistry
Department of Chemistry University of Crete, IRAKLION - GREECE
Sediments collected from Iraklion gulf and Chania gulf (North Crete) were analyzed to assess the status of anthropogenic hydrocarbons input. Aliphatic, aromatic hydrocarbons and molecular markers such as sterane and hopane-type compounds were determined, by GC and GC/MS, for source identification.
1. INTRODUCTION A research program started in our laboratory aiming to study the trans
port of petroleum derived hydrocarbons, within a coastal environment, as well as their incorporation to benthic organisms. For that purpose, measurements of hydrocarbons in the air, water, suspended particles, sediments and aquatic biota, are planed in order to asses the input from tanker discharges, atrrospheric fallout and sewage effluents and also to evaluate the contamination and its consequences to this particular aquatic environment.
The first part of the above program concerns the analysis of sediments in order to identify in them, crude oil-derived hydrocarbons. Sediments are considered as pollutants sinks, because they offer an essential picture of events taking place in the water column. For this reason sediments have been used to study fossil fuel contamination of coastal environment (1,2).
Monitoring studies to identify petroleum derived hydrocarbons in r~ent marine sediments, concentrate on the analysis of hydrocarbon fractions for the presence of n-alkanes, steranes, hopanes and polyaromatic hydrocarbons (1,2,3).
In this paper we repport the application of the approach, of molecular characterization of sedimentary hydrocarbons for the assessment of oil pollution in one considered polluted area, namely the gulf of Iraklion and one unpolluted, the gulf of Chania. Both localities are in the north of the island of Crete. Iraklion is a city of more than 150.000 habitants characterized by a fast developpment. There is an important commercial harbour and an important fuel consuming electricity plant. In the gulf of Chania there is no any important polluting ac.tivity.
2. EXPERIMENTAL Samples. Sediment samples were collected between March and June 1987,
from Iraklion gulf and Chania gulf (north Crete) using a Smith-McIntyre grab sampler (4),in a depth of 30 m under the sea. Samples were picked with a corer from the middle of the grab (2-4 cm) and stored frozen at -20°C until analysis.
- 121-
Isolation and analysis. Hydrocarbon extraction was performed by ultrasonification of dryed sediments (24 h at 40-50 0 C) with (2:1) methylene chloride - methanol (4 x 30 rol). The extracts were reduced under pressure and sulfur was removed from neutral fractions by addition of activated copper. Hydrocarbons isolation was proceeded according to Giger et al (5). Briefly the mixture of hydrocarbons was chromatographied on Sephadex LH-20, with methanol-benzene (1:1). Two fractions were collected. The first fraction was further reduced in a rotatory evaporator and purified on fully activated silica gel. The first fraction, eluted with n-pentane contained the aliphatic hydrocarbons. The second Sephadex fraction was also purified on silica gel by elution with n-pentane followed by methylene chloride. The methylene chloride fraction contained the polyaromatic hydrocarbons.
A Finnigan mass spectrometer, Model 4000, with an INCOS 2000 data system was used for mass spectrometric measurements. The Carlo Erba 4160 gas chromatograph equiped with a Grob-type split-splitless injector, . contained a fused silica capillary column (SE-54, 25m x 0.25 mm) coupled directly to the ion source by a fused silica capillary. Helium was used for the carrier gas with a back pressure of 0.8 atm. The temperature program used was 70°C (1 min), 70-200 oC (10°C/min), 200-280 oc (2°C/min) and 280°C (20 min). The electron impact ionization mode conditions were the following: ionization energy 70 eV; ionizer temperature, 250°C; mass range, 35-590 m/z; scan time, 1,9 sec; electron multiplier voltage, 1700 V.
Quantitative analysis was performed on a Hewlett Packard gaz chromatograph, Model 5890, using the same chromatographic conditions as for the GC/MS analysis. For quantitative analysis l-chloro-hexadecane and 3,6 dimethylophenanthrene were used as internal standards.
3. RESULTS AND DISCUSSION Three sampling stations located in the gulf of Iraklion and two located
in the gulf of Chania were selected for analysis (Fig. 1). These sites were selected for analysis for the following reasons: Hl is a front of the electricity plant, H2 is in front of the commercial harbour and H3 is in front of the airport, where c4 is far from the city of Chania and C5 is in front of the city. Quantitative results are summarized in Table 1.
Figs 2 and 3 show, respectively reconstructed mass fragmentograms representative of the whole group of samples.
The aromatic hydrocarbon fraction isolated from the sediments contained a series of polycyclic aromatic hydrocarbons (PM), (Table 2, Flg. 2B) derived mainly from anthropogenic (combustion) sources (6).
The polluted nature of the Hl,2,3 sediments is evidenced in its saturated aliphatic hydrocarbon distribution (Fig. 2A) which, although still dominated by biogenically derived natural hydrocarbons, showed a slight baseline hump of unresolved components characteristic of oil polluted sediments. Important confirmation of petroleum contamination was obtained from the m/z 191 (Fig. 3A) and m/z 217 (Fig. 3B) mass fragmentograms. This is indicative of petroleum contributions to the sediments as it has been confirmed by the identification of petrogenic molecular markers (3). The steroidal hydrocarbons (m/z 217) essentially constituted by steranes and less by diasteranes. Such complex mixtures are typical for oil pollution. More definitive confirmation was obtained from the hopane distribution (m/z 191) which exhibited a series of extended C32-C35 hopane (22 S & 12) doublet characteristic of oil-derived hydrocarbons.
It seems that no important biodegradation occurs in the aliphatic
-122 -
station ~n-alkanes CPI ~UCM*(ppb) Total PAR (ppb) C19-C26 C2/-C35 (ppb)
H1 1063 1,2 3,0 1623 '11 ,3
H2 696,/ 1 ,5
I
4,1
I 1005 30,/
H3 404,6 1 ,6 2,3 583,3 43,4
c4 130,6 - 4,0 - -
C5 150,5 - 3,2 - -
* UCM: Unresolved complex mixture.
Table 1: Hydrocarbon composition of the sediments corresponding to the stations indicated on Fig. 1.
Peak No Structure assignment
1 Fluorenthene
2 Pyrene
3 Benzo(a)anthracene
4 Chrysene
5 Benzo(b+j+k)fluorenthene
6 Benzo(e)pyrene
/ Benzo(a)pyrene
8 Benzo(ghi)perylene
9 Indeno(1,2,3-cd)pyrene
10,11,12 Dibenzopyrenes
Table 2: PAR/s peak indentification represented in Fig. 2B.
I
I
I
I
composition of the hydrocarbons found in the Iraklion gulf sediments. This may be explained by the continuous input of petroleum products near the electricity plant (Hl) and the harbour (H2).
-123 -
Peak
A
B
C
D
E
F
G
H
I
J
K
L
M
N
o
Structure assignment
18a(H)-22,29,30-trisnorhopane
170(H)-22,29,30-trisnorhopane
17S(H)-22,29,30-trisnorhopane
170,21S(H)-30-trisnorhopane
17a,21S(H)-hopane
isomer of diploptene
not identified
17a,218(H)homohopane 22S
17a,21B(H)homohopane 22R) C32
17a21S(H) bishomohopane 22S
1'itl21S(H) bishomohopane 22R) C33
17a21S(H)-trishomohopane 22S
17021S(H)-trishomohopane 22R) C34
n021S(H)-tetraquishopane 22S
1{U21S(H)-tetraquishopane 22R) C35
Table 3: Peak identification of hopane hydrocarbons represented in Fig. 3A
4. CONCLUSIONS Using sediments as indicators for minitoring of petroleum pollution in
coastal environment is a useful approach. Qualitative and quantitative inforfmation obtained from the composition of aliphatic and aromatic hydrocarbons show that, for the whole area of Iraklion, this problem derives from the same type of fossil fuel products. Some other parameters such as total organic carbon composition, particle size distribution of sediments & ATP are needed in order to provide a better figure of the situation. These parameters are also planed to be measured in the very near future for the sediments ~alyzed for this study.
ACKNOWLEDGEMENTS This work is part of the "Primary Health Care and Nutricient Program"
under the scientific direction of Prof. M. Fioretos (Dept. of Social Medicine, Univ. of Crete) and supported by the EEC and the Ministry of Health an Social Care of Greece. We are gratefull to Dr. W. Giger (EAWAG) for allowing us to perform GC/MS analysis in his laboratory. We thank Professor's A. Eleftheriou group (Marine Biology Laboratory, Univ. of Crete) for sampling.
-124 -
REFERENCES
(1) M. DASTILLUNG and P. ALBRECHT, Mar. Pollut. Bull., 1,13 (1976). (2) A. G. DOUGLAS, P.B. HALL, B. BOWLER and P.F. V. WILLIAMS, Proc. R.
Soc. Edin., 80B, 113 (1981). (3) J. ALBAIGES and P. ALBRECHT, Int. J. Envir. Anal. Chem., 6, 13 (1979). (4) A. ELEFrHERIOU and N. A. HOLME in "Methods for the study for Marine
Benthos", p. 140, Ed. by N. A. Holme and A. D. Mc Intyre, 2nd Ed., 1984. (5) W. GIGER and C. SCHAFFNER, Anal. Chem., 50,243 (1978). (6) R. E. LAFLAMME and R. A. HITES, Geochim.-Cosmochim. Acta, 42, 289
(1978). -
_.
Figure 1 - Sampling stations (*) in North Crete.
-125 -
- ~ I
100
C31
®
C
..
Cn m
,z 85
1903
0
22,4
5 26
'00
~15
32'3
0 35
';5
min
lLm
/z 2
02
s .I 4
®
-X:,z
228
6 I jC
252 X,
z276
m
lz 3
02
ti
. 19
030
22,';
5 26
00
2901
5 3
2'3
0
35,';
5 m
in
Fig
. 2
-R
eco
nstr
ucte
d
mass
-fra
gm
en
tog
ram
s o
f n
-alk
an
es
(A,
m/z
8
5)
an
d
PA
H's
(B
, m
/z
20
2,
22
8,
25
2,
27
6,
30
2)
. P
eak
id
en
tific
ati
on
T
ab
le 2
.
100
®
22,';
5 2&
00
2915
100
®
22,4
5 26
'00
2915
DE
m/ z
191
F , G H 'I..J
·32,
30
32:3
0
K L
N
M
'0
35'4
5 39
'00
"'Yz
217
35:';
5 m
in
min
Fig
. 3
-R
eco
nstr
ucte
d m
ass
-fra
gm
en
tog
ram
s o
f h
op
an
es
(A,
m/z
1
91
an
d ste
ran
es
(B,
m/z
2
17
).
Peak
id
en
tifi
cati
on
T
ab
le 3
.
SESSION II
TRANSPORT OF ORGANIC MICROPOLLUTANTS IN THE AQUATIC ENVIRONMENT
Chairmen W. GIGER and A. MINDERHOUD
The Sandoz accident
Moni toring of the River Rhein - Experience gathered from accidental events in 1986
Predicting transport behaviour of organic pollutants using simple mathematical models
Fate and transport of organic compounds in rivers
Summary
THE SAIIDOZ ACCIDEIT
Bendicht HURNI Amt fur Umweltschutz und Energie
Kantons Basel-Landschaft, Liestal, Schweiz
The cause, course of events, and consequences of the Sandoz warehouse fire, Schweizerhalle, Switzerland, on 1st November 1986, are presented. The flux of matter to the atmosphere, to the soil, to the groundwater and to the Rhine River is discussed, and the chemicals released to the environment are characterized. Some aspects of the clean up operations and their success are also described.
1. Introduction On the 1st November 1986 the accident at Schweizerhalle hit the head
lines worldwide. Since this date, the name Schweizerhalle is a synonym for the risk and danger in the chemical industry, for chemical catastrophes and environmental pollution, as are Bophal or Tschernobyl. It reminds us, that our highly technological world is in ever increasing danger of manmade catastrophes.
2. The warehouse fire The warehouse number 956 was to be found at the southwestern end of
the Sandoz property in Schweizerhalle near Basle, Switzerland. It was about 90 m long, 50 m wide and divided in the middle by a fireproof wall. Originally this warehouse was built for machines and technical equipment. In 1980 it began to be used for the storage of agriculture chemical products. Naturally there was an automatic fire protection system, but as shown by the fire, this equipment was inadequate.
Some minutes after midnight on 1st November 1986 a fire was observed in this warehouse. Later on the reason for this fire was reconstructed with a high degree of accuracy. On Friday afternoon, several hours before the fire was observed, Sandoz employees were busy packing Berlinerblue in this warehouse. They were using a naked gas flame for the sealing of plastic shrink film wrapping. As a result of this work the Berlinerblue most probably caught light and smouldered unobserved for several hours. Berlinerblue is a light-resistant blue pigment used predominantly in the paper- and printing industry. (An association between Schweizerhalle and Tschernobyl can be noted here by the fact that Berlinerblue also is a very efficient antidote to the radioactive Caesium isotope 137).
A short time after midnight the smoulder had developed into a real fire. It was observed and alarm was given. The fire developed in a very short time into a spectacular firework. The fire was so big that foam extinguishers were not effective, even though this foam is designed for fighting against such a type of fire. Water was the only remaining possibility for fire fighting. For several hours the firebrigade fought against the fire with 400 litres of water per second. This large amount of water was taken from the Rhine River. The danger was thereby increased, because in the nearest neighbourhood of the warehouse 956 there was another ware-
- 128-
house containing metallic sodium which could under no circumstances come in contact with water. The protection of this sodium warehouse against the fire required a large volume of water, so the firebrigade played a real game of russian roulette.
At 3 a.m., in the surrounding communities and in the city of Basle, the catastrophe alarm was given. The people were instructed to stay indoors and to shut the windows. A horrible smell spread over a broad area of Basle and threatened to cause a paniC fear in the population. At dawn the fire was finally brought under control, and at 7 a.m. the end of alarm could be given.
3. The flux of matter, or the emissions 1350 t of agricultural chemicals were stored in the warehouse number
956. The warehouse was used very intensively, and Sandoz changed up to 30 t of the stock every week. It took about a fortnight until Sandoz was able to publish a final storelist. Anyhow, the crisis-staff already knew during the fire that mainly insecticides (phosphoric acid esters), and a large amount of urea, were stored here. The list was successively completed by Sandoz. As well as ecologically nonproblematic substances such as emulgators, stabilizers and solvents, very problematic substances also appeared on the list. In addition to 600 t of phosphoric acid esters the following products were of special interest: 8.6 t of an aqueous concentrate of a mercury compound, 1.5 t of phenyl mercury acetate, 2 t of the highly chlorinated insecticide Endosulfan, 2.3 t of the acaricide Tetradifon, and finally, 2.4 t of an ecologically nonproblematic but very intensive fluorescent dye, called Basazolred or Rhodamine. This dye was, and is still, very useful as a tracer in the contaminated terrestial and aquatic environment.
Host of the insecticides were phosphoric acid esters. They are effective by stopping the acetyl cholinesterase. These phosphoric acid esters have no longterm effect because they are chemically and biologically completely degradable. In agriculture, 200 - 500 g per 10 000 m2 of these insecticides provides adequate protection. They are also very toxic to warmblooded species, and their handling requires special precaution. Down to concentrations of 1 mg/l they are toxic to fish.
The organic mercury compound Tillex is a seed treatment product which has fallen into disrepute because it contains mercury. It is still accepted in many countries, even in Switzerland.
The chlorinated insecticide Endosulfan is extraordinarily toxic to fish. For many kinds of fish, the 100 % lethal concentration lies in the region of 1 ~g/l water. The half life in water may be rather long.
The chlorinated acarizide Tetradifon or Tedion has became well known because its chemical structure allows the formation of dioxines under special thermal conditions.
Under the effect of heat, water and oxygen, part of the stored agricultural chemicals underwent a change during the fire. In consideration of transport mechanisms these new products must be included. We distinguish thermolysis (chemical change caused by heat) the pyrolysis (chemical change caused through heat and oxygen) and the hydrolysis. The insecticides Disulfoton and Thiometon, for example,form a colourless oil in their pure form. They have very small vapor pressure and may be distilled only under vacuum. Under normal conditions they decompose below their theoretical boiling point, to form chemical products which may still have some biological activity. Pyrolysis leads to the formation of stable oxidation products as carbon dioxide, water, phosphate, sulfate and nitrate; i.e. to non-problematic substances. The hydrolysis of these phosphoric acid esters is very slow.
- 129-
During weeks or months the ester cleavage produces acids and alcohols, products without the cholinesterase stopping effect.
Chemicals were transported by air and water. Convection currents in the air above the fire, transported the substances to a height of 600 -800 m above ground. A gentle wind from the southeastern direction carried the products towards Basle City. The amount of airborne chemicals was not measurable. With the water used to extinguish the fire, the quenching water, a significant part of the liquid and solid agricultural chemical products were distributed in the area of this warehouse. Some of it was collected by the rainwater run-off drainage system and went directly into Rhine River. Some of it seeped underground contaminating the soil, and the groundwater which is found at a depth of about 14 m. As it later transpired the fire fighting boat pumped contaminated water from the river to fight the fire and protect surrounding buildings, during most of the night. Unwittingly they had spread the contamination to a very wide area. The mass balance is completed by the large amount of heterogenous waste at the site of the fire which was collected and is being stored under strict conditions in about 8000 steel tubes, 200 1 per tube.
4. The imissions Chemicals transported by the air were mainly decomposition products of
agricultural chemicals. Shocking, but also useful as a warning to the population, were the decomposition products of the thiophosphoric acid esters, especially the mercaptanes which are well known for their very intensive and unpleasant odour. In the first days the experts were very uncertain about the possible contamination of the atmosphere with merc~ry. As far as I know neither in the atmosphere, nor in the fallout could any significant mercury concentrations, resulting from this Sandoz accident, be found. Today we know that about 3/4 or more of the total mercury is still present at the site of the fire.
In addition, the formation of the very toxic chlorinated dibenzodioxines and fouranes may be excluded with a reasonable certainty. At the moment there are still some medical and chemical analytical longterm surveys being carried out. However it appears that airborne contamination never provided a serious threat to the population. We appear to have been quite fortunate. The situation looks much worse in the soil. An area of about 10 000 m2 are contaminated to a lesser or greater extent with these agricultural chemical compounds, down to several meters. The contaminated area has been determined by core and surface grab sampling at 40 sites. Several 10 000 m3 of soil are contaminated with about 100 kg of mercury compounds and with more then 1 t of insecticides.
At several points the contaminated quenching water percolated down to the groundwater, at a depth of 14 m. The groundwater system is large and is very important for the chemical industry, the surrounding communities and the city of Basle. By fast and clear instructions to water supply and artificial groundwater recharge plants in the neighbourhood, it was possible to prevent the distribution of the contamination in the groundwater, and to attract it to several groundwater wells near the site of the fire. These protective measures were a great success. However, as long as there are large amounts of these agricultural chemical products in the soil above the groundwater, the danger for the groundwater still exists.
As a result of the unintentional diversion of the quench water to the Rhine River, this river was contaminated, in an until now, unknown magnitude. More than 10 t of agricultural chemical substances entered in the Rhine River during a period of several hours. Contrary to the contamination of the air, the quench water contained mainly unaltered agricultural
- 130-
chemical compounds. This means, that biologically active substances, entered the river. A terrible damage to the aquatic fauna was the result. It took two months after the catastrophe to clear away sludge, rich in insecticides, from the bottom of the Rhine River until the emergency was over. With this cleaning action about 1000 kg of insecticides could be eliminated.
5. The clean up operation: the site, the soil, the groundwater and the Rhine River
With an enormous input of manpower and material, the site of fire was cleaned up. All the waste was separated as well as was possible to avoid additional problems with this anyhow very problematic waste.
The whole elimination of this waste has to be carried out by Sandoz, and the government is responsible for correct execution of this process. This also applies to the decontamination of the soil and protection of the groundwater.
The recolonisation of Rhine River is progressing well. Ten months after the catastrophe in the region of Basle, we again observe fish and small benthic animals such as larval insects, small freshwater shrimps, mussels and so on. However complete restoration of the river ecosystem will take several more years.
-131-
MONITORING OF THE RIVER RHEIN - EXPERIENCE GATHERED FROM
Summary:
ACCIDENTAL EVENTS IN 1986
Henning Friege Landesamt fUr Wasser und Abfall
Water quality of the river Rhein has been gradually improved in comparison to the situation in the 70ies. Thus more sensitive aquatic organisms are now settled in the water body and the banks of Rhein river. Therefore accidental events are now more important for the river possibly deteriorating the water quality. Even more important, drinking water supply may be endangered especially by those substances which are not eliminated in the purification process. In 1986, a number of contaminations of the Rhein water were reported, most of them caused by accidents in chemical production plants. The monitoring strategy for the Rhein in Nordrhein-Westfalen has proven to be succesful, the monitoring system should be enhanced.For rapid detection of contaminants in the river, monitoring activities include sensitive chemical screening methods as well as biological tests. In order to control waste water and cooling water discharge,automized analytical systems should be installed directly at the discharge points of a chemical production plant to detect irregularities as fast as possible. Moreover chemical companies should be enforced to install samplers working automatically. As decisions have to be taken very quickly, valid data on the behaviour of all possible and relevant contaminants with respect to raw water purification should be available on-line.
1. Introduction Water quality of the river Rhein has undoubtedly im
proved as compared to the early 70ies. This development is characterized in Table I by data from the border between Germany and the Netherlands (Rhein monitoring station at Kleve-Bimmen, Nordrhein-Westfalen). As to the biological analysis of benthos, German water quality class II (moderate load) from km 640 to km 700 and at the Niederrhein has been reached, water quality class II-III (critical load) is valid for the more industrialized regions between km 700 (Koln) and km 840 (Wesel). (1)
-132 -
I - '" '" I
Tab
le
I:
An
aly
tical
data
(9
0 P
erz.
,5=
75
% )
fo
r th
e
Rhe
in
(Kle
ve-B
imm
en,
km
865)
as
com
pare
d to
N
RW
's m
iniu
m
qu
ali
ty
req
uir
emen
ts
(90
Perz
.)
0,2
)
Min
imum
q
uali
ty
1970
19
78
1980
19
83
1986
re
qu
irem
ents
02-M
in.
(mg/
1)
3.4
3
.9
5.3
5
.3
6.5
>
4
C5B
(mg
/l)
35*
18
17
23
29
< 3
0
B5B
5 (m
g/l
) 8
.5
5.5
5
.4
5.9
3
.0
< 1
0
NH4-
N
(mg
/l)
3.5
1
.92
1
.41
0
.98
1
. 76
<
2
P (m
g/1)
1
.39
*
1.2
4
0.8
2
1.1
7
0.5
7
<
1
[OX
(1
19/1
) 15
5 61
40
77
AOX
(l1g
/1)
95
CHC1
3 16
41
2
.3
CC1 4
1
.5
6 0
.1
0.3
CH2C
I-C
H2C
l 8
.1
5.4
<
5
< 5
CH3-
CC
13
1.1
1
.6
< 0
.1
< 0
.1
CH
Cl
= C
C12
17
1.6
0
.2
0.1
CC1 2
=
CC1
2 1
.7
1.5
0
.2
0.3
Dic
hlo
rob
enze
nes
<
1.5
<
1.5
<
1.5
<
1.5
Hex
ach
loro
ben
zen
e 0
.05
0
.02
0
.06
0
.01
4-C
hlo
ro-3
-nit
ro-t
olu
en
e
0.3
1
< 0
.1
< 0
.1
< 0
.1
Ben
zo-a
-py
ren
e 0
.06
4
0.0
74
0
.07
1
*197
6
~
I
MES
STEL
LE:
DU
ESSE
LDO
RF
KM
734
PARA
MET
ER:
CHLO
RBE
NZO
L G
EWAE
SSER
: R
HEI
N
EIN
HE
IT
: U
G/L
280
240
200
160
120
80
40
* 0
.5
1986
* AN
AL Y
T I S
CHE
BE
ST! MMU
NGSG
R~~~
~
Fig
. 1
-N
orm
al
co
ncen
trati
on
o
f ch
loro
ben
zen
e
in
the riv
er R
hein
(~
3 ~g
/l)
co
mp
are
d to
le
vels
cau
sed
b
y accid
en
tal
ev
en
ts.
The general situation is more or less satisfying as a step in between, though the goals for some chemical parameters have not yet been attained or are not yet defined (see Table 1). (2)
If water quality has reached a certain degree, sudden release of hazardous contaminants becomes more and more important. This problem is illustrated in Fig. I showing the "normal" and the "accidental" concentration of Chlorobenzene in 1986. Depending on the nature and the concentration of the substance released, the drinking water supply miqht be endangered and/or the ecosystem might be damaged more or less serviously as it was shown for the Sandoz case. (3,4) Moni-toring the quality of surface water systems is normally restricted to a r atively low number of samples per year yielding a bunch of analytical data whichmay be interpreted in terms of distribution~free statistics (5). For example, the 90 perzentile at 75 % confidence level is gained from 13 analysis, whereas 7 samples are sufficient at the 50 % confidence level. Usually, the n~~er of analysis performed depends on the complexity'of analytic work. This
is especially true for organic micropollutants. To detect sudden release of contaminants, new monitoring strategies as well as better instrumentation for the control of waste water discharges are necessary.
2. Control and monitoring activities in Nordrhein-Westfalen In Nordrhein-Westfalen (NRW) waste water discharges as
well as surface water_are controlld carefully. Chemical industry, metallurgical plants, coal tar industry} and food industry are concentrated at the Rhein and some of its tributaries. About 4 million inhabitants of NRW obtain their drinking water from the Rhein purified by a sophisticated process (see below). Waste,water discharges are supervised by the water authorities the number of samples taken depending on the type of the effluent water. For example, samples from a small municipal waste water purification plant are taken twice or three times a year, whereas effluents of a large chemical production plant are examined more than twenty times a year without preliminary announcement by the controlling authorities. The monitoring system at the Rhein consists of three stations where samples are taken continously.
The instrumentation of the monitoring stations is documented in Fig. II. Two stations are situated at the borders of NRW, the third one neighbouring LWA's centr~l Laboratory at DUsseldorf. At normal flow (about 2.000 m /s at Ktiln) I a water molecule needs about 35 hrs. for the full distance. Numerous samples of water and sediment from the Rheim, the channels, and the tributaries are taken by a monitoring vessel.
3. Experience from accidental events in 1986
3.1 Detection, identification and guantification of micropollutants
Under the condition of the "chemical zoo" consisting
- 135-
• WKSt. • WassetkonUonSllllOtI
Monitoring Station: P.C. parameters l3-ray, I -ray monitors Biological tests DOC (cont.) Chlorid PolQrography (cont.) (GC-FID) (GC-PND) (GC- ECD) (Head-space-GC )
Sampling Station: (P.C. parameters) (Biological tests)
Central Laboratory : AES-ICP, AAS GC-FID GC-P 0 GC-ECO Head-space-GC HPLC DAD GC/MS a. -03 y, l3-ray, i -03 Y spectrometers Biological tests
110nitoring Station: P.C. parameters l3 - ray, ~-ray monitros Biological tests (cont.) Chloride (cont . ) DOC (cont . ) AAS GC-FID GC-PND (GC-ECD) (Head-space- GC)
Fig. 2 - Instrumentation of Monitoring Stations in 1987; further planning in brackets
-136 -
of about 80,000 substances sold on the market and an innumerable quantity of by-products and metabolites, analytical work is always restricted to certain "windows", e. g. g.1. c. with an electron capture detector to seize halogenated organic molecules oflow or medium polarity. If maximum information about the nature of the chemicals in a sample and their environmental behaviour is required, many "windows" are needed. In NRW, the monitoring strategy for the Rhein covers -continous measuring of physicochemical parameters (pH, conductivity, turbidity, oxygen),
-semicontinous analysis of chloride, some heavy metals, and daily analysis of a bunch of classical quality parameters,
-continous testing of water fleas (daphnia magna Straus) and goldenorfe ( leuciscus idus).
-continous adsorption of Rhein water by XAD resin followed by daily desorption with dichloromethane and g.l.c screening using flame ionization and P,N-specific detector to detect deviations from the normal range of background concentrations. (6)
Details my be taken from Fig. II. Samples from all three stations are stored for more than a week.
The combination of analytical screening techniques and biotests has proven to be useful. An example is documented in Fig. III. From Table II , one may take that the majority of accidental contaminations could be detected by chemical and/or biological screening tests either already installed or planned for the near future. Quite naturally, there are severe problems which have not yet been solved:
The continous tests with fish as well as with water fleas indicate more deviations from normality than proven by chemical analysis. Sometimes, there may occur toxic effects from substances which are overlooked in chemical analysis (e.g. organic molecules of high polarity, combination effects of substances in a low concentration range), as it has been already mentioned by §ontheimer. (7)
On the other hand, biot~sts c~nnot be "standardized" like usual analytical instrumentation, thus probably leading to misinterpretations. To open more "analytical windows", screening of molecules with high polarity is of interest. It may be taken from Table II that organic acids like 2,4-D can be detected by liquid chromatography. From our experienc~ high pressure liquid chromatography equipped with a detector scanning the u.v. region (array diod detector) should be suitable for this purpose; a special screening method for this instrumentation is projected.
Nevertheless, the combination of analytical screening instrumentation and biotests seems to become a powerful tool for the detection and identification of sudden contaminations in the river. The Dutch water authorities (DBW) are pursueing a similar strategy. (8)
-137 -
9 . 160 9.53;
-~C.=~~--------------- --------- -- ·
11.542
·~.37a
~~ : ~7 2 4 .26 ~4'5IB
-- 24.9~1 25.20;3
25 . 607 -- . 1-i. \i~
13.729
3 . 583 i . 909 111.201
4.574
- 8:~u .. - -------- ---- ----- --_ .
;:m 10 . C70
10.74111 11 . 00a
II. 514 , 1.9,,?01
ltm .. 3. 224
13 . 704 14.064
14.514
IUU ·IUH 16.603
17.010 '1.33'
17.649 17. : 30
19 . 305 '8.~9l
18 . 873
9'~~~99 19 . ns 20.152 0.559
20.B44 2i!~~~5 21 . 6:;a ~!:9n ~~.3a~ .. 72~ .• 17
Fig. 3 - Chromatograms of 24 hrs. composite Rheinwater samples by P,N-specific detector, "normal situation" (right) and pollution by Nitrobenzene at 4ppb (left) using XAD-4 resin.
-138 -
"-' ~
I
Tab
le II
: C
on
cen
trat
ion
s o
f so
me
chem
ical
s re
lease
d
by
accid
en
ts
in 1
986
in
the ri
ver
Rhe
in
vs,
an
aly
tical
cap
ab
ilit
ies
(H =
Hon
nef,
D
= D
Uss
eld
orf
, B
= Bi
mm
en)
Su
bst
ance
m
axim
um
Det
ecta
ble
by
Str
ikin
g acti
vit
ies
rele
ase
d
co
ncen
trati
on
G
C-F
ID
GC
-EC
D
GC-
PND
H
PLC-
DA
D
foun
d (I
l-g/
l)
from
X
AD
-res
in
or
by
neu
tral
a2
idic
o
f w
ater
o
f fi
sh
hea
d-s
pac
e an
aly
sis,
ex
tracti
on
fl
eas
2-M
eth
yl-
ch
ino
lin
8
(B)
? +
+
?
(Ch
inal
din
) N
,N-D
ieth
yl-
N-e
thy
l-N
-p
hen
yl-
6 (B
) +
+
+
+
eth
yle
ne-
diam
in
(An
izo
nb
ase)
1
,2-D
ich
lor-
33
(H)
+
+
(+)
eth
ane
18
(H)
+
+
(+)
Chl
orob
enz.
61e
87
(D)
+
? -
260
(D)
+
(+)
(+)
Dic
hlo
rob
en-
237
(D)
+
+
+
+
zen
es
Tri
chlo
rob
en-
8 (D
) ?
+
zen
es
Ben
zene
35
(D
) +
2
.4-D
ich
loro
-p
heno
x y
-ac
eti
c-
17
(H)
-+
?
acid
P
ho
ph
ori
cest
ers
17
(H
) +
+
+
C
hlo
rop
hen
ole
2
(B)
-(+
) N
itro
ben
zen
e 4
(H)
+
-+
-
+
AOX
+
3.2 Finding the source of pollution Very often, accidental release of hazardous chemicals
is not announced to the responsible authorities. A typical case is demonstrated in Fig. IV. Chlorobenzene at a level of 17 ~g/l (24 h composite sample) was detected by the Dutch monitoring station at Lobith. (Lobith is situated vis a vis to the German station at Kleve-Bimmen. Due to the laminar flow of the Niederrhein, the analytical results of the monitoring stations sometimes differ significantly, if chemicals have been launched downstream of Koln). It could be proven that the substance had been released between km 690 (Koln) and km 734 (DUsseldorf). As to be seen from Fig. IV, it was not possible to find the polluter . In a similar case happe~d in 19B7, the source could be identified by thorough analysis of samples from waste water discharges. The company responsible for the discharge was able to prove that the contaminant had already been in the water pumped out of the river. Later on, a ship was identified which had launched about 5 t of chlorobenzene by cleaning its tanks after unloading just in front"of th_e chemical production plant suspected.
Discharger should therefore be obliged to store waste water samples (e.g. 24 h composite samples) for a week. Moreover, the validity of these samples would be enhanced by a sampling system which cannot be manipulated by the company in question ("behordliche RUckstellproben"). A suitable sampler with 12 bottles (2 1 each) has been developed now being in examination. Each bottle can be stored at most for eleven days and will be cleaned automatically afterwards. The system is sealed. The sampler is equipped with a battery to avoid standstill at power failure. Moreover, a solution must be found for the documentation of samples from ships. As to complex chemical production plants, the companies m~y detect accidental release of hazardous substances to late. Supervision of cooling water discharge is often done by T.o.C. instruments continously working, in many cases there is lack of any suitable instrumentation. But even in the case of T.o.C. monitoring, significant amounts of chemicals may be overlooked. There-f ore, analytical control instruments should be in-stalled directly in the waste water or cooling water canal, reffJectively, i-f relevant amounts of hazardous compound s are ~sed in the plant in question. Bacteria test systems have been develop~d one of it by researchers of the chemical industry (9,10), which may be extremely useful for the detection of accidental release of toxic chemicals. If bacteria are not sensitive enough, daphnia is expected to be a suitable test organism, e.g. for insecticides.
3.3 Evaluation of dangers for drinking water supply About 20 million inhabitants in Germany and in the
Netherlands get their tap water from the Rhein. In Germany, bank filtered water ist used by the overwhelming majority of the water works. Bank filtration is normally followed by oxidation with ozone, filtration, adsorption by activated carbon, and neutralization of aggressive carbon dioxide.
-140 -
Fig. 4 - Contamination of the Rhein with Chlorobenzene (KSt = sampling point of water authorities; Einl. = waste water discharge by chemical industry. the sampling date is given under the line)
-141-
Though this procedure is extremely effective under "normal" conditions, increasing concentrations Q~ certain contaminants may cause risks for drinking water supply due to uncomplete degradation during purification or by damaging the biologicai activity of the bank. By the testing systems described above, potential dangers for two levels of the aquatic food chain will be detected. Properties of unusual contaminants with respect to the water purification process may be assessed by test filter systems. (11) It was demonstrated that the most important phosphoric esters from the Sandoz accident were eliminated in the river bank at an overall concentration of 10 ~g/l (12). But htis result was yielded afterwards; furthermore, five- or tenfold higher concentrations would probably have damaged the capability of self puri fication.
At the time of the Sandoz accident, data for the elimination of the compounds released were not available with the exception of Parathion. But even in this case, data were contradictory (l~). Besides the contaminants announced by the responsible company, severalmetabolites were identified (3). Probably, there have been many more break-down products.
Taking the Sandoz accident as an example, it is obvious that bank filtration should be stopped as a precaution to avoid risks for drinking water supply, if elimination data are not available for the compounds in question, or if proper assessment of the behaviour of these compounds is not ;possible.
4. Concluding remarks Since the early 70ies, the ecosystem of the river Rhein
is on the way of recovery, but the possibility of future danger remains because of accidents connected with a river used so intensively by industry, traffic and settlements. Therefore, the following measures are claimed for the future:
Monitoring activities should be enhanced. Obiously, it is not necessary to analyze each contaminant at each concentration level. Therefore, screening methods as well
as biological tests should be used to detect irregularities which may be followed up by more sophisticated methods.
Companies should be obliged to store waste water samples for a longer period to facilitate the search for a polluter, if necessary. Moreover, a sampler which cannot be manipulated from outside could be a tremendous incentive to avoid accidental release of toxic substances.
More data concerning the behaviour of chemicals in the water purification process (bank filtration, adsorption on active carbon)should be available immediately.
Water quality of the river Rhein must be improved gradually to regain a good ecological state. One should have in mind that this improvement depends on a variety of factors of which organic miropollutants are an important one; but "ecology" of a river means more than mere water quality. Therefore, general measures for the support of self-purification capacity and reactivation of former biotops (13) are also needed.
-142 -
References
(1) Landesamt fUr Wasser und Abfall NRW: GewassergUtebericht 1986, DUsseldorf 1987.
(2) Landesamt fUr Wasser und Abfall NRW: Weitergehende Anforderungen an Abwassereinleitungen in FlieBgewasser, DUsseldorf 1984
(3) Landesamt fUr Wasser und Abfall NRW: Brand bei Sandoz und Folgen fUr den Rhein in NRW, DUsseldorf 1986
(4) Deutsche Kommission zur Reinhaltung des Rheins: Deutscher Bericht zum Sandoz-Un fall mit MeBprogramm, 1986
(5) STOCK, H.D., KLtls, H., REINKE, H., ALBERTI, J. and ANNA, H.: Anwendung der verteilungsfreien Statistik und Informationstheorie auf die Kontrolle von AbwasserEinleitern, Chem. Ing. Techn. 2lL 679 (1979)
(6) ANNA, H., GORTZ, W. and ALBERTI, J : Konzeption und Aufbau einer zeitlich dichten RheinUberwachung, Vom Wasser ~ 93 - 105 (1985)
(7) SONTHEIMER, H.: Der Rhein im Jahr 1982. 9. Arbeitstagung der IAWR, KHln 1964
(8) RIWA (Samenwerkende Rijn-en Maaswaterleidingbedrijven): Jahresbericht '84 - Teil A: der Rhein, p. 74-77, Amsterdam 9185, and personal communications from DBW.
(9) PAGGA, U. and GUNTHNER, W.: The BASF toximeter, a helpful instrument to control and monitor biological waste water treatment plants. Wat. Sci. Technol. llL 233 - 238 (1981)
(10) PILZ, U. and AXT, G.: Weiterentwicklung eines kontinuierlich arbeitenden Bakterientoximeters Vom Wasser ~ 91 - 100 (1984)
(ll) SONTHEIMER, H. and VOLKER, E.: Beurteilung von Abwassern aus der Sicht der Trinkwasserversorgung. Z. Wasser-Abwasser-Forsch. l2L 9 - 13 (1986)
(12) ARW, Jahresbericht 1986, KHln 1987
(13) FRIEDRICH, G. and MULLER, D., Rhine, in B.A. WHITTON (Ed.), Ecology of European Rivers, Oxford 1984.
(14) FRIEGE, H., ALBERTI, J., GRUBERT, G., STOCK, H.D., STOCK, W., REUPERT, R., PLtlGER, E., WILLEMSEN, H.G., KNIE, J., JUNKE, 1. and HALKE, A.: RheinUberwachung und Einleiterkontrolle - Auswertung der StHrfalle im Jahr 1986, Vom Wasser, in press.
-143 -
PREDICl'ING'mANS:roRI' BEHAVIaJR OF CR:iANIC roLllJI'ANl'S USIK; SIMPIE ~CAL KXEIS
P.S. Griffioen* am D. van de Meent**
*Institute for Inlam Water Management am.waste Water Treabnent ro Box 17, 8200 AA Wystad, the Netherlarrls
**National Institute of PUblic Health am Envira1mental Hygiene ro Box 1, 3720 B1i. Biltb:wen, the Netherlarrls
SI.mI!my
'lW sillple JOOdels are presented for predictirg transport behaviour of pollutants, in case of accidents in a river. '!he first JOOdel is especially suitable for predictirg transport in the neigbcmilood of the accident, wle the seoorrl JOOdel can be used for la-ger distancies, incl~ inflow, a.rtflow am velocity variatiems. Relevant JOOdel parameters are discussed am attention is payed to the description of elimination processes. FUrt:hel:nore the effects are discussed of exc::hanJe processes, not acx::oonted for by the JOOdels.
1 INmOI::UCl'ION Accidents in am alcn;J a river are generated by a rumtler of
causes, such as leakages, fires am 1:x:Iat accidents. As a consequenc:e, various types of dlemicals can be spilled wdl pose a threat to aquatic life am plblic water Slg)ly. In order to take effective am save ooonter measures am in order to infonn the plblic, predictiems are necessary aboot transport behaviour am e>!peCted oc:n::ent:ration levels. For that p1I'pose JOOdels can be used. B.lt still the prediction of transport am ooncentration levels is not an easy task. A rumtler of p,ysical, chemical am biolc~ical processes detenni.ne the behaviour of chemicals in a river. )ok)reo\Ter, lack of infonnation aboot the actual situation limits the accuracy of JOOdel predictiems.
'!his paper presents two JOOdels for predictin;J transport am oc:n::ent:ration, in case of accidents in a river system. Together they are a usefull tool in actual situatiCllS. '!he min feature of the JOOdels is their sillplicity. Ebth JOOdels are analytic expressiems am therefore easy to ilIplemern on a persooal carp.rter am fast in use, even for large river systens.
'!he JOOdels are presented in the next dlapter. Primary, attentien is focussed en the description of mi.xinJ am transport. Specific prqJerties of the spilled pollutant are taken together in an overall first order elimination process, incl~ adsoIIr tion am subsequent sediJrentation. A J1Dre detailled description of the eliminatien processes is given in dlapter 3, together with a discussion aboot dispersion coefficients. Because
-144-
a'lly elimination prooesses are oonsidered, the effect of exd'lan:Je prooesses, irx::ludin:J resuspension, is briefly diaJSSed in chapter 4.
2 MIXING AND 'lRANSKRl' OF roIWrANIS IN RIVER SYS'lDlS Spreadirg of spilled pollutants is dlaracterized by two
p:rooesses namely advectioo am dispersioo. Mvectioo is the transport of pollutants with the main flow: they are transported with a velocity equal to the cross-se=tiCX1al averaged velocity. Dispersion is mainly caused by velocity am OCI'Celltration variatialS CNer the cross-se=ticn due to fricticn losses. M:Jstl.y, dispersicn caused by tuJ:bll.eroe is of a lower order of magnit::OOe. '!he effect of dispersioo is a decrease in cxn:leI1traticn gradients.
In general, spreadiIg of spilled pollutants passes three pmses. In the first {ilase transport is daninated by lIlixin} CNer the depth, Wle the secxni {ilase is daninated by lIlixin} CNer the cross-se=ticn (lateral dispersioo). In the third {ilase, lagitudinal lIlixin} detemines the cxn:leI1traticn profiles.
2.1 IATERAL DISPERSICN If the depth of the river is relative small with respect to
the width of the river (order 20 or lOOre), the first {ilase is neglegible in the sense that it takes place instarrt:enea.l. '1herefore, initial spread.in:J can be described by a two-dimensional lOOdel., t:aki.n.J into aocoont lateral am lorgitudinal dispersioo am advecticn.
'!he distaIre fran the acx:ident to the locaticn downriver where the pollutants are well mixed CHer the cross-secticn, is called the lIlixin} lergth. Deperdin:J 00 the situaticn, the lIlixin} lergth can be So-100 km. '1herefore, the mass distrib..Iticn CNer brarx::hes may be different fran the clisdlarge distrib..Ition. '1his is inp:>rtant infonnatioo for further calculations rut also for 0CA.II'lter measures.
In this secticn a lOOdel. is presented for the description of the lateral Plase. Consider the situation depicted in the figure below:
x=O t -vy-=Q
d-- * y
b
->x
->u
* = location of the acx:ident
Here X am y are resp. the lorgitudinal am lateral coonlinate, U the average flow velocity, oonstant in the neiglll:nlrhood of the accident am b is the (mean) width of the river. ~ that the acx:ident took place at a distaIre d fran me of the banks (see figure) am that the spill took place at time t=o. ~ further that the pollutants are totally reflected at the river banks. '!hen the depth averaged cxn:leI1traticn C(x, y, t) at a distance x fran the accident, SCI'llE!It.bere at a point y (~ySb) in the cross-section atrl at time t, is given by:
-145 -
IVh (x-ut) 2
(1) C(x,y,t) exp[-kt] exp[- -- ] 4rt/DxDy 40xt
GO (y-2nb+d) 2 (y-2nb-d) 2
* E (exp[- 1 + exp[- J) ~ 4Dyt 4Dyt
t.ilere M is the mass spilled into the river, h is the (mean) water depth am Ox am Dv are resp. the lcn;Jitudinal am lateral dispersioo coefficients. 'Dle term exp[-kt] stams for first otder elilninatioo processes am k is the first otder elilninatioo rate constant. 'Dle SUllllBtiOO aoc:nmts for reflectioo at the river banks. Because of the nature of the expcI'letIts ooly a few SUllllBtioos are relevant in acblal. situatioos (fran -10 to +10).
Based 00 the above equatioo, it can be shcMl (Van Mazijk, 1987) that the mixiJ'g lergt:h L is ~tely given by:
L 11:1 0.4 u1j lOy For distarx:ies greater than L the ~tioo profile is uniform C:Ner the cross-sectioo.
Figure 1 shows sane results ootained with equatioo (1). With the given parameter values, the lIIi.xinq lergth. L is aboot 54 Jan.
- '1'-0 m
\1"'1.0 (nVs) M=100 (kg)
tn.N
, .. " .. V- 200 "'
D,r200 (m'/s), ~.3 (m'/s) tF4 (m), b=200 (m) d=O , k=()
"" ...... .. .
, , , , ,
Distance from the barl< (m)
Figure 1. Sane results dJtained with equatioo (1;.
2 • 2 em: DIMENSlOOAL 'ffiANSroRl'
- X-5 km
- - . X- 20 km
....... X"' 55 k.m
At distarx:ies greater then the lIIi.xinq lergth., the pollutant is well mixed CNer the cross-sectioo. '!hereafter, transport can be described by a cne-dilnensiooal nw:xiel, taJd.rg into ac:x:nmt ooly advectioo am lcn;Jit:utinal ctispersioo.
In this sectioo three equatioos are presented for predictirq the cross-sectiooal averaged oonoent:ratioo cbrm river fran an
- 146-
aocident. All three equatioos are based 00 stationary (that is time iroeperrlant) hydrologic oc:ntitioos. 'lhey differ, however, with respect to the assunpticns aJ:x::ut velocity am cross-sectional variatioos alcrg the river, namely:
9 constant velocity am cross-sectioo alcrg the river. 12 constant velocity am varyirg cross-sectioo. g varyirg velocity am cross-sectioo.
9 If flow velocity am cross-sectioo are' constant (00 inflow), the cxnoentratioo C(x, t) at a distance x am time t, is given by:
MIA (x-utp (2a) C(x,t) = -- e:xp[- --] e:xp[-kt]
2./7rDt 4Dt
where M is the mass spilled, A am u resp. the cross-sectioo am flow velocity, D the lcrgitu:iinal dispersioo coefficient am k the first order eliminatioo rate constant.
12 If inflow ally effects the cross-sectioo bIt rot the flow velocity, then C (x, t) is given by:
Ml1.o 00 (x-ut} 2
(2b) C(x,t) = -- -- e:xp[- --] e:xp[-kt] 2./7rDt Q(x) 4Dt
where 1.0 am 00 resp. are the cross-section am d.isd1arge at the locatioo of the spill (x=O) am Q(x) the d.isd1arge at a distance x. '!he factor QdQ(x) is a dillutioo factor. '!he inflowirg water has concentratioo zero. Notice that (2a) is a special case of (2b) .
g If both the velocity am cross-sectioo vary in down river direction, thenC(x, t) is given by:
MlA(x) (u(x) (T(X)_t)}2 (2c) C(x,t) = -- e:xp[- ] e:xp[-kt]
2j7rDtc 4Dtc
Here, A(x) am u(x) are resp. cross-sec::tion am velocity at a distance x. T(x) is the time needed to travel alcrg with the main flow fran the locatioo of the spill CNer a distance x. 'Iherefore, T(x) is called the travel time. If the flow velocity is constant, the travel tine satisfies:
x = u T(x) whien means that the travel time is sinply equal to distance divided by velocity (c:::c:IIpare with (2a) am (2b». In real situatioos, the river may be subdivided into stretches with constant hydrologic dlaracteristics. '!hen the travel time aver a stretch is equal to the lergth of the st:ret.d1, divided by the velocity. 'l1ll:'c:u3h sunrnation, the total travel time is c:btained between each two points alorg the river. '!he factor tc in (2c) acx:amts for flow variatioos. If there are 00 variatioos, then tc is equal to t. In case of variatioos fran st:ret.d1 to stretch, a oorrectioo term nust be ad:led, eaen time a variation is passed. '!he oorrection tents are equal to:
[{u(X)/Ui}2 - 1] Ti where ui am Ti are resp. velocity am travel tine on stretdl i.
- 147-
Notice again that (2.b) is a special case of (2.c). Figures 2 am 3 shclw sane results obtained with equation
(2). In figure 2, velocity am discharge are 0CI'lStant. Figure 3 shcM; the maximJm ccn::art:ratien when travellirr;J damriver. '!here is inflow at x=50 kIn.
Remarks - Equations (1) am (2) are based en .inst:anteneoos spills.
Extension to spills of finite duratien is rather straightforward by suitable SUlllllatien (in fact integratien) of the equations.
- If the spilled pollutant floats en water, there is no mixin;J rNer: the depth. In that case, the results of equations (1) am (2) shoold be JIIlltiplied by the depth h, in order to cirt:ain surface cxn::Srt:ratien (expressed as mass per writ of area). M::Ireover, average surface velocities shoold be used instead of cross-secticnal averaged velocities.
- Aooord:in] to Eql8.tien (2), the travel time is equal to the time at which the centre of the pollutien passes a fixed locatien. Qle lIllSt realize that this is rot equal to the JOCIIIel'lt at which the maxinum ccn::art:ratien is dJserved at that location. rue to dispersion am elilninatien, the maxim.Im conoentration is dJserved sane time before the centre passes. Hc7Never, in general this time differeD:le is neqle:Jible.
- If equatien (1) is integrated rNer: the width of the river am the result devided by the width, equation (2a) is ootainerl.
,moW
-K-o.O --_. K:cQ2' lid!-
Figure 2. CorX:lentration profiles at distancies x=55 kIn am x=100 km am k=O am k=O.2 (lId)-
·.~----~u~----~----~----~
- ~=O.O
Figure 3. Maxinum concentration when travelliIg damriver.
- 148-
3 DISPERSlOO COEFFICIENl'S AND ELIMlNATIOO ~ When usirg the m:xie1s ale DUSt specify dispersion coeffi
cients am the el:i:mination rate 0CI'lStant. '!he latter is depenjant on the type of P:r<:X:leSses relevant for that particular pollutant.
3.1 DISPERSlOO COEFFICIENl'S Dispersion coefficients are in fact m:xie1 ~ in the
sense that they are the resultants of P:r<:X:leSses rot acx:nmt:ed for by the m:xie1. '1herefore, the value of a dispersion coefficient shoold be seen in the cxntext of the m:xie1.
'!he value of the dispersim coefficient varies fran river system to river system. Because of di.sd1arge variations am dlarges in bottan rc::u;jmess, it can even vary with time am locatim. Values reported in the literature for the lorgitudinal dispersim coefficient vary fran aboot 10 (m> Is) for SDDOth dlannel. flow, up to aboot 700 (m> Is) for roogh river flow.
In this section a semi -enpirical expressim is presentai whien relates the dispersion coefficient to bottan rc::u;jmess am flow velocity. '!his equation can be used for extrapolation am int.etpolation urrler varioos oorrlitions, rut because of it's E!1pirical nature one shoold first calibrate the expression with in situ experin'ents.
1II::xxlrdirg to Fisdler (1979), the dispersion coefficient D can be written as:
D = Q h u* ~ Q is a dimensionless prqx>rtionality constant, h the water depth am u* the shear velocity, whien is related to the water velocity u accordirg to:
u* = u Jg/C Here, 9 is the acx::eleration of gravity, am C the Olezy-ooefficient whien is deperrlant on the rc::u;jmess of the river bottan. C can be calculated with the fo:rnul.ae of Colebrook-Ni.kuradse:
C = 18 log{l2h/d} (mVs) ~ d is a rc::u;jmess measure. For flat bottans d is equal to the particle dianeter; for river bottans with riWles d is approxilllately equal to one or ONe tines the height of the riWles.
'!he above expressions can be used for the lorgitudinal as well as the lateral dispersion coefficient, only the value of Q
is different. Values reported in the literature differ very lII.ldl. For lCDJitudinal dispersion the experimental detennined value of Q rarges fran aboot 20 to 500 (Fisdler, 1979). '!he value for lateral dispersim is aboot a factor 40 to 100 smaller. Clearly, it is advisable to perfonn in situ experin'ents urrler varioos oorrlitions am to evaluate the dispersim coefficient (or the prqx>rtionality 0CI'lStant) within the context of the m:xie1. '!hat is, one has to oarpare measured am calculated concentration levels.
3.2 RIJMINATIOO ~ Pollutants spilled into a river are subject to Iilase trans
fer ani transfonnatim P:r<:X:leSses. In this section it is examined whete.r these P:r<:X:leSses can be described by an overall first order el:i:minatim process, as is suwosed in equations (1) am (2). Furt.her1rore, the overall el:i:mination rate constant is
-149 -
related to DDre specific process parameters, althalgh realizin:} that this doesn't solve the prOOlem of ootainin;J reliable values. '!he follCMin} rotatioos will be used:
ct : total pollutant CXD:leJltration in water (massjvolume) Oi : CXD:leJltration of pollutant in dissolved fODll (mass/volume) s : particle ocncelltration in water (massjvolume) Cp CXD:leJltration of pollutant in adsomed form in the water
(mass per lD'lit of particle mass) (J) : ~tial of pollutant in adsomed fOnD at the bottan
(mass per lD'lit of area) It is assumed that Oi am Cp are in equilibrium, thus:
Cp/Oi = I<'p where I<'p is the particle-water partitial ooefficient (volume per lD'lit of particle mass). F\n:'thernr:)re it is assumed that settled polluted particles dal't exdlarge with pore water.
water-bottgn transfer a) Sedimentatial
It is easy to see that the pollutant sed:llnentatial flux (mass per lD'lit of area per tmit of time) equals:
sed-flux = - v S Cp = - v S Fp ct with Fp=I<'p/(1+SI<'p) am v the (mean) particle settlin:} velocity. 'lhus if the cx:noentration in equation (1) am (2) is interpreted as the total cx:noentration ct, then the first order eliInination rate constant for sed:llnentation is given by:
k-sed = (v S Fp)/hm where 11m is the thickness of the layer over wen the particles are well mixed.
b) Resuspension If all settled particles have the same prOOability to
resusperrl, then the mean resuspension flux of polluted particles is prc:portional to the raII'Itler of settled polluted particles. '!he prc:portionality constant (rotation: P, l/P is the mean resi.den:le time on the bottan between settlin:} am resuspension) depen:ls C8'1
the shear velocity u* (see 3.1) am the relative density 6 of the particles. A ocmoon ~d1 is to take the follCMin} depenierx::y:
f = q (U*/llo - 1)/6 if u* ~ 1Io* else P = 0 where 1Io is the critical shear velocity am q an enpirical constant (l/t). '!he value of 1Io * is depen:Jant 00 the type of bottan am ran;JeS fran 0.15 (1o/s) for rtu;Jh bott.cm; to 0.005 (1o/s) for DDre flat bottans.
It follows that the pollutant resuspensioo flux equals: res-flux = P (J)
Because (J) is the integral result of sed:llnentatioo am resuspensial, it's nat possible to express (J) as a linear fwrtioo of ct, both at the same time. Only un:ler steady state ocn:liticns this is possible b.1t lucky, accidents are no statiooary ~e. 'lhus resuspension can't be taken into acxnmt by a first order eliminatioo process for ct.
water-air transfer: volatization By the classical dooble film theoty, the flux of pollutants
thrtu;Jh the water-air interface is equal to: vol-flux = vtot (P/H - Oi)
where P is the partial pressure of the pollutant in air am H
-150 -
Henry's constant. '1he overall mass transfer coefficient vtot is given by:
ljvtot = ljvw + RI'/ (H va) where T is the absolute teuperature, R the gas constant ani vw ani va resp. the mass transfer coefficients in water ani air (lergth per unit of time). '1hese latter are related to mlecular diffusion ani to the shear velocity at the water-air interface Widl is oormally a power :fun::tion of the wind speed at sane reference height. If PfH is small cx::mpared to Q:l (the accident took place in the river), then the volatization flux can be written as:
vol-flux = - vtot Ftl ct with F'd=lj(l+SRp). '1hus the first order elllnination constant for volatization becaoes:
k-vol = (vtot Ftl) /h
Transfonnation Prooesses Because of the wide variety of transfonnation processes a'lly
sane general remarks will be made. Transfonnation processes act on both the adsorbed ani dis
solved fraction. If these processes can be described prcperly by a (pseu:lo) first order overall transfonnation process, then the overall process rate constant for transfonnation is:
k-transf = rd Ftl + l:p S 11> (dissolved) (sorbed)
where both rd ani rp may be the sum of varioos (pseu:lo) first order process constants.
In general transfonnation processes deperrl on envi.rornnental factors ani therefore also rd ani l:p. For instance, Plotolysis ciepenjs on irradiation at the water surface ani extinction in water, together with extinction pIqlerties of the pollutant. Biotic transfonnation processes deperrl on bacterial activity resultirg in process constants Widl deperrl on bacterial CXII'lOE!l'l
'i::.ration. '!he general ClWroadl is to express the first order rate
constants as the product of a specific, CCIlpOUl'l:i ani process depen:1ant constant with a factor representirg envi.rornnental effects: rd(p) = rd(p)o f(T,IiI,bacteria,light, ... ).
'1he table below gives an overview of the results of this section.
flux k
sediJoontation -v11>sct v 11> Sfhm
resuspension f3Q)
volatization - vtot Ftl ct vtot Fd/h
transfonnation rd Ftl + l:p S 11>
- 151-
4 EXCliANGE FROCESSES: ST1\GNANI' ZONES AND RESUSPENSION In the previcus dlapter it was shorm that resuspension can't
be covered by the lOOdel.s. '!he same holds for the exc:h.an:Je with stagnant zones present in the river am caused by irregularities or small dams briJ.d in order to regulate the water fla.r.r (typical for the dutdl situatioo). M1en passin] these zooes, mass is transferred thrcu;Jh lateral dispersioo am given back by the same mechanism. '!he bottan is in fact also a stagnant zcne, ally the exc:h.an:Je is regulated by two different type of prooesses.
It is }XlSSible to incorporate these processes, b.rt: then no larger an analytic sollutioo is }XlSSible, wdl means that a rumerical solution nust be SCU]ht.
In this dlapt:er ally the results of sane investigatioos are presented, wdl clarify the effect of exc:h.an:Je prooesses, t0gether with sane rules for assessin] the effects. '!he investigatiens are reported elsa.here (Griffioen, 1987).
Figure 4 shows the difference between transport with am wit:hoot stagnant zooes. '!he result wit:hoot stagnant zones is ootained with equation (2). Clearly, stagnant zones lower the peak ClCI'Dmtratioo, cause tailin] am delay the nanent at wdl the maxim..nn ClCI'Dmtration is d:lseIved. In fact these effects are already known in dlranat:.ograply.
Figure 4. An illustration of the effects of stagnant zones.
'!he maxim..nn delay is deperrlant on the exc::harge rate. It can be shown that the maxim..nn delay is achieved for extreme exchan:Je rates am that the maxim..nn delay is given by:
f/(l-f) T(x) where T(x) is the travel time (see (2.2» am f the fraction of the total cross-section, considered to be stagnant. '!he maximum delay is thus a fraction of the travel tine.
For bottan exdlarge, the situatioo is less trivial because of the two different processes. If the resuspension rate is large (P large: la.r.r particle densities, high shear velocities), the maxim..nn delay can be assesed by:
T(x) (v Fp s/hm)/P which means that for high resuspension rates the delay is zero. If however P is small (high densities, la.r.r shear velocities), the maxiJrum delay is proportional to:
T(X) P v Fp S/hm expressin] the fact that for fj=(J there is also no delay (al thoogh
- 152-
sane mass will never pass). F\rt:he.ntDre it is seen that the delay is inversely prqx>rtional to the thickness 11m of the mi.xin;J layer.
5 ax:tI5IOO 'l\«) analytical m::ldels are presented for predictiRJ transport
behavic:m' of pollutants in case of accidents. '1lle main featw:e is their sblplicity bIt still a variety of hydrologic oorxlitialS can be 0CJYered ani several eliJDinatioin processes can be taken into accamt. Also st:ret:ci1in;J ani tailirg by lateral ani lCDjimtinal dispersiat can be described ani the effects of stagnant ZaleS can partly be assesed by sblple rules.
'!he limitatialS lie in the fact that stationary hydrologic oorxlitialS are asslJmed.
Fischer, H.B. et al (1979). Mixin:.J in Inlani ani Coastal waters. Academic Press, New Ym.
Griffioen, P.S. (1987). '!he differen::Je between travel time ani transport time: the effect of stagnant ZaleS. To awear. Mazijk, A. van (1987). Verspreidirg van stoffen in de Rijn en de CXlI'1SE!C)Ileles voor de representativiteit van meet:t:;unten. H2O (20) 1987, nr 9.
-153 -
FATE AND TRANSPORT OF ORGANIC COMPOUNDS IN RIVERS
1) Selection of study site and validation of sampling and analytical procedures.
C.D. YATTS and K. MOORE YRc Environment, Medmenham Laboratory, Henley Road, Medmenham,
PO Box 16, Marlow, Bucks, SL7 2HD.
Summary
A study of the transport and fate of organic pollutants in rivers has been initiated by the selection of a site for study from among seven candidates. Analysis was carried out on the water, suspended particulates and sediments at the selected site to assess the representativeness of the sampling procedures and determine a range of compounds suitable for the detailed transport and fate study. The selected compounds were mostly anthropogenic in origin and included several potential EC list 1 and related chemicals. Calculations were carried out to determine the analytical requirements for the detailed study and the proposed research programme is outlined.
1. INTRODUCTION Yithin the aquatic environment the transport and fate of organic
pollutants is thought to be largely dependent on their physicochemical properties. Pollutants are subject to a range of physical and chemical processes which determine their ultimate distribution in the aquatic environment. These are illustrated schematically in Figure 1.
Sorption may occur to both suspended particles (1) and bottom sediments (2, 3), the former maintaining the pollutant within the water column, while the latter may effectively immobilise it. There may also be some association with dissolved organic material (4) present in the water body, effectively increasing the "solubility" of the pollutant in water. Uptake of pollutants by the biota living within the water column may result in bioaccumulation and hence transfer of the pollutants into the food chain. Loss of pollutants may occur by volatilisation from the surface of the water body (5), or by transformation via a number of chemical and biochemical routes. In some cases the product of such reactions may be of greater concern than the precursor compounds, but in others the end result of these transformations may be that the compound is completely mineralised. Such transformations may occur by indirect (6) and direct photolysis (7), oxidation, hydrolysis (8) and microbial biotransformation (9).
A knowledge of the detailed effects of these processes on the transport and fate of organic pollutants in water bodies is essential to a realistic evaluation of the potential effects on downstream biota and drinking water supplies. Such" knowledge can assist in providing an objective and rational framework for decisions on chemical use and disposal.
Several approaches have been made to the problem of obtaining information on these complex phenomena, including the laboratory
- 154-
measurement of partition coefficients (1-3) and the construction of environmental models, for example, Q~ASI (10), TOXFATE (11) and EXAMS (12). A detailed review of the approaches to design of mathematical models of aquatic environments was presented by Burns (13). It is important to remember that the refinement of such models ultimately depends on the measurement of pollutants in the real environment (14). A third approach is to select a range of organic pollutants and measure the concentrations of these compounds in the various environmental phases in a real aquatic environment. Potential problems with this approach include obtaining representative samples of the environmental phases for analysis and developing analytical methods of sufficient sensitivity and specificity. Consequently, relatively few reports have appeared in the scientific literature which have adopted this approach. One study of this type (15, 16) used the effluent from a small speciality chemicals manufacturing plant as a point source discharging organic pollutants to a river. Samples of the effluent, receiving water and river sediments were analysed at a number of points downstream of the discharge and showed the presence of a range of compounds with widely differing physicochemical properties. It was found that the aqueous concentrations of these compounds followed the rules of simple dilution and that those compounds with the highest octanol/water partition coefficients (log Kow) were strongly associated with the particulate matter in the water and were found in the sediment at the greatest distance from the discharge. A similarly detailed study was carried out for linear alkylbenzene sulphonates (LAS) using a municipal sewage effluent as a point source (10). LAS was found to have a very short life in this environment due to rapid biodegradation in water and it was shown that the results obtained from measurements in a real environment agreed quite closely with those predicted from the Q~ASI fugacity model.
Both of the examples given above demonstrate that useful information can be obtained on the fate and transport of pollutants in rivers from measurements on real aquatic environments and that these results can be enhanced by comparison with predictions from mathematical models using physicochemical parameters. The study reported here was conceived with the objective of examining the fate and transport of organic pollutants in rivers using a similar approach to that adopted by Lopez-Avila and Hites (15) above. A short list of seven potential sites for study was drawn up and these were examined chemically to allow final selection of the study site. Field and laboratory studies were carried out on the selected site to allow selection of the range of compounds which would be studied and for which analytical methods would be validated.
2. EXPERIMENTAL
2.1 Sampling Assessment of the seven potential study sites involved collection of
manual composite wa~er samples (5 litres) into pre-cleaned all glass bottles. These were collected over a period of 2 hrs from the river upstream of the effluent discharge and from the sewage treatment works (ST~) effluent. The samples were filtered (~hatman 451 paper; 20-25 ~ particle retention) on site, acidified to pH2 with conc. HCI and dichloromethane (250 ml) added.
For the detailed study of the selected site, manual composite water samples (5 litres) were collected over a period of two hours from the river upstream and downstream of the effluent discharge and from the ST~ effluent. The water samples were acidified to pH2 with cone. HCI on site and pressure filtered through a glass fibre filter (15.0 em Vhatman GFF; 0.7 ~ particle retention; previously cleaned by heating for 6 hrs at
-155-
450°C) immediately on return to the laboratory. The filter papers and associated particulate matter were stored at -20°C prior to extraction and the filtered water samples at 4°C. Surficial river sediment samples were collected from near both banks and in midstream at points above and below the effluent discharge. Sediments were sieved through a 2 mm stainless steel screen and transferred to pre-cleaned wide neck glass jars with screw tops lined with aluminium foil for transport to the laboratory in a cold-box. On arrival at the laboratory the sediment samples were stored at -20°C prior to analysis.
2.2 Extraction Water samples (acidified to pH 2 on site) were extracted after
addition of a spiking mixture to give a concentration of 100 ng 1-1 of the following compounds: d3-1,1,1-trichloroethane, d6-benzene, d5-chlorobenzene, dl0-p-xylene, d5-phenol, d8-naphthalene, d34-hexadecane and dl0-phenanthrene, by stirring overnight with dichloromethane (250 ml). The solvent was removed using a separating funnel and dried by freezing out the water. The dry extract was concentrated (to -5 ml) using a Kuderna-Danish apparatus with a macro-Snyder column and to a volume of about 1 ml using a micro-Snyder column and a concentrator tube (10 ml). Further concentration was carried out, if necessary, using a stream of dry, pure nitrogen. Selected water samples were then basified to pH 12 with 6N aqueous potassium hydroxide solution and extracted again by stirring overnight with dichloromethane (250 ml). The solvent extract was concentrated as described previously.
Glass fibre filter papers with associated suspended particles were Soxhlet extracted overnight with dichloromethane (250 ml). The solvent extract obtained was then processed as described previously.
Frozen sediment samples were allowed to thaw at ambient temperature, centrifuged, the supernatant water decanted and the sediment (about 75 g wet weight) transferred to a pre-extracted Soxhlet thimble where they were mixed with an equal volume of pre-extracted anhydrous sodium sulphate. Soxhlet extraction was carried out overnight using dichloromethane (300 ml) and the solvent extract processed as described previously.
2.3 Gas chromatography (GC) GC was carried out on a 60 m DBl bonded phase fused-silica capillary
column (J & W Scientific) of 0.25 ~ film thickness and 0.25 mm internal diameter installed in a Varian 3300 gas chromatograph. Injection was performed via a Varian on-column injector, the column oven temperature maintained at 30°C for 4 minutes and then linearly increased at 8 °C min-1 to a final temperature of 300°C. Detection was by a flame ionisation detector (FlO) maintained at 330°C.
2.4 Gas chromatography-mass spectrometry (GC-MS) GC-MS was carried out on a Hewlett Packard 5710A
gas chromatograph directly coupled to a VG Analytical 7070E double focussing mass spectrometer. The gas chromatograph was fitted with an OCI-2 on-column injector (SGE) and a fused silica capillary column (DB-1; 60 m x 0.32 mm; 0.25 ~ film thickness; J & W Scientific). Data acquisition and processing were performed by a Super-Incos data system (Finnigan MAT, UK) with mass calibration using perfluorokerosene. All mass spectra were obtained under electron impact (EI) conditions using 70 eV electron energy, 200 ~A trap current and 6kV accelerating voltage. The mass spectrometer was continuously scanned over the mass range 20-700 u at 0.5 sec decade-1 with a dynamic resolution of about 2000 m/lllll.
-156-
2.5 Total organic carbon (TOC) determination Five replicate water samples (50 ml) were collected in glass
stoppered bottles and filtered through ashed (450 °C for 6 hrs) glass fibre filters (Whatman GFF). TOC was determined using an organic carbon analyser (Dohrmann DC80).
2.6 Determination of Organic carbon, hydrogen and nitrogen content of sediment and suspended particles
Suspended particles were filtered off from the effluent, upstream and downstream river water samples using a membrane filter (0.65 micron pore size), removed by scraping and dried (110 oC, overnight). Stored frozen sediment samples were allowed to thaw at room temperature, treated with 12N hydrochloric acid overnight to remove carbonate, centrifuged and the supernatant liquid discarded. The sediment was then washed with distilled water, centrifuged, the supernatant liquid decanted, dried (110 °C, overnight) and milled to a fine powder. The dried particles and sediment were examined using a C, H, N analyser (Carlo-Erba).
2.7 Chloride analysis Triplicate samples were collected in 100 ml plastic bottles. These
samples were then analysed using an ion chromatograph (Dionex Model 12) fitted with a Dionex AS-4A anion separator column and a conductivity detector using a carbonate/bicarbonate eluent.
2.8 Boron analysis Triplicate samples were collected in 100 ml plastic sample bottles.
These samples were then analysed by an air-segmented continuous flow technique on an AutoAnalyser II, using azomethine-H to form a coloured complex which was measured spectrophotometrically at 420 nm.
3. RESULTS AND DISCUSSION
3.1 Selection of study site Details of the seven provisional study sites are given in Table I.
All of the sites represent STWs receiving a mix of domestic and industrial effluent and discharging the treated effluent to lowland rivers. The STWs vary in capacity from 5.5 to 95 x 10 3 m3 day-l average flow and the receiving rivers from 5.3 to 1584 x 10 3 m3 day-l average flow. Dilution of the treated sewage effluent in the receiving river water varies from ~1:1 to ~40:1 for the sites studied. In order to minimise the amount of data processing for this site selection stage of the work only single upstream river and STW effluent filtered composite samples were collected and analysed by solvent extraction and GC-MS.
Table I. Details of the seven provisional study sites
Average daily flow (m 3 )
Site number River Sewage treatment works
1 1584 x 10 3 42 X 10 3
2 50 X 10 3 23 X 10 3
3 5.3 X 10 3 5.5 X 10 3
4 191 X 10 3 6 X 10 3
5 28 X 10 3 37 X 10 3
6 450 X 10 3 95 X 10 3
7 >13 x 10 3 9 X 10 3
-157 -
An ideal site for a detailed study of the transport and fate of organic pollutants has to meet several requirements:-(a) A wide range of compound types present at high and relatively
constant levels in the point source discharge that are absent (or present at very much lower levels) in the upstream river.
(b) Low dilution of the point source discharge by the receiving river. (c) No major inputs of organic pollutants for a distance of several
kilometres downstream from the point source. (d) Access to the river at several points downstream of the point source
discharge that allow collection of both water and sediment samples.
In viev of these requirements the GC-MS data were examined for compounds present in the STY effluent extract at reasonable concentrations that were absent in the upstream river water extract. Two of the sites (2 and 3) showed substantially more compounds than the others and were therefore prime candidates for further study. Unfortunately, most of the compounds could not be identified from the mass spectral data for Site 3. This was not altogether surprising since the STY was known to receive waste water from a specialty chemical manufacturer and thus the compounds may be of obscure structural types. Site 2 showed a greater number of compounds than Site 3 most of which were readily identified and with a wide range of structural types. Since Site 2 also had a low dilution factor of the STY effluent in the receiving river (about 2:1, see Table I) and an approximately 10 km stretch downstream of the STY effluent discharge which did not receive any major input of organic pollutants it was chosen for more detailed study. The time of travel for the 10 km stretch of river varied between 6-18 hours for the usual range of river flows.
3.2 Selection of organic pollutants Site 2 was analysed in more detail to assess its suitability for the
fate and transport study, to select the organic pollutants which would form the basis of the study and to assess how representative the manual composite samples and sediment samples were. Average daily flows over a one year period for the STY effluent and receiving river (see Table I) suggested that the effluent was diluted by a factor of about 2:1 in the downstream river. This approximate figure was checked by examining samples of upstream and downstream river water and STY effluent for 2 determinands which are present at very low concentrations in unpolluted river water and high concentrations in STY effluent, namely chloride and boron (17; see Table II). Mass balance calculations using the measured concentrations of both of these determinands indicated that the STY effluent flow contributed one third of the downstream river water flow, in agreement with the average daily flows reported in Table I. This is an acceptable dilution factor for this study and should enable the selected compounds to be fairly easily traced downstream of the STY effluent discharge. Contribution of pollutants downstream of the STY is from diffuse inputs, ie land runoff and atmospheric deposition, with no major point source input for some 10 km at which point the river flows into a major river.
Assessment of the representativeness of the sampling procedures and selection of the organics for study was done by collecting manual composite samples of the STY effluent, upstream and downstream river water and upstream and downstream sediments. The eff~uent and river water samples were obtained by taking 1 litre grab samples at ten minute intervals and compositing the 1st, 4th, 7th etc samples, the 2nd, 5th, 8th etc samples and the 3rd, 6th, 9th etc samples. This provided interleaved composites and a good check on the representativeness of the
- 158-
Tab
le II
. B
ulk
par
amet
ers
mea
sure
d at
Sit
e 2
Par
amet
er
Tem
p'
pH'
Mea
n2
Mea
n2
Mea
n3
W'e
ight
4 o
f M
ean
2 M
ean
2 M
ean
2
( 0
c)
bo
ron
ch
lori
de
Toe
o
f p
art
icu
late
s o
rgan
ic
carb
on
n
itro
gen
h
yd
rog
en
conc
co
nc
filt
ere
d
per
li
tre
con
ten
t o
f co
nte
nt
of
co
nte
nt
of
(mg
1-1
) (m
g 1
-1)
wat
er
of
wat
er
part
icle
s o
r p
art
icle
s o
r p
art
icle
s o
r (m
g 1
-1)
(mg
1-1
) se
dim
ents
se
dim
ents
se
dim
ents
S
ampl
e (%
) (%
) (%
)
Ups
trea
m
6.9
8
.0
mea
n m
ean
mea
n 8
.6
mea
n m
ean
mea
n ri
ver
<0
.1
42
.9
4.9
1
5.9
6
1.2
1
2.2
2
wat
er
SO
? SO
0
.2
SO
0.4
SO
0
.71
SO
0
.02
SO
0
.06
Oow
nstr
eam
8
.5
7.7
m
ean
mea
n m
ean
19
.9
mea
n m
ean
mea
n ri
ver
0.4
3
55
.1
9.7
3
1.4
6
4.2
8
4.4
3
wat
er
SO
0.0
3
SO
0.3
SO
0
.8
SO
2.1
8
SO
0.4
4
SO
0.3
9
::;;
Sew
age
11
. 3
7.2
m
ean
mea
n m
ean
39
.3
mea
n m
ean
mea
n '"
eff
luen
t 1
.12
8
0.0
1
7.9
4
0.6
7
5.8
0
5.8
6
I
SO
0.0
1
SO
0.2
SO
3
.5
SO
1.4
7
SO
0.3
9
SO
0.3
3
Ups
trea
m
mea
n N
ot
mea
n ri
ver
1.2
8
dete
cte
d
0.2
1
sed
imen
t5
SO
0.1
0
SO
0.0
1
Oow
nstr
eam
m
ean
Not
m
ean
riv
er
0.5
4
dete
cte
d
0.1
6
sed
imen
t 5
SO
0.0
7
SO
0.0
1
Fo
otn
ote
s:
One
m
easu
rem
ent
mad
e w
ith
W'h
atm
an
pH st
ick
on
2.1
2.8
6
Mea
n o
f th
e an
aly
ses
of
thre
e
sam
ple
s M
ean
of
the
an
aly
ses
of
fiv
e
sam
ple
s 4
Tak
en
from
on
e g
rav
imetr
ic d
eter
min
atio
n o
f p
art
icle
s re
tain
ed
by
a
gla
ss fi
bre
fi
lter
(W'h
atm
an
GFF
) 5
Sam
ple
co
llecte
d
from
nea
r to
th
e ba
nk
no
t ap
pli
cab
le
SO
= s
tan
dard
dev
iati
on
sampling procedure. The water samples were filtered to give suspended particulate samples and extracted under both acidic (pH 2) and alkaline (pH 12) conditions. Capillary GC-FID and GC-MS chromatograms obtained for the three composite upstream water samples were quite similar for the identifiable compounds, both qualitatively and quantitatively as also were the chromatograms for the three upstream particulate samples, the three downstream water samples, the three downstream particulate samples, the three STY effluents and the three STY effluent particulate samples. This is illustrated by the GC-MS total ion current chromatograms for the extracts from the three downstream particulate samples (Figure 2). Thus the manual composite sampling procedure used appears to provide representative samples of the STY effluent and river water and the extraction procedure used provides reproducible recoveries of a broad range of organics.
Sediment samples were collected from near to each bank, where the sediments generally consisted of fine particles and from the centre of the stream where coarser-grained sediments were present. The chromatograms obtained were very similar qualitatively but showed lower amounts of all of the organics in the extract of the mid-stream sediment compared to both side-stream sediments. This probably reflects the lower organic carbon content of the coarser grained mid-stream sediments.
Table III shows the major compounds identified by GC-MS in the extracts of the acidified water and suspended particulate samples. Interestingly, the major identified compounds in the basified water extracts, eg caffeine, were also present at similar levels in the acidified water extract. Only low levels of compounds which were not amenable to identification were present solely in the basified extracts. All of the compounds listed in Table III were present at concentrations of about 5-200 ng 1-1 in the water phase. Quantification of the compounds in the particulate phase was more difficult since no internal standards were present in these extracts. On the basis of relative peak heights on GC-MS'compared to standards, some of the compounds were present at concentrations up to a few micrograms per litre in the STY effluent particulates.
Of the major compounds present in the water and particulate samples only £-dichlorobenzene, ~, ~ and E xylene and the fatty acids were present among the major compounds (10 - 500 ~g.kg-l concentration) identified in the sediment extracts. The major compounds in the sediment extracts which were virtually identical for the upstream and downstream sediments were normal, branched and cyclic alkanes and polycyclic aromatic hydrocarbons, which are the expected main components of organic extracts of surficial river sediments (18, 19) and are indicative of anthropogenic inputs.
Selection of compounds for the fate and transport study centred on those that were present only in the STY effluent and downstream river water samples, although some compounds that were present also in the upstream samples were also considered. Table IV shows the compounds selected and their approximate concentrations in the samples examined. The fatty acids, oleic, palmitic and stearic acid were found in all of the water and particulate samples, but were included in the study because they were present at very high concentrations (equivalent to micrograms per litre) in the STY effluent particle extracts. Similarly, the steroidal compounds coprostanol (a widely used marker compound for sewage contamination - see, for example, 20) and cholesterol were present in all of the particle extracts but at much higher concentrations in the STY effluent and downstream samples. Four of the compounds selected, the dichlorobenzene isomers and tributylphosphate, are potential EC list 1 substances and two others, tris-(2-butoxyethyl)-phosphate and
- 160-
4-chloro-3,5-dimethyl phenol, are closely related structurally to potential list 1 substances. Styrene, the alkylbenzenes and 2,6-di-isopropylphenol are also anthropogenic in origin, either being used as intermediates in organic synthesis or occurring as by-products. The fatty acids, squalene and the sterols are all biosynthesised and occur widely, for example in human excreta.
Table III. Major compounds identified in water and suspended particulate samples
Compound Upstream Water Particle
STW effluent Water Particle
Downstream Water
Particle
Alkanes (n-,branched and cyclic)
C3Alkylbenzenes C10 to C13-Alkylbenzenes Benzaldehyde x Benzothiazole x Caffeine x 4-Chloro-3,5-dimethylphenol Cholesterol Coprostanol Cyclohexanone x Dichlorobenzene, 0, m and £ 2,6-Di-isopropylphenol Hexadecanal Hexanal x l-Methyl-4-(1-methylethyl)-
3-cyclohexen-l-ol x Octadecanal Octadecenal Octene Oleic acid x Palmitic acid x Pentachlorophenol x Phenol Squalene Stearic acid x Styrene Tri-n-butylphosphate 2,4,6-Trichlorophenol «,«,4-Trimethyl-3-
cyclohexene-l-methanol x Tris(2-butoxyethyl)phosphate Tris(2-chloroethyl)phosphate x Xylene, £, ~ and £ x
x
x x
x
x x
x x
x
x x x x
x x x
x
x
x x x x
x x x x
x x x x
x
x
x x
x
x x
x x
x x
x
x x x x
x x x
x
x
x x x x x
x x x x
x x x x
Footnote: This list does not include the deuterium labelled standards which were added to the acidified water samples after filtration.
3.3 Analytical requirements
x
x
x x
x
x x
x x
x x
A study of the transport and fate of these compounds requires their accurate quantification in the dissolved, suspended particulate and sediment phases. To this end a range of recovery experiments must be
- 161-
carried out using water, sediment and particulates (or glass fibre filters) fortified with authentic standards of the compounds. Spiking levels can be chosen on the basis of approximate measured concentrations (see Table IV) or where the compound has not been detected in a particular phase (ie water, particulate or sediment), the amount of compound expected in that phase, calculated on the basis of equilibrium partitioning.
Table IV. Compounds selected for fate and transport study
Compound
ClO -C13 Alkyl benzenes
4-Chloro-3,5-dimethyl phenol
Cholesterol Coprostanol Dichlorobenzene,
0, m, and p 2,6-DI-isopropyl
phenol Oleic acid Palmitic acid Squalene Stearic acid Styrene Tri-n-butyl
phosphate Tris(2-butoxy ethyl) phosphate
- = not detected
Upstream \later Particle
100-1000 100-1000
5-50 <100 10-100 100-1000
100-1000 10-100 100-1000
Concentrations (ng 1-1) ST\l effluent Downstream
\later Particle \later Particle
20-200 10-100
10-100 10-100
20-30xl03 1O-15xl03 35-45xl0 3 l7-22xl0 3
10-100 10-100 10-100 10-100
10-100 2-10x103 10-100 1-5xl03 10-100 10-20x103 100-200 5-10x103
2-l0xl0 3 1-5x103 10-100 10-20xl03 10-100 5-lOxl0 3 ca 100 10-100 ca 100 10-100
100-200 10-100
The sediment (or particulate) water partition coefficient, Kp, is defined as the ratio of the concentration of pollutant in the sediment (or particulate) phase (Cp) to the concentration of pollutant in the water (Cw) ie Kp = Cp/Cw and hence log Cw = log Cp - log Kp and log Cp = log Cw + log Kp.
Partitioning to sediment (or particulate) is related to the fractional organic carbon content of the sediment (or particulate) and it is, therefore, useful to define a fractional organic carbon partition coefficient, Koc, which is the ratio of the concentration of pollutant in the sediment (or particulate) in ~g.kg-l organic carbon to the concentration of pollutant in the water in ~g.1-1. Koc is related to Kp by the simple expression, Koc = foe x Kp, where foe is the fractional organic carbon content of the sediment (or particulate). It has been found experimentally for a range of pollutants that Koc is related to the octanol/water partition coefficient of a pollutant by the equation log Koc = 0.989 log Kow - 0.346 (21). Thus Koc can be calculated from the Kow for the compound of interest. Kow values were obtained from the
-162 -
scientific literature or calculated using standard methods (22). Substituting into the above equations the measured values of water or suspended particulate concentration and the measured fractional organic carbon content of the downstream particles (Table II: foc = 0.3, log foc = - 0.52) or sediment (Table II: foc = 0.0054, log foc = - 2.27) allowed calculation of the theoretical concentrations of the compounds in the phases in which they were not measured.
Table V shows the concentrations of the selected compounds in the downstream water (a), particulate (b) and sediment (c) phases calculated from approximate measured concentrations and octanol/water partition coefficients. These calculations indicate that it will be necessary to measure 4-phenyl dodecane and squalene at the low pg 1-1 level in water in order to determine whether equilibrium partitioning is actually taking place and this represents a very difficult analytical problem which may be insoluble even by GC-MS with mUltiple ion detection (MID). However, these are the two worst cases and it should be a more straightforward task to determine the concentrations of the other compounds in water. Calculated particulate concentrations range from 1.1 pg mg- 1 for styrene to 720 pg mg-1 for tris-(2-butoxyethyl)phosphate calculated from approximate water concentrations and the lower levels will require the use of GC-MS in HID mode to achieve sufficient sensitivity of detection. Sediment concentrations predicted from the calculations vary over several orders of magnitude from 20 pg g-1 for styrene to 20 ~g g-1 for coprostanol and will again require GC-MID to provide the required sensitivity of detection at the lower concentrations. It is likely that derivatisation of the more polar compounds will be necessary to achieve sufficiently low detection limits using GC-MS in MID mode. Derivatisation will both improve the chromatography of the fatty acids, sterols and phenols and result in a greater proportion of the total ion current being borne by high mass structure-specific fragments.
In view of the high calculated sediment concentrations (several ~g g-1) for the fatty acids (palmitic, oleic and stearic), squalene and the sterols (cholesterol and coprostanol) it is surprising that only the fatty acids (10-500 ng g-1; Table IV) were detected in the sediment extract. Two possible explanations for this discrepancy are (i) the downstream surficial sediments are highly scoured by the river water and hence are "diluted" by deposited upstream sediment, and (ii) these compounds are undergoing a non-equilibrium partitioning, a phenomenon which has been previously reported for PAH compounds in sediments (23).
Table V. Calculated concentrations of the selected compounds in water, particulate and sediment phases.
(a) Vater
Compound Approximate log Kow log Koc log Kp Calculated conc in water conc
particles (ng 1-1) (mg kg-1)
oleic acid 50-250 5.36 4.96 4.44 1800-9100 palmi tic acid 250-500 6.2 5.79 5.27 1300-2700 stearic acid 250-500 8.23 7.79 7.27 13-27 4-phenyldodecane 0.5-5 8.76 8.32 7.80 0.0079-0.079 cholesterol 500-750 9.5 9.05 8.53 1.5-2.2 squalene 50-250 9.86 9.41 8.89 0.064-0.32 coprostanol 875-1125 10.71 10.25 9.73 0.16-0.21
-163 -
(b) Particulate
Compound Approximate log Kow log Koc log Kp Calculated conc in particle water conc
(ng 1-1 ) (ng. mg-1 )
styrene 10-100 2.95 2.57 2.05 0.0011-0.011 £,-dichlorobenzene 5-50 3.37 2.99 2.47 0.0015-0.015 o,-dichlorobenzene 5-50 3.38 3.00 2.48 0.0015-0.015 iii, -d i chlor,obenzene 5-50 3.48 3.10 2.58 0.0019-0.019 4-chloro-3,5-dimethyl
phenol 10-100 3.50 3.12 2.60 0.004-0.04 tri-n-butylphosphate 10-100 4.00 3.61 3.09 0.012-0.12 2,6-di-isopropylphenol 10-100 4.72 4.32 3.80 0.06-0.6 tris(2-butoxyethyl)
phosphate 10-100 4.78 4.38 3.86 0.072-0.72
(c) Sediment
Compound Approximate log Koc log Kp Calculated conc in sediment conc water (llg. g -1)
(ng. 1-1 )
styrene 10-100 2.57 0.30 2.0x10-s-2.0x10-4 £,-dichlorobenzene 5-50 2.99 0.72 2.6x10-s-2.6x10-4 o,-dichlorobenzene 5-50 3.00 0.73 2.7x10-s-2.7x10-4 iii,-dichlorobenzene 5-50 3.10 0.83 3.4x10-s-3.4x10-4 4 chloro-3,5-dimethyl-
phenol 10-100 3.12 0.85 7.1x10-s-7.1x10- 4
tri-n-butyl phosphate 10-100 3.61 1.34 2.2x10- 4-2.2x10- 3
2,6-di-isopropylphenol 10-100 4.32 2.05 0.0011 - 0.011 tris(2-butoxyethyl)
phosphate 10-100 4.38 2.11 0.0013 - 0.013 oleic acid 1800-9100* 4.96 2.69 0.9-4.5 palmitic acid 1300-2700* 5.79 3.52 4.5-9 stearic acid 13-27 7.79 5.52 4.5-9 4-phenyl dodecane 0.0079-0.079* 8.32 6.05 0.009-0.09 cholesterol 1.5-2.2* 9.05 6.78 9-14 squalene 0.064-0.32* 9.41 7.14 0.9-4.5 coprostanol 0.16-0.21* 10.25 7.98 16-20
* Calculated from particulate concentration, see (a)
Similarly, the detection of E-dichlorobenzene at a level of 10-500 ng g-l in the sediment compared to a predicted concentration of 0.03-0.3 ng g-l indicates either that non-equilibrium partitioning is occurring or there is another source of £-dichlorobenzene contributing to the sediment. In fact the latter is most likely since the upstream sediment extract contains similar amounts of p-dichlorobenzene (10-500 ng g-l) to the downstream sediment. -
4. CONCLUSIONS
1. The site selected for detailed study is very nearly ideal, with a major point source (STY effluent) contributing one third of the
- 164-
downstream flow, good access for sampling and no major downstream organic inputs for a distance of nearly 10 km.
2. The manual composite sampling method used provides representative samples of the effluent and river water.
3. A wide range of pollutants are present in the ST~ effluent that are absent or present at much lower concentration in the river upstream of the discharge.
4. The physicochemical properties of the selected pollutants cover a broad range with log Kow values from about 3 to 11.
5. Measured and calculated concentrations of the compounds in the water, particulate and sediment phases are sufficiently high in most cases to be able to measure their concentrations in real samples accurately and determine partition coefficients from these field measurements.
5. FUTURE ~ORK Recovery studies will be carried out for the selected compounds by
spiking authentic standards into water, sediment and particulate samples at realistic concentrations measured or calculated previously. Once the analytical methods have been validated, composite samples of effluent, upstream and downstream river water and sediment at several points will be collected from the site and analysed for the selected compounds. The river flow will be measured at the time of sampling and the collection of downstream water samples staggered to allow sampling of the same plug of water as it moves down the river. The results obtained will be used to determine sediment/water and suspended particulate/water partition coefficients which will be compared with partition coefficients calculated previously from log Kow values to ascertain if the results are compatible with current sorption modelling techniques. Measured field concentrations will also be compared with concentrations predicted by aquatic environmental modelling techniques, eg EXAMS and this will enable an assessment to be made of the ability of models to predict "real" environmental concentrations of the selected organic pollutants.
6. ACKNO~LEDGEMENTS
The authors thank Mr M Righton for assistance with the early part of the work, Mr M Fielding and Dr A J Dobbs for helpful and stimulating discussions, Mr T Gibson for obtaining the GC-MS data and the director of ~ater Research Centre, Environment laboratory for permission to publish.
REFERENCES
(1)
(2)
(3)
(4)
(5)
DI TaRO D M (1985) A particle interaction model of reversible organic chemical sorption. Chemosphere 14, 1503-1538. KARICKHOFF S ~, BRO~ D S and SCOTT T A (1979) Sorption of hydrophobic pollutants on natural sediments. ~ater Res 13, 241-248. SCH~ARZENBACH R (1985) Sorption behaviour of neutral and ionizable hydrophobic organic compounds. In Proc 4th European Symp on Org Micropollut Aquat Environ, 168-177. CARTER C ~ and SUFFETT I H (1983) Interactions between dissolved humic and fulvic acids and pollutants in aquatic environments. In Fate of chemicals in the environment edited by R L Swann and A Eschenroeder. ACS Symp Ser 225, 11, 215-229. MACKAY D L, PATERSON S and SCHROEDER ~ H (1986) Model describing the rates of transfer processes of organic chemicals between atmosphere and water. Environ Sci Technol 20, 8, 810-816.
- 165-
(6) JENSEN-KORTE U, ANDERSON C and SPITELLER, M (1987) Photodegradation of pesticides in the presence of humic substances. Sci Tot Environ 62,,335-340.
(7) MANSOUR M and KORTE F (1986) Abiotic degradation pathways of selected xenobiotic compounds in the environment. Stud Environ Sci 29, 257-269.
(8) FAUST S D (1977) Chemical mechanisms affecting the fate of organic pollutants in natural aquatic environments. Adv Environ Sci Technol 8, 2, 317-365.
(9) RICHARDS D J and SHIEH W K (1986) Biological fate of organic priority pollutants in the aquatic environment. Water Res 20, 9, 1077-1091.
(10) HOLYSH M, PATERSON S, MACKAY D, BANDURRAGA M M (1986) Assessment of the environmental fate of linear alkylbenzene sulphonates. Chemosphere 15, 1, 3-20.
(11) HALFON E (1986) Modelling the fate of Mirex and Lindane in Lake Ontario, off the Niagara river mouth. Ecol Modell 33, 1, 13-33.
(12) BURNS L A and CLINE D M (1985) Exposure analysis modelling system. Reference manual for EXAMS 2. NTIS PB Report, 85-228138.
(13) BURNS L A (1983) Fate of chemicals in aquatic systems. Process models and computer codes. In Fate of chemicals in the environment, edited by R L Swann and R A Eschenroeder. ACS Symp Ser 225, 25-40.
(14) IMBODEN D M (1986) Mathematical modelling of the behaviour of organic micropollutants in the aquatic environment. Comm Eur Communities (Rep) EUR Org Micropollut Aquat Environ, EURI0388, 460-464.
(15) LOPEZ-AVILA V and HITES R A (1980) Organic compounds in an industrial wastewater. Their transport into sediments. Environ Sci Technol 14, 11, 1382-1390.
(16) JUNGCLAUS G A, LOPEZ-AVILA V and HITES R A (1978) Organic compounds in an industrial wastewater. A case study of their environmental impactL Environ Sci Technol 12, 1, 88-96.
(17) WAGGOTT A (1969) An investigation of the potential problem of increasing boron concentrations in rivers and water courses. Water Res 3, 749-765.
(18) READMAN J W, MANTOURA R F C, LLEWELLYN C A, PRESTON M R and REEVES A D (1986a) The use of pollutant and biogenic markers as source discriminants of organic inputs to estuarine sediments. Intern J Environ Anal Chern 27, 29-54.
(19) GIGER Wand SCHAFFNER C (1978) Determination of polycyclic aromatic hydrocarbons in the environment by glass capillary gas chromatography. Anal Chern 50, 2, 243-9.
(20) READMAN J W, PRESTON M Rand MANTOURA R F C (1986b) An integrated technique to quantify sewage, oil and PAH pollution in estuarine and coastal environments. Mar Poll Bull 17, 7, 298-308.
(21) KARICKHOFF S W (1981) Semi-empirical estimation of sorption of hydrophobic pollutants on natural sediments and soils. Chemosphere 10, 8, 833-846.
(22) HANSCH C and LEO A J (1979) Substituent constants for correlation analysis in chemistry and biology. John Wiley, New York.
(23) READMAN J W, MANTOURA R F C and RHEAD M M (1984) The physico-chemical speciation of polycyclic aromatic hydrocarbons (PAH) in aquatic systems. Fresen Z Anal Chern 319, 126-131.
- 166-
- ~ I
AIR
WA
TE
R
air
par
ticl
e •
... ,
de
posi
tion
"."
:::.
""
(wet
and
dry
l •.
.• .'
.
Wet
I)
() de
posi
tion
II
b 6
b 6
Wat
er
Fig
ure
1.
Pro
cess
es
aff
ecti
ng
ch
em
ica
ls i
n t
he
riv
er
en
vir
on
me
nt.
chem
ical
de
grad
atio
n (h
ydro
lysi
s &
ox
id a
tion
I
Figure 2. Total ion current chromatograms for the downstream particulate samples
100.0
RIC
100.0
RIC
100.0
RIC
500 1000 1500 2000 2500 3000 3500 4000 SCAN 8 :06 16:12 24:18 32:24 40:30 48:36 56:42 64:48 TIME
- 168-
POSTER SESSION II TRANSPORT
Sediment-water partition coefficients of hydrophobic chemicals in the presence of third phase material
Environmental fate of organosilicon chemicals
Pollution of Saronicos Gulf (Athens, Greece) by fossil fuel hydrocarbons
The Sandoz/Rhine accident The environmental fate and transport of twenty-one pesticides introduced to the Rhine River
Occurrence and leaching of pesticides in waters draining from agricultural land
Polychlorinated biphenyls in the Kupa river, Croatia, Yugoslavia
SEDIMENT-WATER PARTITION COEFFICIENTS OF HYDROPHOBIC CHEMICALS IN THE PRESENCE OF THIRD PHASE MATERIAL.
S. MARCA SCHRAP and ANTOON OPPERHUIZEN
Laboratory of Environmental and Toxicological Chemistry, University of Amsterdam, Nieuwe Achtergracht 166, 1018 WV Ams'terdam, The Nether lands.
Summary
Aqueous solubilities of three chlorobenzenes in sedimentwater systems are significantly enhanced by the presence of third phase material originating from the sediment. Because the aqueous concentrations of the chlorobenzenes are influenced by this third phase material, their sorption coefficients will also be influenced. Besides the experimenta 1 cond i t ions, the characteri st i cs of the sediment determine the third phase concentration. Pretreatment of the sediment, such as drying in an oven can change these characteristics significantly.
1. I NTRODUCTI ON
It is often assumed that sorption of hydrophobic chemicals can be seen as a partitioning of the chemical between the sediment and water. The sorption coefficient is defined by:
(1)
where Kp denotes the sorption coefficient (l/kg) and Cs and C the concentrations of the test compound on the sediment (g/kg1 and in the water (g/1), respectively. The sorption coefficients are assumed to be independent of the concentrations of the compound in the water and on the sediment and of the concentration of the sediment in the water.
Although support for this model has been provided (1,2), the decrease of sorption coefficients with increasing particle concentration (3), and an increase of sorption coefficients in consecutive desorption experiments (4) seem to contradict the partition model of sorption on sediments. In a recent study however (5) we have shown that these observations are a manifestation of an incomplete separation of the sediment and water in sorption experiments rather than a deviation from the sorpt ion mode 1. The not-separated part of the sed iment (the third phase) increases the concentration of the test compound in the aqueous phase, and an underestimation of the sorption coefficient will thus follow.
- 170-
2. AQUEOUS SOLUBILITY MEASUREMENTS
For three chlorobenzenes (1,2,3 tri-, penta- and hexach lorobenzene) the aqueous so 1 ubi 1 it i es were measured in distilled water and in overlying water from a sediment-water suspension. Suspensions (3 gram sediment, dry weight, in 200 ml water) were shaken for 48 hours at 21 0C and then centrifuged with 2000 rpm for 20 min. From the overJying water 150 ml was used for the solubil ity measurements. Sediments were taken from a fresh water lake (Oostvaardersplassen, The Netherlands) and contain 2.9% organic carbon. Both oven dried (2000C for 2 hours) and wet sediments were used in the experiments.
The aqueous solubilities were determined with a generator column as was previously described (6). The column contained Chromosorb (100-120 mesh) which was impregnated with 3% of a test compound. Each test compound was studied individually. With one column first the aqueous solubility in distilled water was determined, then the solubility in overlying water from wet sediment and finally the oven-dried sediment. After the experiments with overlying water the solubility in distilled water was determined again as a control of the system.
In each experiment at least 6 samples of 20 ml were taken, which were extracted immediately with 2 ml toluene. Th i s was injected ina gas chromatograph wh i ch was equ i ped with an electron capture detector.
Table 1. Aqueous solubilities of 1,2,3 tri-, penta and hexach lorobenzene in d i st i lled water, and over lyi ng water of wet and oven-dried sediment.
Compound distilled overlying overlying chloro- water water wet water oven-benzene sediment* dried sed.*
( g/1) ( g/1) ( g/1)
1,2,3 tri 9.8.103± 2.4.103 9.9.103± 2.4.103 13 .8.103±2. 7 .103
penta 2.2.102± 0.2.102 2.8.102± 0.3.102 4.1.102± 0.2.102
hexa 2.8 ± 0.7 5.9 ± 1.4 10.4 ± 2.1
* obtained after centrifuging 15 gil sediment/water suspension (dry weight)
In table 1 the aqueous solubilities in water and in overlying water are given for the three chlorobenzenes. As can be seen a significant increase of the concentrations in the aqueous phase was found when an amount th i rd phase, originating from the sediments, is present in the water. It
- 171-
can a 1 so been seen that over lyi ng water obta i ned from wet sediment and from oven-dried sediment does not enhance the aqueous solubilities of the polychlorinated benzenes to the same extent. Since the suspensions were shaken with the same sed iment concentrat i on (based on dry wei ght) and were separated with the same technique, it is clear that the thirdphases of oven-dried and wet sediments are not comparable. Whether the amount or the properties of the third phase material are altered is not clear. Since this third phase containing overlying water is obtained under the same experimental conditions as those in which sediment/water sorption coefficients are usually determined, the presence of a third phase should be considered when sorption coefficients are determined using measured concentrations of the test compounds in the aqueous phase.
3. ~ORPTION MODEL
Assuming that the sorption process can be seen as a partition process and assuming that the presence of third phase materi ali nf 1 uences the concentrat ion in the aqueous phase~ the experimentally determined sorption coefficient (KpeXp ) can be described by (7):
Kpexp = Cw + Ctp
(2)
where Ctp denotes the concentration of the compound associated with the third phase {gil). So Ctp is the concentration of the compound on the third phase (Cm in g/kg) multiplied by the third phase concentration (s in kg/l) :
Ctp = Cm . s (3)
The component's sorption coefficient between water and the third phase (Kpm) is then given by:
Kpm = Cm / Cw
Combining equations 1 to 4 gives:
Kpexp = 1 + Kpm' s
If it is assumed that Kp equals Kpm then written as:
1 Kpexp =
l/Kp + s
-172 -
(4)
(5)
equation 5 can be
(6)
From this equation it can be seen that the experimentally determi ned sorpt ion coeff i c ient decreases with an i ncreas i ng third phase concentration. This is illustrated in figure 1 where for three hypothetical compounds the experimental sorption coefficients are plotted against the third phase concentration. From this figure it is clear that experimentally determined sorption coefficients are more influenced by third phase material if the sorption coefficients are higher. Hence I even very low third phase concentrations can significantly influence the experimentally determined values of compounds with very high sorption coefficients.
exp log Kp
4+----~
3 ---- - -- ------------- __ ' •
2
.. ...............
"-
1 ...... .... _. __ .... - .... -._-.. _ ..... --... -.••....... .....• ........... _._ •
o
-1
-2
-3
.....
-6 -5 -4 -3 - 2 -1
- Kp= Kpm =10 ___ Kp" Kpmlo103
_ Kp = Kg",z104
o logS
Fig 1. Influence of the third phase concentration (kg/l) on the experimenta 1 sorpt ion coeffi c i ent for three hypothet i ca 1 compounds.
- 173-
In order to determine the sorption coefficient from equation 6 the third phase concentration must be known. This can not be measured directly, because of the unknown composition of the third phase. The amount of third phase (m),
m = s. V (7)
with V the volume of water (1), can be expressed in terms of the sediment's characteristics. Hereby it may be assumed that a fixed fraction of the sediment can act as third phase material (0). This fraction is dependent on the sediment used.
0= mtot/M (8)
Here M denotes the weight of the sediment and mtot the weight of the part of the sed iment wh i ch can act as th i rd phase material. However, during suspension of sediment in water only a fraction of the third phase material (9) will be in the water phase, the rest (1-9) remains in the sediment. After separation of the sediment and the water containing third phase, the amount of th i rd phase in the water phase (m) will be:
m = 9 mtot (9)
Combining the equations 6 to 9 shows that in experiments with compounds with sufficiently high Kp the measured sorption coefficients are dependent on M,O ahd 9.
4. CONCLUSIONS
The presence of a th i rd phase in the aqueous phase in sediment-water systems enhances the aqueous solubilities of test compounds. This effect depends on the third phase concentration and the affinity of the compound with the third phase material. The third phase concentration is on the one hand dependent on the experimenta 1 cond it ions (concentrat ion of the sediments and the separation techniques used), and on the other hand on the properties of the sediment. The fraction of the sediment that can act as third phase material and the fraction of that third phase that will be in the water after every suspension are characteristics of the sediment. Pretreatment of the sediment, such as drying in an oven, causes a Significant change in the properties of the aqueous third-phase and/or the sediment. Hence, sediment-water partition coefficients determined with differently treated sediments may not be comparable, even when the same sediment and the same separation techniques are used.
Acknowledgement This work was supported by the Institute for Inland Water
Management and Waste Water Treatment, Ministry of Transport and Public Works, The Netherlands.
-174 -
REFERENCES
(1) CHIOU C.T., PETERS L.J., FREED V.H. Science 1979,206, 831- 832.
(2) CHIOU c.T., PORTER P.E., SCHMEDDING D.W., Environ.Sci.Technol. 1983,lZ,227-231
(3) O'CONNOR D.J., CONNOLLY J.P, Water res., 1980,li, 1517-1523
(4) HORZEMPA L.M., DITORO D.M., Water res.,1983, 851-859. (5) SCHRAP and OPPERHUIZEN, submitted for publication. (6) OPPERHUIZEN A., GOBAS F.A.P.C., VAN DER STEEN J.M.D.,
Environ.Sci.Technol., in press. (7) GSCHWEND P.M., WU S.C. ,Environ.Sci. Technol., 1985, l2.....
90-96.
-175 -
ENVIRONMENTAL FATE OF ORGANOSILICON CHEMICALS
Antoon Opperhuizen, Guus M.Asyee and John R.Parsons Laboratory of Environmental and Toxicological Chemistry
University of Amsterdam Nieuwe Achtergracht 166, 1018 WV Amsterdam, The Netherlands
Summary In a-biotic samples obtained from the natural aquatic environment evidence has been reported for the presence of ,silicones from antropogenic sources. It has also been reported however that a-biotic transformation may play an important role in the environmental fate of these chemicals. In particular clay catalyzed hydrolysis may be important. Hitherto little is known about the ability of (micro)organisms to transform silicon compounds. Since most organosilicon compounds are che micals with very low aqueous solubilities and high solubilities in organic solvents it was expected that these chemicals would accumultae to high concentrations in organisms. This however, is not what has been found in laboratory tests with fish or mammals, or in biota sampled in the natural environment. The low bioaccumulation factors may be explained by high rates of elimination of the short chain silicones and by a lack of uptake for the long chain chemicals.
INTRODUCTION During the last four decades the production and use of
organosilicon compounds has grown very fast. In 1974 the annual production was estimated being 150 metric tons (1) , while in 1984 this was estimated to be 1500 metric tons (2). The total production is composed of silanes, silicon fluids, resins and elastomers. By varying the structure of the organosilicon compounds an enormous series of substances with different physico chemical properties can be made. This, in combination with the low chemical reactivity make that organosilicon compounds are used in an increasing number of applications.
The chemical stability in combination with the low aqueous and high organic solvent solubility of most of the organosilicon compounds and the high production volume means that these chemicals should be of concern to environmental chemist, toxicologist and biologist. In the present paper the literature on the environmental fate of organosilicon compounds will be reviewed.
RESIDUES IN THE ENVIRONMENT Despite the large annual production volume, analyses of
organosilicon compounds are usually not incorporated in the present chemical monitoring programmes. In various studies the presence of these compounds in the naturql environment has been demonstrated.
In atmospheric samples organosilicon compounds are only rarely identified. In a few cases concentrations from local sources have been reported. With respect to non-point sources
- 176-
polydimethylsiloxanes were found in Arctic aerosols in concentrations ranging from 0.1 to 20 ng/m3 (3).
Residues in the aquatic environment have been found in a number of studies. During a 6 month period in 1978, octamethylcyclotetrasiloxane (04) has been found in the lower region of the Rhine River at level between 100 and 1000 ng/l (4). Concentrations between 1 and 100 ng/l were measured regularly in the River Lee near London (5). In French fresh water samples polydimethylsiloxanes were found together with many other micropollutants in concentrations between 0.01 and 100 ng/l. In nearly all samples organosilicon compounds were found, and in several cases these compounds accounted for 30 % of the total amount of the micropollutants (6). Local pollution of the aquatic environment with methylsilicon compounds had been reported for a matura.tion pond in South Africa in which a sudden increase of algae sedimentation was found. Estimated maximum concentrations were as high as 0.1 mg/l (7). Furthermore in waste water from a petrochemical industry a silicone surfactant was found. Although actual concentration were not reported, they were estimated as being in the mg/l range (8).
In water sampled in rivers from the Nagara River watershed (Japan) concentrations of organosilicon compounds between 2.0 and 54.2 ~g/l were measured (9). In sediments from these rivers concentrations between 0.3 and 5.8 mg/kg were found (9). In aquatic sediments of Pontomac River (Washington,D.C.) silicone concentrations between 0.46 and 3.07 ( mean 1. 38) mg/kg have been found (10), and the sediments in New York Bight, contained concentrations ranging from below the detection limit to 50 mg/kg (11). Aqueous surface micro-layers collected in Delaware Bay and in Chesapeake Bay, contained significantly lower concentrations, Le. 34.1 and 30 gil. In the Chesapeake Bay sediments the concentrations of the organosilicon compounds ranged from 0.3 mg/kg in the upper region of the bay to 36.1 mg/kg in the mouth of the bay (10). In a recent study with Puget Sound sediments it has been demonstrated that silicones found in the superficial sediments (sediment age 15 years ) are not present in older sediments ( sediment age 60 years )(2). Since organosilicon production started in 1943, these results are in agreement with the assumption that all organosilicon compounds found in the environment originate from antropogenic sources.
In biotic samples organosilicon pollution has been reported in at least two cases. In the gills of Gammurus duebeni a variety of micropollutants were found. Amongst these silicones were identified (12). In addition, in expired air from a heterogeneous non-smoking human population hexamethylcyclotrisiloxane and other silanated compounds have been found among some 350 other compounds (13). Since these silicones did not originate from metabolic transformations, nor from natural sources, they were considered to be environmental pollutants.
A-BIOTIC FATE OF SILICONES Distribution processes
Little is known about the environmental transport organosilicon compounds. This is both due to a lack
-177 -
of of
monitoring data and to uncertainties and contradictions in the available data sets.
Most silicones have relative high molecular weights and low vapour pressures, so it is often assumed that transfer of these chemicals to the atmosphere will be limited and sorption onto soils or sediments is considered to be more important. However, the vapour pressure of many silicone fluids is in the range of 10-5 to 10-0 rom Hg, these magnitudes being comparable to those of polychlorinated biphenyls and DDT for example. For these latter compounds atmospheric transport contributes significantly to their global fate. The detection of silicones in the Arctic air samples supports the hypothesis that atmospheric vapour and aerosol transport may be important for silicones (3). Since it has been suggested that compounds with vapour pressures higher than 10-6 rom Hg are predominantly present in the gas phase in remote regions, it is likely that vapour phase transport is important for small chain silicone oligomers. For compounds with saturation vapour pressures lower than 10-8 rom Hg, such as polydimethylsiloxanes with 18 or more siloxane units, aerosol transport is probably more important.
Most silicones have low aqueous solubilities and relatively high solubilities in organic phases. Also their octan-1-01/water partition coefficients of most of these chemicals are high (14). These hydrophobic/lipophilic properties resemble those of many hydrocarbon micropollutants. These similarities attest to the hypothesis that silicones tend to concentrate in atmospherehydrosphere boundaries and in hydrosphere-sediment boudaries ( 3,7,10). Whereas the transfer from the hydrosphere to the atmosphere may be fast, the transfer from sediments to hydrosphere seems to be very slow (15 ). In addition, since transport of organosilicon compounds in wet soil is very slow, sediment"s can act as a sink of these micropollutants.
Little is known about the role land soils play in the fate of silicones. In a few studies it has been shown that silicones are fairly mobile in dry or damp soils (16). No information is available however on the rates of exchange of organosilicon compounds between lithosphere and atmosphere. Transformation processes
With respect to photochemical degradation, usually a distinction is made between liquid and volatile organosilicones, since the physico-chemical state of the latter chemicals make fast complete degradation unlikely. After irradiation of methylsiloxane resins with wavelengths longer than 281 nm for more than two months, cross-linking of the polymers and the formation of Si-CH2-Si linkages have been found (17). Irradiation for two weeks with wavelengths shorter than 281 nm caused the formation of Si-CH2-CH2-Si and Si-OH linkages, but the environmental significance of the latter products is unknown. After irradiation of polydimethylsiloxanes and polyphenylmethylsiloxanes to wavelengths between 340 and 360 nm, CH3 and R-CH2 radicals were formed as the results of two quantum reactions (18). After investigation of the mechanisms of photochemical transformation of several volatile methylsilicon species is was concluded that pure air and sunlight are not
- 178-
sufficient to cause significant degradation. Small additions of 03 or H20 however caused significant degradation, probably due to the formation of singlet oxygen .from excited ozone (19,20). Irradiation of octamethylcyclotetrasiloxane showed the formation of more than 30 products, among which various condensation products. The persistence of volatile organosilicon compounds is comparable to those of aliphatic hydrocarbons (19,20). The atmospheric persistence of the non-volatile compounds however is assumed to be much higher.
It has been suggested that silicone oils in natural waters hydrolyze to form silanols (21,22), but evidence for this transformation reaction is lacking. It was shown however that after irradiation of a diluted (CH3)2Si(OH)2 solution in a natural water demethylation occurred, while irradiation in pure water did not show any reaction (23). Photochemical degradation of other organosilicones in aqueous environments has also been found (24,25). It was suggested that the presence of nitrate or nitrite as sources of hydroxyl radical via singlet oxygen, is required to allow the organosilicon chemicals to react after irradiation. The hydroxyl radicals are highly reactive species responsible for attacking the Si-CH3 bonds (23).
Degradation studies of organosilicones in soils are mainly focused on the liquid polydimethylsiloxanes. In a study with PDMS (100 centistokes) coated on a neutral soil, a half-life of 10 days has been measured (1). Further investigation of a variety of silicone fluids revealed that the decay of these chemicals on soils follows first order kinetics with half lives between a few minutes and several weeks (26). Mechanistic investigations showed that rearrangement reactions and hydrolysis yielded low molecular weight volatile cyclic and linear oligomers, as well as hydroxyl-products which are probably more soluble in water. Hence, it was concluded that ,these'siloxanes are removed from the soil after a catalytized transformation. It was shown that the components responsible for the catalytic function of soils are clays. It has been known for a long time that acids and bases cleave Si-O-Si linkages. Active clays are also able to cleave these bonds (27), in contrast to sands and humus, and thus the amount of clay of the soil is important for the rate of degradation of organosilicones in soils. In addition, although hydrolysis requires the presence of some water, high levels of hydration of the soil reduce the soil's catalytic activity (23). Furthermore it has been shown, that both the type of the clay material and the temperature of the soil are important to both the mechanism and the kinetics of the degradation (27).
BIOTIC FATE PROCESSES. Distribution-processes
Although few studies have been carried out to investigate the bioaccumulation capacity of organosilicon compounds in the lower trophic levels of the food chain (28), it is clear that long chain polymeric silicones are not taken up by organisms. However, most commercial formulations contain a small but significant amount of small chain oligomers, the
-179 -
polydimethylsiloxanes being the most important class of these oligomers. Hitherto only the bioaccumulation of polydimethylsiloxane oligomers has been investigated. In preliminary bioaccumulation studies with cyclic and linear polydimethylsiloxanes it has been shown that although some uptake is found, these chemicals do not significantly accumulate in fish, neither after dietary nor after aqueous exposure (14,29,30). This is despite their hydrophobic properties, which are comparable to those of polychlorinated biphenyls. Recently it has been shown that the lack of accumulation of the long chain oligomers, twelve or more silicon atoms, may be due to a lack of membrane permeation (30). For the short chain oligomers however, significant uptake has been found, but high rates of elimination prevented accumulation. The elimination half lives for all oligomers were less than 6 days (30). Although it was previously suggested that uptake may be limited to the gills or to the exterior of the fish (29), recently residues in the liver were also found (30).
Laboratory studies with mammalian species have not indicated that polydimethylsiloxane fluids bioaccumulate (21), although experiments with rats revealed that some uptake occurs. After exposure of rats to s single oral dose of 03' 04 and 05 , significant levels of 04 and 05 were found in the liver and some trace level in fat tissues (21). The 05 oligomer had a significant higher accumulation potential than the 04' Exposure of the rats to a mixture of linear oligomers showed that maximum accumulation in fat occured for the M04M oligomer, but in the liver for the M06M oligomer.
In a series of experiments with Antifoam A and M ( both are polybimethylsiloxanes polymers containing respectively 10 % and 0.02 % 03 to 06 cyclic oligomers) with mice, rats, rabbits, dogs, monkeys and humans, it was found that the higher molecular weight siloxanes were not appreciablely retained in the organs and primarily excreted via the feces (1). The smaller molecular weights siloxanes ( six siloxane units or less ) however, were found to be absorbed by the organisms in small amounts. Biotransformation processes.
Little has been reported on the biotransformation of organosilicon compounds, in particular for the polydimethylsiloxanes. From studies on the biotic fate of labelled phenyl-trimethyl silane and phenyldimethylsilane in rats some major metabolic routes for the siloxanes may be deduced (21,31,32). The urinary metabolites of the phenyltrimethylsilane after single oral administration were hydroxylated-products of both the alkyl and aryl moieties (31,32). The phenyldimethylsilane formed phenyldimethylsilanol, which could undergo condensation to form a disiloxane during the process of purification. Oealkylation as well as oxidation have been suggested to be the main metabolic processes of the organosilicon compounds, although clear evidence is still lacking (21,31,32). From the studies with the phenyl-methyl silanes it can be proposed that cleavage of the Si-O linkage or oxidation of the methyl groups may be the principal steps in the biotransformation of the chemicals. Evidence for the cleavage of
-180-
the Si-O linkages has been provided by the identification of mono- and dihydroxy siloxanes after administration of the polydimethylsiloxane D4 to rats (21). In experiments with Antifoam A in humans, organosilicon compounds more polar than the cyclic oligomers were found (15). The chemical structures of these compounds was not elucidated.
The detection of hydroxy derivatives of organosilicon compounds after oral administration of organosilicones, may suggest that these compounds are formed by non-enzymatic acidic hydrolysis in the stomach. In recent studies with fish however, it has been shown that after both dietary and aqueous exposure of series of cyclic and/or linear polydimethylsiloxanes by gas chromatography a homologous series of compounds can be found, of which the structures are unknown, together with various rearrangement products (30). Hitherto however no definite evidence is found whether or not these compounds are products of biotransformation reaction inside the fish.
A number of studies have shown that microorganisms may grow on the surface on silicone rubbers and resins, but are unable to degrade them (1). In a few cases it has been reported that growth of microorganisms of silcone polymer coatings causes detoriation of the product, probably due to some degradation of some components of the industrial product (33-35). In studies with dimethyl silicones about 50 species of bacteria, fungi and yeast were unable to show any degradation (36). Although in various other studies with microorganisms no biodegradation has been found (37,38), it should be noted that so far most studies used bacterial growth, C02 production or 02 consumption as parameters. These parameters however are not very sensitive for poorly degradable compounds, and are usually unable to show cometabolism or partial degradation of compounds.
CONCLUSIONS. Since World War II the production of organosilicon
compounds has increased, the annual production levels now being higher than those of PCB's have ever been. However, the number of studies on the environmental fate of these compounds is very limited. For instance there is no monitoring programme for environmental concentrations, little is known about the ability of organisms to biotransform these chemicals or about the mechanisms and rates of exchange between atmosphere-hydrosphere and lithosphere. On the other hand, available information suggests that, despite the hydrophobic nature of these chemicals, bioaccumulation is not significant and that a-biotic transformation may degrade them at significant rates. However, the presence of organosilicon compounds originating from antropogenic sources in sediments and other environmental samples show that these chemicals do not have efficient sinks in the natural environment. Hence, despite the fact that most organosilicon compounds are only slightly toxic ( if at all ) to aquatic and other organisms (28,38), more attention should be paid to the fate of these compounds at least as long as the production volumes grow.
- 181-
REFERENCES 1. HOWARD,P.H.; DURKIN,P.R.; HANCHETT,A.,(1975), Environmental
hazard assessment of liquid siloxanes (silicones) EPA-report 560/2-75-004.
2. PEL~ENBARG,R.E.; TEVAULT,D.E.,(1986), Environ.Sci.Techno1. 20, 743-744.
3. WESCHLER,C.J. ,(1981), Atmos.Environ. 15, 1365-1369. 4. DeGROOT,R. (1979), H20, 12, 333-336 • 5. WAGGOTT,A. (1981), Trace organic substances in the River Lee,
in Chemistry in water reuse, vol 2, (Cooper,w.J. ed), Ann Arbor Science, pp 55, Ann Arbor.
6. FERRAND,R.; MAZZA,M.;PAYEN,P. (1978), Multidetection approach to analysis of organic pollutants in water; Methods and comment on results from aquatic pollutants: Transformation and biological effects ( Hutzinger,O.;VanLelyveld,L.H.; Zoeteman,B.C.J. eds), Pergamon Press, pp 87, Elmsford.
7. VANDERPOST,D.C.;TOERIEN,D.F. (1974), Water.Res. 8, 593-599. 8. MATSUI,S.;MURAKAMI,T.;SASAKI,T.;HIROSI,Y.;IGUMA,Y. (1975),
Prog.Water.Technol. 7, 645-649. 9. WATANABE,N.;NAKAMURA,T.;WATANABE,E.,SATO,E.;OSE,Y. (1984),
Sci.Total.Environ. 35, 91-97. 10.PELLENBARG,R.E. (1979), Environ.Sci.Technol. 13, 565-569. 11.PELLENBARG,R.E. (1979), Mar.Pollut.Bull. 10, 267-269. 12.MORRIS,R.J.; DAWSON,M.E.; LOCKWOOD,A.P.M., (1982),
Mar.Pollut.Bull. 13, 13-18. 13.KROTOSZYNSKI,B.K.; O'NEILL,H.J. (1982), J.Environ.Sci.Health
A17, 855. 14.BRUGGEMAN,W.A.; WEBER-FUNG,D. OPPERHUIZEN,A.,VANDERSTEEN,J.;
WIJBENGA,A.; HUTZINGER,O. (1984), Toxicol.Environ.Chem. 7, 287-296.
15.FRYE,C.L., (1983), Soap,cosmet.Chem.Spec. 59, 34. 16.COX,T.S.; INGEBRIGHTSON,D.N. (1976), Environ.Sci.Technol. 10,
598-601. 17.DELMAN,A.D.; LANDY,M.; SIMMS,B.B. (1969), J.Poiym.Sci.
AI, 375-378. 18.ZHUZHKOV,E.L.; AKHUNDOVA,L.A.; VOEVODSKII,V.V. (1965),
Kinetr.Katal. 6 , 637. 19.LENTZ,C.W. (1980), Ind.Res.Dev. 22, 139-143. 20.ABE,Y.; BUTLER,G.B.; HOGEN-ESCH,T.E., (1981),
J.Macromol.Sci.Chem. A16, 461-471. 21.FUNG,D.M.L., (1982), Poly(dimethylsiloxanes), surfactants and
polychlorinated biphenyls. Ph.D. Thesis, Guelph. 22.VANDERPOST,D.C. (1978), Wat.Pollut.Control. 78, 520-524. 23.FRYE,C.L. (1978), Fifth International Symposium,
Organosilicone Chemistry, Karlsruhe, Germany. 24.FRYE,C.L. (1980), J.Organometallic Chem. 17, 252-260. 25.BUCH,R.R.; LANE,T.H.; ANNELIN,R.B.; FRYE,C.L. (1984),
Environ.Toxicol.Chem. 3, 215-222. 26.BUCH,R.R.; INGEBRIGHTSON,D.N. (1979), Environ.Sci.Technol.
13, 676-697. 27.SAKIYAMA,M.; OKAWARA,M. (1965), J.ORGANOMET.CHEM. 2,473-477. 28.AUBERT,M.;AUBERT,J.;AUGIER,H.;GUILLEMAUT,C. (1985),
Chemosphere 14, 127-138. 29.HOBBS,E.J.; KEPLINGER,M.L.; CALANDRA,J.C. (1975), Environ.
-182 -
Res. 10, 397-406. 30.0PPERHUIZEN,A.; DAMEN,H.G.W.; ASYEE,G.M., VANDERZTEEN,J.M.D,
HUTZINGER,O. (1987), Toxicol.Environ.Chem., 13, 265-285. 31.FESSENDEN,R.J.; HARTMAN,R.A. (1970), J.Med.Chem.13, 52-54. 32.FESSENDEN,R.J.; AHLFORT,C. (1967), j.Med.Chem. 10,810-812 33.INOUE,M. (1973), Plast.lnd.News Japan, 19, 17. 34.ZHARIKOVA,G.G.;MARKELOVA,S.I.;BOBKOVA,.T.S.;LANDAU,N.S.,
SMOLINA,G.S.;SILAEV,A.B. (1971), Prikl.Biokhim.mikrobiol. 7, 236. .
35.BlLAI,V.I.;KOVAL,E.Z.;PASHENKO,A.A.;KRUPA,A.A.; SVIDERSKII,V.A. (1978), Geol.Khim.Nanki , 4 , 322.
36.YANAGI,M.; ONISHI,G. (1971), J.Soc.Cosmet.Chem. 22, 851. 37.SHARP,R.F.; EGGINS,H.O.W. (1970), Int.Biodetor.Bull,6,19. 38.FIRMIN,R.; FRYE,C.L.; RAUM,A.L.J. (1984), Ecotoxicol.test.
Mar. Environ. , 1, 591-623.
-183 -
POLLUTION OF SARONICOS GULF (ATHENS.GREECE) BY FOSSIL FUEL
HYDROCARBONS
A.Mylona, N.Mimicos, NRC DEMOCRITOS, Athens-GR E.Stephanou, Lab. for Environ. Chemistry, UNIV.of CRETE,Iraclion-GR
Abstract
Sediments sampled within Saronicos gulf (Greece) were used as geoaccumulators for the assesment of the status of petroleum pollution in this area. Molecular markers, namely were studied to identify the source of this pollution. The results were compared with those obtained from a spectrofluorometric analysis.
INTRODUCTION Sediments are considered as pollutants sinks because they provide
an substantial picture of events arriving in the water column. In this respect sediments have been used to study the petroleum contamination of coastal environment (1). Petroleum derived hydrocarbons can be contributed by different sources, e.g. atmospheric fallout, urban and industrial effluents, accidental or intentional tanker discharges. The problem of fingerprinting of hydrocarbons in the marine environment, to identify the sourc~ of oil spillages (accidental or intentional), has been of major concern in the last years and many analytical methods were developped for this purpose (2).We repport here the application of two analytical techniques, namely the study of molecular markers (hopanes) using CGC and CGC/MS analysis and spectrofluorometric analysis.
Saronicos Gulf presents a particular interest as it is considered a havely polluted area, the main pollution sources of which are: - ~wage outfall·, at Keratsini bay, which discharges the major part of the
domestic waste and industrial effluents of·the Athens Metropolitan area 6 . (,3,S.10 habitants).
- Port of Piraeus, one of the most important of the Mediterranean Sea. - Outfalls af about 30 major factories among which are 2 rafineries,
2 shipyards, steel mills etc.
EXPERIMENTAL
§~!!!E~!~~ Samples of the sediment surface were taken in 1986 at locations i~
Saronicos Gulf (Fig.1). All sediments samples were collected by a O,lm Van Veen grab sampler. After retrieval of the sampler the water was allowed to drain off avoiding disturbing the surface layer of the samples. Samples of approximately SOD ml were taken carefully from the top S cm layer through the hatch in the top of the sampler and immediately wrapped in aluminium foil and stored at -lSoC until analysis.
- 184-
!~£~~!!£~_E£~£~~~~_~£E_§E~£!£~~~~~E~~~!E!£_~~~~~~!~ All reagents used were p.a. grade. Especially the solvents used
(n-He, n-Pen and MeOH) were redistilled in glass apparatus in order to insure the best quality. Adsorbents (S.G. and Alumina) as well as boi~ ling stones were extracted with CH3Cl-MeOH (2:1) in Soxhlet apparatus.
Before analysis the samples were freeze-dried and sieved through 200 mesh sieve. 5-20 g of the 200 mesh homogenated sediment samples were saponoficated with methanolic KOH for 1,5 hours followed by extraction with n-pentane. The n-pentane phase was reduced in a Rotavapor to 0,5 ml, the sample was transfered to a glass vial for further reduct-ion to 0,2 ml using purified dry N2• Separation of alkanes from aromatics using an. Alumina S.G./column was followed. The synchronous spectra, by means of Perkin-Elmer MPF-44A spectrofluorometer,(~A=23nm) of the aromatic fractions were taken in order to obtained a deeper insight of its composition.
!~£~~!f~~_E£££~~~£~_~££_g~g_~~~_g~gL~§_~~~~~~!~ Sample extraction and compounds isolation were proceeded according
to Giger (3). Briefly hydrocarbons were Soxhlet~xtracted from dry sediments with methylene chloride. The solution was reduced under pressure and sulfur was removed from neutral fractions by addition of activated copper. The mixture of hydrocarbons was chromatographied on Sephadex LH-20. Two fractions were collected. The first fraction was further reduced in a rntavapor and purified on fully activated silica gel. The first fruction eluted with n-pentane, contained the aliphatic hydrocarbons. The second Sephadex fraction was also reduced in a rotavapor and eluted with n-pentane and then methylene chloride. This fraction, eluted with methylene chloride,contained the PAH.
A Finnigan mass spectrometer Model 4000 with an INCOS 2000 data system was used for mass spectrometric determinations. The Carlo Erba 4000 equiped with a Grob - type split-splitless injector, contained a glass capillary column (SE-54, 25 m x 0,25 mm) coupled directly to the ion source by a fused silca capillary. Helium was used for the carrier gas with a back press~re of 0,8 Atm. The temperature programme used was 70°C (1 min), 70-200 C (10°C/min), 200-280oC (2°C/min) and 2800 C (20 min). For the electron impact ionization mode the conditions were the following: ionization energy 70 'eV; ionizer temperature, 250oC; mass range, 35-590 m/z; scan time, 1,9 sec; electron multiplier voltage, 1700 V. For quantitative alalysis 1-chloro-hexadecane and 5,6-dimethyl-phenathrene were used as internal standards.
RESULTS AND DISCUSSION
Sampling stations located in Saronicos Gulf were selected (Fig.1) for comparison between the GC/MS and spectrofluorometric analysis. In table 1 are summarized the GC/MS and UV-spectrofluorometric analysis results.Figs 2 and 3 show the synchronous spectra of aromatic hydrocarbon fractions of surface sediments samples. The spectra of the first fraction+(n-pentane:dichloromethane, 7:3) show the presence of naphthalenes while the presence of aromatic campounds containing three and four fused rings are shown in fig 3 (2 fraction:dichlomethane).
tFig.2
-~-
In addition, the presence of more highly condensed aromatics, in lesser amounts, is indicated by fluoresence bands at 400, 440 and 460 nm (Fig.3). Such a wide range of aromatic compounds is typical of the aromatic composition of petroleums. The GC/MS analysis gave the same results. Fig.4 and 5 show,respectively, n-alkanes distribution and reconstructed mass fragmentogram of hopanoid hydrocarbons mixture, representative of the whole group of samples. The n-alkane distribution presented in Fig.4 exhibit a bimodal profile. The first mode ranging between n-C 20 and n-C 26 has no carbon number preference. The second ranging, between n-C 27 and n-C35 is characterized by an odd number slight predominance, maximizing at n-C 31 • Such a pattern corresponds to a dual origin: The first distribution indicated a crude oil contribution and the second a terrestrial plant wax imput. Fig.5 shows hopanoid hydrocarbons mixtures obtained by mass fragmentographic reconstructions of m/z 191 ions (2). Also the hopane distribution reflects both biogenic microbial and anthropogenic (petroleum) origin. The presence of C32-C 35 extended hopane 22R and S doublets is characteristic of crude 011 pollution (4). The aromatic hydrocarbon fraction isolated from the sediments contained a series of polycyclic aromatic hydrocarbons (PAHs) derived mainly from anthropogenic (combustion) sources (5) (Fig.6).
These results suggest that there is an overall pollution by the same type of petroleum products in the studied area. The constancy in the sedimentary hopane distribution supports this interpretation.
REFERENCES
1. J.W. Farrington and B.W.Tripp, Geochim.Cosmochim. Acta 41, 1627 (1977)
2. J.Albaiges and P.Albrecht, Intern. J.Environ.Anal.Chem. ~, 171 (1979). --
3. W.Giger and C.Schaffner, Anal.Chemistry 2Q, 243 (1978).
4. M.Dasti11ung and P.Albrecht, Mar.Pollut. Bull.~, 13 (1976).
5. L.Viras, P.Siskos and E.Stephanou, Intern.{.Environ.Anal.Chem. ~, 71 (1987).
Table I. U.V.F.aud GC/MS Analysis Results of Surface Sediments Samples of Saronicos Gulf.
UVF-results GC/MS Results Sample in llg/g d.w. 1: n-alkanes CPI Hopanes Station ppb C20-C26 C2T C35 C32-C35 (22R&S)
Sd/8/86 2,49 1682 0,5 2,5 Present
Sd/9/86 5,26 446 0,6 1,1 Present
Sd/13/86 4,83 3079 0,9 2,1 Present
-186-
'" c: 2 c:
III u c III u III III Lo j
i:L
III >
III u C III
'u\ III Lo j
lL
III > +' .2 III a:
-.:.-
"",-- '.. ' .. ~ . • ~. y '"",,_AJ"'~
.:,.= ' .. ,';;;;"evs ,l ~ , ....... .
- i ..... kt ... ..,- i . •
""", ',:. . .,
Fig. 1 - Map of Saronicos Gulf
Fi g,2
290
Fi g. 3
290
" , / I .. ~ .. ,
330
330
I
, I
I I
I
370
__ nm
410 450 _ nm
-187 -
~-\
"'- \ , ,
, "
, "\
" '. '. \
490
, , , , , , ,
" , \
, ,
,
"
, , , , "
- , .::: ...
530
100
22-45 2600 29:15 32:30 3W 3900 ';2115 .;8:';5 min
Fig. 4 - Typical reconstructed ion chromatogram (m/z:85) of n-alkanes
100
mJz 191
52:00 TIME
Fig. 5 - Hopanes (m/z:191) ion chromatogram
X 202
~228
252
"£Z276
26:00 32:30 39:00 ';5:30 min
Fig. 6 - Reconstructed ion chromatograms (m/z:202, 228, 252, 276) of PAH
- 188-
SUMMARY
THE SANDOZ/RHINE ACCIDENT
The Environmental Fate and Transport of Twenty-one Pesticides Introduced to the Rhine River
P.D. Capel and W. Giger
Swiss Federal Institute for Water Resources and Water Pollution
Control (EAWAG) CH-8600 Dtibendorf, Switzerland
On 1 November 1986 a fire at a Sandoz warehouse resulted in a massive input of chemicals to the atmosphere, Rhine River and surrounding soils. The wave of chemicals in the river, predominately Disulfoton and Thiometon, was monitored as it traveled to the North Sea. This chemical wave resulted in the death of benthic organisms and fish, especially eels. From the measured and estimated concentrations, Etrimfos, Endosulfan and Formothion were probably the lethal chemicals. The measured water concentrations from the period directly following the accident suggest that some of the chemicals were lost from the main wave before it reached the North Sea. Based on the physical, chemical and biological properties of the pesticides and the hydraulic characteristics of the Rhine, the fate and transport of the chemicals have been modeled. It is predicted that most were quickly flushed from the river. The exceptions are the mercury compounds and possibly Endosulfan. Other than these there should be no residual of the chemicals in the water, sediment or biota.
On 1 November 1986 a fire at a Sandoz AG warehouse in Schweizerhalle, an industrial area near Basel, Switzerland resulted in chemical contamination of the atmosphere, the surrounding soils and the Rhine River. Pesticides and other chemicals were discharged into the Rhine River causing massive kills of benthic organisms and fish, particularly eels for 200 km downstream (1). It is estimated more than 500,000 fish were killed (2). The warehouse, which was completely destroyed by the fire, contained pesticides, solvents, dyes, and various raw and intermediate materials. The majority of the more than 1300 tonnes of stored chemicals (3) was destroyed in the fire, but large quantities were introduced into the atmosphere, into the Rhine River through runoff of the fire-fighting water and into the soil at the site. Public and private reaction to the fire and subsequent chemical spill was strong. Even though this was "one of the worst chemical spills ever" (4), the nature of the chemicals and the powerful self-cleansing mechanisms of the river have made the predictions of a long-term "dead" Rhine unfounded. The recovery of the Rhine from this accident is well underway, but the problems from chronic chemical contamination still remain.
As with most rivers which drain industrial and agricultural areas, the Rhine suffers heavily from chemical pollution. The large chemical and min-
- 189-
ing industries located within the basin contribute extensively to this contamination. It has had problems for decades with heavy metals, dissolved solids, and organic chemicals. Over the past decade, the river's condition has improved tremendously due to statutory and voluntary controls on industrial inputs. Heavy metals (As, Cd, Cr, Cu, Hg, Pb, Zn), monitored at a minimum of· six stations, have shown steady decreases over the time period 1975-1985 (5). Even though the Rhine's condition has improved considerably in the previous decade, the problem of chemical contamination still exists. Based on average flow and concentrations, it is calculated that in 1985 2,500 tonnes of Zn, 270 tonnes of Cr and Pb, 8.9 tonnes of Cd and 3.1 tonnes of Hg flowed out of the Rhine.
SUMMARY OF ACCIDENT The warehouse contained at least 90 different chemicals, 21 of which
were pesticides, around which most of the environmental concern centered. The other chemicals seemingly posed only minor environmental problems because they degraded quickly and/or have low toxicity.
The fire probably started in a lot of Prussian blue dye, that had been packaged the previous day. The flames were detected at about 00:30 on 1 November 1986 and were extinguished by 06:00. The fire-fighting water (7-10,000 m3) was discharged into the Rhine, carrying chemicals into the river at Rhine km 159.1. Some entered the water column directly; some fell to the river bottom as dense, immiscible, chemical globules (mixtures of pesticides, dyes and solvents). The exact mass of the chemicals entering the Rhine is unknown. Estimates have been made by the German, Swiss and French authorities based on measured water concentrations.
The plume of chemicals, intensely colored due to Rhodamin B dye, hugged the south shore of the river until it reached a dam at km 163.8. After that dam, transverse mixing of the chemicals was nearly complete (8).
From the measured and estimated chemical concentrations at Village-Neuf, a prediction can be made as to which chemicals most likely killed the fish and other aquatic life. A comparison of water concentrations to the E~50 and LC50 values suggest that Endosulfan, Formothion and Etrimfos are probably responsible for the death of the biota. Mercury, DNOC, Fenitrothion and Parathion could have also contributed. Probably, there was also a synergistic toxic effect resulting from the mixture of pesticides. A strong correlation was found between the sum of the organophosphorous pesticides water concentration and the measured Daphnia toxicity (6). The chemicals mentioned above, contained in the same parcel of water as the major organophosphorous compounds (Disulfoton and Thiometon), were primarily responsible for this effect. '
The total mass of each of the monitored chemicals diminished as it flowed downstream. There are at least two explanations for this observation. Either the environmental removal processes (biological, physical and/or chemical) were fast enough to decrease the mass of chemicals in the Rhine or the losses are due to the hydraulic characteristics of the river. It is known that there are stagnant zones in the river, in the old river bed, behind the dams and in the French canal system, in which the chemicals were observed (7). If a portion of the chemicals were caught in these stagnant areas, then the main wave of chemicals would be diminished as it moved down the river. The portion of the chemicals retained in the stagnant areas would slowly find its way back to the main channel and be removed from the river. The net result would be a longer residence time of the chemicals in the water.
-190 -
ENVIRONMENTAL FATE PROCESSES There are a finite number of processes which must be considered to de
scribe or predict a chemical's environmental transport and fate. Each of these processes will be considered first in general terms, then specifically for these pesticides and the conditions of the Rhine River in November, 1986. The water chemistry, temperature and residence time of the chemicals in the water will have a significant impact on the fate processes.
After a chemical substance enters the water, a number of processes will act to decrease its concentration. Chemical (hydrolysis, photolysis, oxidation) and microbiological transformations can occur. The substance can be transferred to the atmosphere via volatilization. It will distribute itself between the water, suspended solids, and living biota. The particles, with their associated chemicals, can sink through the water column and be incorporated into the sediments. Marco-biota can metabolize the chemical or accumulate it within its tissue. The latter, however, results in a very minor loss. The fate of the products of chemical and biological transformation will also be governed by the same processes.
The hydraulic characteristics of the Rhine River contribute an important self-cleansing mechanism from chemical contamination. The residence time of water from Basel to the North Sea is about 12 days. Water currents scour the river bed, eliminating significant long term sedimentation of fine particles in the main channel, except in areas behind many of the dams in the upper stretches of the river. This scouring action and the short water residence time continually purge contaminants from the river into the North Sea. If a chemical is not strongly sorbed, nonvolatile, and is biologically and chemically recalcitrant, it will remain unchanged in the water column. It will undergo dilution and be removed from the river at a rate about equal to the water residence time. A chemical which is strongly sorbed will have a longer residence time in the river. Its transport will be influenced by sedimentation to the river bed, resuspension and subsequent particle transport down the river.
ENVIRONMENTAL FATE PREDICTIONS OF THE PESTICIDES IN THE RHINE A one dimensional model of chemical movement and fate in the Rhine
River, which incorporates hydraulic and environmental processes, has been developed (8). Disulfoton and Thiometon have been used to validate the model, since the greatest number of field (Rhine) measurements were made for these two compounds. For both, the chemical processes are much slower than biodegradation. A zero-order decay constant best fit the observed Rhine data (1.5 x 10-8 g m- 3 sec-1 for Disulfoton) (8). The disappearance rates in the Rhine, calculated by the best fit of the field data, and the biodegradation rates in Rhine water observed in the laboratory agree within -10% (8,9). The predicted and model concentration profiles for the four Rhine stations are illustrated for Disulfoton and Thiometon (Figure 1). Both the loss of chemical mass and the observed tailing of the concentration/time profiles of Disulfoton and Thiometon are adequately described by the model. The effect of the convective period (initial spreading) and the exchange of water between zones of flowing and stagnant water. Retainment of the chemicals in the areas of stagnant water diminished the concentration in the maximum wave. These retained chemicals slowly bleed out of the stagnant areas and were eventually flushed to the North Sea.
This strong confirmation of the model allows predictions to be made for the other chemicals. Based on the environmental half-lives of biodegradation, hydrolysis, photolysis and volatilization fate of the chemicals
-191-
<0
:J 0 DISU LF OTON c.n c 0
.... )0 .,. E ;>
c X 2 0 ~ ::E: N ~ C Q ..-. '0
0 Q) 20 ::;:: C c 0 C
0 .. :::c ~
c .r::::. .. ....... u c 0 u
10
0
14 . 1
Date (Hoven-be •• 1986)
20 :J
TH IOMETON 0 en c 0
~ .... '" ...... ;> 1 S E e 2 X ~ 0 2 ::E: " N ~
'0 c
e 10 0 4-0 ::;:: Q)
~ C c -, C .. " 0 u :::c c 0 u
14.11
Date (Hovell'btr. 1ge6)
FIGURE 1 - Measured (dashed) and modeled (solid) time/concentration profiles of Disulfoton and Thiometon in the Rhine.
-192 -
have been estimated. Some of the chemicals were completely lost within the river; others were totally recalcitrant. All of the chemicals, except those which are completely removed, are predicted to have time/concentration profiles of essentially the same shape as Disulfoton and Thiometon. Table I summarizes the mass of each of the pesticides passing the four German sampling stations from Figure 1. For those chemicals which passed out of the Rhine, the concern is transferred to their impact on the North Sea. The weakness of these estimates lie in the accuracy of the environmental rate constants. Usually the rate of biodegradation is known with the least degree of certainty.
Disulfoton, Thiometon and Etrimfos were measured in the sediments after the accident. Since Disulfoton has only a weak tendency to sorb iKoc -1600 mL/gOC) and the others even less (Thiometon Koc-340, Etrimfos Koc 570 mL/gOC), these chemicals should not have been found in the sediments, if sorption was the controlling process. A more reasonable explanation (than simple sorption and sedimentation) would be the movement of the chemical globules along the river bottom due to current action. The presence of globules implies an extremely non-homogeneous sediment concentration. The majority of the water burden left the Rhine by November 12th (Figure 1), but the chemicals in the sediments had only moved about 170 km (1/5 the distance) by then (6). This slow moving sediment burden, continually being reintroduced into the water, probably contributed to the asymmetry of the time/concentration profiles (Figure 1).
The four metal-based pesticides must be considered separately from the organic ones. The two zinc-based pesticides hydrolyze very quickly. The mass of zinc resulting from the accident is unimportant compared to the typical daily load passing Basel because the accident contributed less than 3% of a daily load. The mercury compounds are of more concern. Both of these will dissociate in water and exist as organomercury cations (phenylmercuric and ethoxyethylmercuric cations). They both will strongly sorb to particles or to dissolved humic material. The observations of elevated mercury levels in the sediments near Basel confirm this (6). The mercury should have a relatively long residence time in the sediments, but eventually it will be transported down river by current action or biologically transformed to neutral organomercury compounds (i.e. methylmercury), which have the potential to be transferred to the water and be bioconcentrated. The sediments near the site have mercury concentrations which are only about twice the normal "background" concentrations of Rhine sediments.
CONCLUSIONS Fortunately, this accident has not proved to be the long term eco
logical disaster that it was originally predicted by some. The Rhine river purged itself of all the chemicals (with the possible exceptions of mercury and Endosulfan) within a few weeks or months. There should be no residuals in the water, fish, or sediments for these chemicals. Some fish already returned to the Rhine at Basel by the Spring of 1987. However, problems of chronic chemical inputs still remain and this needs to be examined and dealt with in a continuous and systematic manner.
We need to learn from this tragedy and try to avoid other accidents which may have longer lasting effects. The suddenness and severity of chemical spills should be countered with fast reaction and remediation by environmental scientists. The ability to understand and predict a chemical's environmental behavior is crucial to this action and must be based on quantitative data (physical, chemical and biological) describing
-193 -
environmental processes. Unfortunately, there are tremendous gaps in this data. Much of the necessary information for these compounds and a myriad of other toxic chemicals are not available in the literature. Those data which do exist are, at best, difficult to obtain and many times inadequate, incomplete or conflicting. A well accessible data bank storing environmentally useful information on anthropogenic chemicals should be created. We would then be better prepared to contend with and predict the outcome of the next "ecological disaster".
REFERENCES
1) Rich, V, (1986) Nature, 324:201. 2) Deininger, R.L. (July, 1987) JAWWA, 78-83. 3) EAWAG (1986) Erster Zwischenbricht tiber Bestandesaufnahme, Oologische Beurteilung, empfohlene Massnahmen und Absichten ftir weitere Untersuchungen nach dem Schadenfall Sandoz im Rhein bei Basel, Dtibendorf, Switzerland, 12 December 1986. 4) Anon. (1987) Environ. Sci. Technol., 21:5. 5) Internationale Kommission zum Schutze des Rheins gegen Verunreinigung (1985) Tatigkeitsbericht 1985, ISSN 0173-6531. 6) Deutsche Kommission zur Reinhaltung des Rheins (1986) Deutscher Bericht zum Sandoz-Unfall mit Messprogram, December 1986. 7) Rapport du Comite d'Experts sur la Pollution Transfrontiere du Rhin (1986) Paris, 16 December 1986. 8) Reichert, P. and Wanner, O. AIHR-CONGRESS-IAHR, Lausanne, Switzerland, 1987. 9) EAWAG (1987) Zweiter Zwischenbericht tiber Verhalten der Chemikalien im Rhein, biologischer Zustand und Wiederbelebung des Rheins nach dem Brandfall in Schweizerhalle, Dtibendorf, Switzerland, 31 August 1987.
TABLE I - Estimated Mass (kg) of the Pesticides Passing Four Rhine Stations
Dichlorvos Etrimfos Fenitrothion Formothion Parathion Propetamphos Quinalphos
Captafol DNOC Endosulfan Metoxuron Oxadicyl Scillirosid Tetradifon
Estimated Input (kg)
1 290
2.5 3
50 160
6
1.6 1800
40 190
1900 0.3
40
Maximiliansau (km 362)
0 260
1.6 ?
46 160
5.2
0 1700
23 170
1900 ? ?
Mainz (km 498)
0 240
1.4 ?
45 160
4.9
0 1650
18 160
1900 ? ?
Bad Honnef (km 640)
0 230
1.2 ?
43 160
4.6
0 1600
15 155
1900 ? ?
Lobith (km 865)
0 210
1.0 ?
41 160
4.4
0 1550
13 145
1900 ? ?
--------------------------------------------------------------------------? denotes that not enough physical, chemical and/or biological was found to
make adequate predictions. ---------------------------------------------------------------------------
-194 -
OCCURRENCE AND LEACHING OF PESTICIDES IN WATERS DRATh-rNG FRG1 AGRICULTURAL LAND
S. REKOLAINEN National Board of Waters and EnviroI'llleIlt, Finland
SUJl1lla.rY
The occurrence and leaching of SClIle widely used pesticides were studied in 1985-1987 in an experiIrental field, in a small agricultural drainage basin and in a river highly loaded by agriculture. Phenoxy herbicides were found in mst of the samples and the highest concentrations were measuped in midsummer, soon after the application. High concentrations were usually associated with heavy rainfall events. Trace anounts of SClIle highly toxic insecticides could also be detected in the SaIOO sarrples.
1. It-.1TRODUCTION The mst important processes to be taken into account when evaluating
the effects of pesticides on aquatic ecosystems are leaching, degradation, accumulation and also toxicity. Pesticides with lOltl solubility in water and which are tightly adsorbed to soil particles can be washed out only by surface runoff. However, nany pesticides with high water solubility are nowadays widely applied. These can also be leached by subsurface drainage systems or they can infiltrate to ground water.
The time lapse between pesticide application and the first rainfall event and also the intensity of this rainfall are the mst important factors affecting the losses of pesticides fran agricultural land. A considerable part of total pesticide losses occurs during the first rainfall (3, 4). In addition to the chemical properties of pesticides and weather conditions, soil properties have a remarkable effect on pesticide leaching.
2. STUDY ARFAS Concentrations of pesticides were measured in an experiIrental field
in the years 1986-1987 and in a small drainage basin and in the river Aurajoki in 1985-1986.
The area of the experiIrental field (Kotkanoja) was 0.5 ha and it was divided into four equal parts. The surface runoff was collected fran the whole field, but the subsurface runoff was collected fran the four parts separately. The pesticide sarrples were taken in 1987 fran both surface and subsurface runoff waters, but in 1986 from the surface runoff only.
The LOyt~eenoja drainage basin is situated in western Finland. Its area is 5.6 kID , of which 77 % consists of agricultural land. The discharge is measured continuously and the pesticide samples were taken during high flOltl periods in sumner and early auhum.
The river Aurajoki is sit~ated in south-western Finland. The area of its catchr.ent basin is 885 kID , of which 41 % is agricultural land. Agriculture is the mst important single factor affecting the total transport
- 195-
of nutrients and suspended solids in the river.
3. RESULTS Phenoxy herbicides (MCPA, dichlorprop, rnecoprop) were found in mst
of the samplE!s (Tables 1-4). The_~iIrum concentrations of phenoxy herbicid~T measured were 4,5 ug 1 dichlorprop in the river Aurajoki and 3,2 ug 1 MCPA in Uiytfuleenoja, where the cultivation of sugar beet, cabbage, carrot, and other crop plants is intensive. Traces of some special herbicides (rnetamitron, terbuthylazin, trifluralin, dinoseb) and insecticides (dirnethoate, fenitrothion, malathion) were also detected. In the river Aurajoki sirnazine was also observed very frequentlY1 The concentrations of the insecticides varied fram 0.01 to 0.6 ug 1 . Of the pesticides studied only parathion ~d isophenphos were not detected in any samples (detection limit 0.01 ug 1 ).
Table 1. Detection of pesticides in surface runoff samples) fran the Kotkanoja experimental field.
Total Not MaxiIrum number of detected conc~tration samples u9: 1
Dichlorprop 8 6 0.45 MCPA 8 6 1.0 Dirnethoate 8 6 0.12
Table 2. Detection of pesticides in subsurface runoff samples fran the Kotkanoja experimental field.
Total Not MaxiIrum number of detected conc~tration samples u9: 1
Dichlorprop 8 1 0.22 MCPA 8 1 0.12 Dirnethoate 8 6 0.06
Table 3. Detection of pesticides fran the Uiytfuleenoja drainage basin.
Total Not MaxiIrum number of detected conc~tration samples u9: 1
Dichlorprop 8 4 0.3 Mecoprop 8 6 0.02 MCPA 8 6 3.2 Dirnethoate 8 3 0.6 Fenitrothion 8 3 0.06 Metamitron 8 4 0.11 Dinoseb 8 4 0.07 Trifluralin 4 0.02 Malathion 4 2 0.35 Terbuthylazine 4 1 0.17
The highest concentrations were measured in July and August after the application period, but phenoxy herbicides were also detected in late autUIIn1 in Aurajoki in 1985.
- 196-
Table 4. Detection of pesticides from the river Aurajoki.
Total Not Maximum number of detected conc~tration samples uS 1
Dichlorprop 11 5 4.5 Mecoprop 11 4 0.15 l«:PA 11 7 2.6 Dimethoate 11 7 0.03 Fenitrothion 6 6 Metarnitron 6 6 Dinoseb 11 6 0.03 Trifluralin 6 6 Malathion 6 6 Parathion 6 6 Terbuthylazine 11 10 0.03 Isophenphos 6 6 Endosulfan 6 5 0.02 S:i.rrazine 11 5 0.7
4. DISCUSSION It is obvious that widely used phenoxy herbicides occur in rivers
situated in intensive agricultural areas at least during the growing season. The observed concentrations are in good agreement with the observations made in agriculturally loaded rivers in SWeden (1, 2). The observed concentrations are considerably lower than the toxic levels for aquatic organisms, but the consequences of continuous low-level exposure are unknown. In spite of the relatively high degradation rate of the organiC phosphorus insecticides dimethoate and fenitrothion, trace aIOOunts of these compounds were detected in certain samples. Because of their high acute toxicity they may have same direct effects on aquatic ecosystems.
REFERENCES
(1) BRINK, N. 1985. Bekfutpningsrredel i aar och grundvatten. Miljo och Framtid 12: 10-13.
(2) KREUGER, J. &. BRINK, N. 1987. Bekfutpningsrredel i aar. Forskningsredogorelse 1986. MiJreographed.
(3) WAualOPE, R.D. 1987. Tillted-bed s:irrulation of erosion and chemical runoff from agricultural fields: II. Effects of formulation on atrazine runoff. J. Environ. Qual. 16: 212-216.
(4) WHITE, A.W. Jr., ASMUSSEN; L.E., HAUSER, E.W. & TURNBULL, J.W. 1976. Loss of 2,4-0 in runoff from plots receiving s:irrulated rainfall and from a small agricultural watershed. J. Environ. Qual. 5: 487-490.
-197 -
Summary
POLYCHLORINATED BIPHENYLS IN THE KUPA RIVER,
CROATIA, YUGOSLAVIA
1 Z. 3MIT, V. DREVENKAR and M. KODRIC SMrT
Medical Centre Sisak, Department of Sanitary
Chemistry and Ecology, Sisak, Yugoslavia
1Institute for Medical Research and Occupational
Health, University of Zagreb, Zagreb, Yugoslavia
The presence of polychlorinated biphenyls (PCBs) was investigated
in samples of water, suspended particles, sediments and fish from
the Kupa river, Croatia, Yugoslavia, along a river stretch
extending up to 10 km upstream and 200 km downstream of the primary
contaminated karst region. The PCB levels detected in the samples
collected downstream ranged from 1 to 52 ng dm-3 for water, from 50 -1 8-1 to 190 pg kg for suspended particles and from to 39 pg kg for
the sediment. A wide range of PCB concentrations, from 0.1 to
42.3 pg g-l, which were measured in edible portions of different
fish confirmed a long-term contamination of the river with PCBs. As
the Kupa river may be classified among low to moderately contaminated
waters, it is essential that the investigations of the presence and
behaviour of PCBs in the river and its environment be continued.
1. INTRODUCTION
Recently a very serious contamination with polychlorinated biphenyls
(PCBs) of a relatively narrow karst area in Slovenia, in the north-west
of Yugoslavia, has been reported (1-3). It is due to improper disposal of
waste by an electrocapacitor manufacturing plant taking place even since
-198 -
1962. The PCB migration from the waste tip into the karst ground water
and in the source water has been established. The PCBs have been further
introduced into the small Lahinja river, a tributary of the Kupa. The
Kupa river flows partly along the border line between two adjacent
Yugoslav republics Slovenia and Croatia. In Croatia the Kupa river and
the connected ground waters are water sources for the public water
supplies in the rural and urban areas lying along a longitudinal river
segment of about 200 km downstream from the primary contaminated area.
In the absence of data on PCB contamination of this part of the Kupa the
present investigations were initiated in 1985. In this contribution the
results are presented of PCB determination in samples of the river water,
suspended particles, sediments and fish, which have been collected in the
period from July 1985 to March 1986 at four different locations along the
river.
2. EXPERIMENTAL
2.1 Sampling
The sampling of the river water, suspended particles and sediment
was performed along the Kupa river at various distances from the primary
contaminated area. Location 1 was about 10 km upstream and locations 2,
3 and 4 were about 10, 100 and 200 km downstream of the Lahinja river
mouth.
Samples of the river sediments were collected from the upper 10 cm
layer of the sediment profile. The suspended particles were separated
from water samples by filtration of the water using Whatman No. 1 filter.
Samples of different kinds of fish caught along the monitored
segment of the Kupa river were kept frozen at -20 °c before analysis.
2.2 Procedures
Samples of 2 dm3 of the river water, 5-10 g of air-dried sediments,
air-dried suspended particles separated from 10 dm3 of the river water
and 2-5 g of edible fish portions were treated according to the procedure
recommended by US Environmental Protection Agency and with usual,
slightly modified, standard methods (4, 5). The combined extracts obtained
by multiple extraction of samples with n-hexane were purified on a silica
gel - sulphuric acid column. If necessary, n-hexane eluates were
additionally washed with concentrated sulphuric acid. The eluates were
evaporated to dryness under a stream of nitrogen, the compounds
redissolved in 1 cm3 of n-hexane and analysed by gas chromatography with
an electron capture detector.
The detection limits for PCBs in the analysed samples were 1 ng dm- 3
in water samples, 1 pg kg-1 in sediment samples, 50 pg kg-1 in suspended
particles and O. 1 pg g -1 in fish samples.
The PCB recovery from water samples was tested by spiking 1 dm3
deionized water samples with 500 ng of Pyralene isolated from discharged
capacitors by n-hexane extraction. The recovery calculated on the basis
of Aroclor 1260 was 26-3~/o.
2.3 Gas chromatographic analysis
Each sample was chromatographed on two columns of different polarity.
The compounds were identified by their retention times as compared to
standards prepared by dissolving known amounts of Aroclor 1260 in
n-hexane. PCB concentrations were calculated by summation of heights of
eight major peaks on the first column and of 14 major peaks on the second
column.
To evaluate PCB patterns in selected sample extracts and in the
Pyralene extracted from discharged capacitors a high resolution gas
chromatograph-mass spectrometer-computer system (HRGC-MS) was used.
3. RESULTS AND DISCUSSION
The PCB profile determined by HRGC-MS analyses in extracts of
sediment and fish and in Pyralene (Fig. 1) was almost identical to the
previously reported profile of the PCBs extracted from a sediment sample
collected in the heavily contaminated area in Slovenia (2). The absence
of PCB residues in samples of water, suspended particles and sediment
collected upstream of the Lahinja river mouth (Table I) also points to
the contaminated karst area in Slovenia as the source of the PCB presence
in the Kupa river.
The amounts of PCBs in unfiltered water samples collected in July
1985 were approximately one order of magnitude higher than those
determined in the water samnles collected later (Table I). This could be
- 200-
TIC
0/0
50
o 0.01
Fig.1
\
X =5
X =6 .
C12 H1O- X CI X
X=4 --..
X=2 X=3 X=7
7111 ~ ~--4.57 9.55 14.53
time (min)
A typical total ion current chromatogram of PCBs in extracts of sediment and fish from the Kupa river
a consequence of entirely different specific hydrological conditions i.e.
of the input of the snow-melting and rain-off waters from the primary
contaminated area into the Kupa river at the time of sampling.
While PCB concentrations in the river water samples collected from
September to December 1985 were almost identical at all three locations
downstream of the Lahinja river mouth, the PCB amounts detected in the
suspended particles andsedimenmdecreased with the increasing distance
from the primary contaminated area (Table I). By tentatively performed
simultaneous analyses of PCB residues in the Kupa river water and in the
suspended particles separated from the same water samples a distribution
coefficient of the order of magnitude up to 104 between particles and
water was determined. Suspended particles were obviously an important
route in the transport of PCB residues over a considerable distance from
the primary contaminated area. Their varying amounts in the run-off waters
from this area could also greatly contribute to the periodical decrease/
increase of the PCB level in the Kupa river water.
The results presented in Table I indicated a significant accumulation
of PCBs on the river sediment. The sediment could continue to be a stanmmg
- 201-
source of contamination of the water environment, and could be activated
especially during.the intensive rising of the river water.
A wide range of PCB concentrations and the tentatively calculated
bioaccumulation factors of the order of magnitude from 104 to 106 in
edible portions of the fish samples from the monitored segment of the
Kupa river confirmed a long-term river contamination with PCBs (Table I).
Compared to the current US Food and Drug Administration tolerance limits
of 2 pg g-1 for edible portions of fish (6) the fish in the Kupa river
could be considered to be moderately to highly contaminated.
The results of this work showed that the concentrations of PCB
residues in the Kupa river along a 200 km longitudinal segment downstream
of the Lahinja river mouth were lower by 1-2 orders of magnitude in the
river water and by 3-4 orders of magnitude in the river sediment than in
the source water (300 ng dm-3) and respective sediment (55000 pg kg- 1) in
the contaminated karst area (2). According to the World Health Organization
criteria for PCB concentrations in fresh waters (7) the Kupa river may be
classified among the low to moderately contaminated waters. Owing to thffir
well-known persistence and bioaccumulation in the environment the ~
of PCBs in the monitored segment of the Kupa river should be expected to
last for many years to come. Therefore the present investigations should
be continued through systematic analyses of PCB load profiles in the river
depending on the hydrometegrological conditions in the primary contaminated
area and downstream as well as of PCB concentration in different profiles
of the river sediment and in various biological samples collected in the
river environment.
ACKNOWLEDGMENT
This work was financially supported by Republic Committee of Water
Economy of SR Croatia. The assistance of Dr. V. Svob and Mrs. M. Cvetko
during the HRGC-MS work is gratefully acknowledged. The extended version
of this paper is accepted for publication in Chemosphere.
REFERENCES
(1) Brumen, S., Medved, M., Voncina, E., Jan, J. and Malnersic, S., Hrana
i ishrana 25 (1984) 179-193 (in Croatian)
(2) Brumen, S., Medved, M., Voncina, E. and Jan, J., Chemosphere Jl (1984)
1243-1246.
- 202-
(3) Herlander, D., Zdrav. vestn. 55 (1986) 137-139 (in Slovenian).
(4) US Environmental Protection Agency, Manual of Analytical Methods for
Analysis of Pesticides in Humans and Environmental samples, EPA
600/8-80-038.
(5) Standard Methods for Examination of the Water and Waste Water, APHA
-AWWA-WPCF, Washington, 16th Ed. (1985), p. 538.
(6) US Food and Drug Administration, Tolerance for Polychlorinated
Biphenyls, Code of Federal Regulations 21 Part 109.
(7) World Health Organization, Polychlorinated Biphenyls and Terphenyls,
Environ. Health Crit., Vol. 2, WHO Geneva 1976.
Table I
PCB concentrations in river water, suspended particles, sediment and in
the edible portions of the fish along the monitored segment of the Kupa
river (based on Aroclor 1260)
Sampling location
PCB concentration
Sample range
pg dm- 3(N)
RIVER WATER
July 1985 4 4 - 52 (20)
Sept - ND (4)
-1* pg kg
____ :_~~~_22~2 ______ ~~2~~ __________ : __ ~_i2~2 ____________________________ _ SUSPENDED
PARTICLES
Dec 1985
SEDIMENT
Sept -
- Oct 1985
FISH
July 1985
1
2
3 4
2
3 4
- March 1986 - 4 N= number of samples; ND= not detected;
- 203-
ND
190
135
50
ND
39
7 8
0.1-42.3(30) *mean of two samplings.
SESSION III
TRANSFORMATION OF ORGANIC MICROPOLLUTANTS IN WATER
Chairmen J. ZEYER and A. BJORSETH
Biodegradation of chlorinated aromatic chemicals in continuous cultures
Anaerobic degradation, processes and test methods
The fate of organic compounds in the environment
BIODEGRADATION OF CHLORINATED AROMATIC CHEMICALS ill CONTINUOUS CULTURES
J.R. Parsons, D.T.H.M. SlJm and M.C. Storms Laboratory of EnVlronmental and TOxlcologlcal Chemlstry, Umverslty of
Amsterdam, Nleuwe Achtergracht 166, 1018 WV Amsterdam
Chemostat (contmuous) cultures of mlcroorgamsms offer a number of advantages for the study of the blOdegradation of xenobiotic chemicals compared to the more commonly used batch systems. Chemos tat cultures enable the mfluence of enVlronmental conditions and adaptation on blOdegradatlOn to be studied, as well as simpllfying kmetic studies.
The cometabolic degradatlOn of chloroblphenyls by a Pseudomonas stram m batch and chemos tat cultures was compared. In batch cultures there was ready degradatlOn of mono- and dichloroblphenyls, but not of tetrachloroblphenyls. In contrast, tetrachlorobiphenyls were degraded in chemostat cultures exposed contmuously to the compounds by means of a generator column. Degradation was apparently enhanced by the more favourable conditlOns in chemostat cultures, and not by adaptatlOn of the bacterla.
First order blOdegradatlOn rate constants for the tetrachloroblphenyls were calculated from the steady state concentrations m the cultures and the exposure concentratlOns. The rate constant for 2,2',3,3'-tetrachloroblphenyl was much higher than those for the 2,2',5,5'- and 2,2',6,6'- lsomers, suggesting that steric hmdrance of 2,3-dioxygenatlOn controls thelr degradatlOn rates.
INTRODUCTION Chlorinated aromatic chemicals are an lmportant group of organic
micropollutants. Although there have been many studies of thelr blodegradation [1), most of these have been llmited to the less chlorinated compounds. For example, biodegradation studles of polychlorinated blphenyls (PCB's) have in general concerned congeners containing up to four chlorine substituents [2,3), although Bopp recently reported the isolation of a Pseudomonas strain able to degrade biphenyls Wl th up to six chlorine substituents ['1
Batch culture techmques are most commonly used to study blOdegradation. Thelr main advantage is their experimental simplicity. However, it is not always appreclated that the conditions m batch cultures bear little resemblance to those in the environment. In batch cultures bacteria are initially exposed to large excesses of nutrients, resulting in rapid growth (often at their maximum rate) after a lag or adaptation period. Depletion of nutrlents and bulld up of metabolic products results m an ever changmg envlronment until growth is halted by exhaustion of an essential nutrlent or accumulation of toxic metabolites. In contrast, microorganlsms grow in the environment at less than maXlmum rates as they are always exposed to a nutrient limitation.
Contmuous culture systems, such as chemos tats, make lt possible to grow mlcroorganisms under controlled conditlOns. In a chemostat, microorga-
- 206-
nisms grow at a rate determmed by the rate at which medium 1S supplied [5), and under a continuous nutrient limitation, which controls the culture density. It is possible to expose a chemostat culture to different nutrient limitatlOns (e.g. carbon or nitrogen limitations).
Chemostat cultures enable continuous exposure of a growing culture to test compounds, thus increasing the chance of adaptation to the chemical. The influence of environmental conditions can also be studied. Another advantage of chemostat cultures is that the1r constant biomass concentration simplifies the measurement of biodegradation kinetic~.
To date, chemostat and other continuous culture systems have been relatively llttle used 10 biodegradatlOn research. Continuous flow reactors have been used as models of wastewater treatment systems to study the degradation kinetics of, for example, 2,4-dichlorophenol [6] and pentachlorophenol [7]. Veerkamp et al. compared the transformation of chlorobenzoic aC1ds 10 batch and chemostat cultures of a Pseudomonas strain [8]. They observed that 2- and lj,-chlorobenzoates were metabolized in chemostat cultures but not in batch cultures: 3-chlorobenzoate was transformed in both batch and chemostat cultures. Liu studied the influence of medium flow rate on degradatlOn of the PCB mlxture Arochlor 1221 in continuous cultures [9]. All the seven major components of the m1xture were degraded at low flow rates. Increasing the flow rate resulted in increasing concentrations of some of the dichlorobiphenyls in the cultures.
We report here some of the results obta1ned in a study of the degradation of PCB's in chemostat cultures of a Pseudomonas stram, and a comparison w1th their degradation in batch cultures of the same stram.
MATERIALS MID METHODS The work reported here was carried out usmg a Pseudomonas strain
isolated from soll with biphenyl as carbon substrate. This strain also grows on 2- and lj,-chlorobiphenyl, but not on more highly chlormated biphenyls. In order to obtam blOmass concentratlOns not limited by substrate solubility, benzoic acid or 3-methylbenzolc aCid were used as carbon sources. Cultures with a biomass concentratlOn of ca. 600 mgll were grown 10 carbon limited medium of the followmg compos1tion (per 1):
1.00 g benzoic aCid or 3-methylbenzolc aCid lj,.03 g Na2HPOlj,.12H2o 2.1lj, g KH2POlj, 1.00 g (NHlj,)2s0lj, 0.20 g MgSOlj,.7H20 0.07 g Ca(N03)2.lj,H20 0.03 g iron (Ill) ammOnium citrate 1 ml trace elements solutlOn (30 mgll Na2Blj,07.10H2o. 20 mg/l CO(N03)2. 10 mg/l ZnSOlj" 3 mgll (NHlj,)6Mo~2lj" 3 mg/l MnSOlj,. 2 mg/l Nie12 and 1 mgll CuSOlj,. 5H20)
Batch cultures experiments were carried out 10 air-tight sealed erlenmeyer flasks to 11ml t losses of PCB's by volatllizatlOn. PCB's were added as solutlOns 10 1 ml acetone or methanol once bacterial growth was vis1ble. IncubatlOns were continued until the cultures were well into the stationary phase of their growth curve (usually lj,8 or 72 h after 1Onoculation).
Experlments 10 chemostat cultures were carried out in a system similar to that described prevlOusly by Veerkamp et al. [8]. The aeration was limited to ca. 20 mljmm to a vOid excessive volatilizatlOn of PCB's. The chemicals were added to the cultures either as a pulse (dissolved 10 1 ml acetone or methanol) or contlOuously US10g a generator column (Fig. 1). The
- 207-
columns contained 15 g Chromosorb GAl! (It-5-o0 mesh) loaded with 100 mg It-chlorobiphenyl (It--CB) and ca. 5 mg each of 2,2',3,3'-, 2,2',5,5'- and 2,2',o,o'-tetrachlorobiphenyls (2,2',3,3'-, 2,2',5,5'- and 2,2',o,o'-CB, resp.).
Samples of cultures (typically 50 ml) were extracted with equal volumes of hexane. The hexane layer was removed after centrifugation and reduced in volume to ca. 1 m!. This extract then underwent clean-up by being eluted
water -
water .
medium
•
chromosorb + PCB's
I----i-r. eff luent
-++-- culture
septum
~~::!:==tI--. stirrer bar
at 250C L...L... __ -'-'
Fig. 1. Apparatus for contmuous exposure of a chemos tat culture usmg a genera tor column.
- 208-
through a column contammg ca. 1 g each of 100-120 mesh sHica + lJ,0% W /w H2S04 and silica + 331. w /w 1 N NaOH. Analysis was by GC-ECD (Tracor 550, 2i~
Dexil 300 GC on 160-130 mesh Chromosor'b 750, 1 ill X 2 mm or Hewlett-Packard 5830A, ECD, CpSll 5CB, 25 m x 0.32 mm). The recoveries of the PCB's were checked by analyzmg samples contammg Rnown quanti ties of these compounds and were routinely above 901..
Control experiments were carried out under stel'lle but otherwise lden tical conditions.
Chloroblphenyls were obtamed from Analabs (North Haven, Conn., U.S.A.). Solvents wepe distilled before use.
RESULTS AND DISCUSSION Batch cultures of Pseudomonas strain JB1, grown on benzoic aCid, showed
ready degradation of mono- and dlchloroblphenyls, but there was no eVldence for degradatlOn of tetrachloroblphenyls (Table I). The lacR of degradatlOn of tetrachloroblphenyls may be caused by them havmg very low degradatlOn pates, or by their degradatlOn reqUlrmg other enzymes than those mvolved in the degradation of lower chlormated congeners.
TABLE I DEGRADATION OF CHLOROBIPHENYLS BY BENZOATE-GROWN BATCH CULTURES OF STRAIN JB1
PCB
2-lJ,-2,5-lJ"lJ,'-2 ,2' ,3,3'-2,2' ,5,5'-2,2' ,6,6'-
H not detected.
Initial conc. (lJg/l)
3730 1520
120 37
200 200 190
Final conc. (lJg/l)
Cul ture Control
lJ,90 3220 n.d.!! 1690
0.lJ, 120 2.0 23
210 210 220 210 200 200
The first series of experiments in chemostat cultures was carried out by adding acetone 01' methanol solutions of chloroblphenyls to a culture of strain JB1 growing on benzoate. Typical results obtamed for lJ,-chlorobiphenyl (lJ,-CB) and 2,5- and 3,5-dichloroblphenyl (2,5-CB and 3,5-CB) are shown m Fig. 2, as are the results of a control experiment with 2,5-CB. The line marRed D indicates the calculated dHutlOn rate. There was no evidence for sigmflcant volatillzation of 2,5-CB. Neither was there evidence for volatllizatlOn of lJ,- and 3,5-CB in Similar experiments. These results indicate that all three chlorobiphenyls were degraded, and that their degradation followed first order Rinetics. This is expected in a chemostat culture m which the biomass concentration is kept constant and the concentrations of substrates are well below theH' half-saturation (KM) values.
Under these conditions, the rate of disappearance of a pulse of a chemical from a chemostat culture is given by
-de dt : (kb + kv + D) C ( 1 )
- 209-
where C 1S the concen t l' atlOn of tl1e chem1cal at t1me t, kb the f1rst order biodegradation rate constant, D the dllutlOn rate constant and Kv the vola tiliza tlOn rate constant (where applicable). D can be calculated from the medium flow rate and the volume of the chemostat (D = f/V). The values of kb calculated for the three chlorobiphenyls from the data shown m Fig. 2 are 15.8, 0.13 and 0.54 h-1 for 4-CB, 2,5-CB and 3,5-CB, respectively. There was no eVidence for degradation of the tetrachloroblphenyls in such experiments (data not given).
1000
100
u
10
1 o 2 4
o 4-CB o 2,S-CB • 2,5 -CB (control) .. 3.5-CB
• 0 : - 0.046
t(h) 6 8
Fig, 2. Degradation of 4-, 2,5- and 3,5-CB in chemos tat cultures.
-210-
The extremely low aqueous solublli tIes of the tetrachlor'Oblpheny Is means that a large proportlOn preCIpItates In the chemostat when a solutlOn In an orgamc cosolvent IS added. A generator column was used to expose a culture continuously to a mIxture of tetrachloroblphenyls, In order to aVOId problems arlSlng from theIr very low solub1l1 ty. MedlUm was pumped through a column contalmng Chromosorb coated wIth a mIxture of 4-, 2,2',3,3'-, 2,2',5,5'- and 2,2',6,6'-CB, and placed ImmedIately above the chemostat (Fig. 1). This techmque made it possible to expose a culture contlnuously to chiorobiphenyis In the dIssolved state for several months. ComparIson of the concentrations of chlor'Oblphenyls In the medium and those In the culture revealed whether degradatlOn occurred.
At a steady state in such an experIment, the rate at WhICh a chemIcal enters the chemostat IS equal to the rate at which It IS lost by dllutlOn, degradatlOn and volatilIzation (where applICable):
(2)
where Cm and Cc are the concentratlOns of the chemIcal in the medlUm and chemostat, respectively. The biodegradatlOn rate constant can be calculated from
= !{v (3 )
In an experiment with a benzoate-grown culture, no 4-CB could be detected In the culture while its medIum concentration was ca. 1.1 mg/l (detection level ca. 1 jJg/I, corresponding to !{b > ca. 44 h-1). The concentrations of tetrachlorobiphenyis in the culture were generally lower than those in the medium (FIg. 3). The largest concentration difference was found for 2,2',3,3'-CB, indicating that this compound was degraded most rapIdly.
The concentrations of tetrachlorobiphenyls In the medium and the culture varied during the course of the experIment (FIg. 3). Nevertheless, if the concentrations in the culture are in a steady state, SImIlar values of !{b should be calculated. However, these values varied considerably for the benzoate-grown culture. Changing the growth substrate to 3-methylbenzoate after 80 days improved the reproducibility of !{b' The mean values of Cm' Cc and kb for the 3-MeBA-grown culture are given in Table II.
The reason for the improved reprodUCIbility in cultures grown on 3-methylbenzoate is not clear, but IS POSSible that growth on 3-methylbenzoate, but not on benzoate, requires som~ of the same enzymes as degradation of chlorobiphenyls.
The fIrst reaction In the most common degradation pathway for chlorobiphenyls is dioxygena tion In adjacent ortho and meta posItions [2,3). 2,2',3,3'-CB has two sets of unsubstituted ortho and meta pOSitions, whereas 2,2',5,5'- and 2,2',6,6'-CB have none. Therefore, It is likely that the much lower degradation rates of the latter compounds compared to that of 2,2',-3,3'-CB is caused by steric hindrance of the initial dioxygenation reaction.
- 211-
20
10
20
10
o
20
10
Ch 2-u
o
o
o
50
50
2,2',3,3'-CB
t (d)
2.2:5,5'-CB
100 150 I (d)
2,2',6,6'-CB
100 150 t (d)
Flg, 3, Contlnuous exposure of a chemos tat culture to a mixture of tetra-chlorobiphenyls, 0 - concentration in medlUm;
• - concentratlOn in the culture ,
- 212 -
TABLE II CONCENTRATIONS AND BIODEGRADATION RATE CONSTANTS IN A 3-METHYL-BENZOATEGROWN. CULTURE
PCB Cm (~g/l ) Cc (~g/l ) kb (h-1 )
2,2' ,3,3'- 6.64 ± 0.43 0.46 ± 0.14 0.580 ± 0.180 2,2' ,5,5'- 16.23 ± 1. 52 11.65 ± 2.63 0.018 ± 0.009 2,2' ,6,6'- 5.25 ± 0.47 3.14 ± 0.57 0.028 ± 0.009
The enhanced degradation of the tetrachlorobiphenyls in chemostat cultures, compared to that in batch cultures may have been the result of adaptation of cultures exposed continuously to these compounds. Alternatively, the dlfference in the conditions In the cultures, espec1ally the continuous carbon limitatlOn in the chemostat cultures may have been respons1ble.
The poss1ble role of adaptation m the degradation of the tetrachloroblphenyls was investigated by comparmg the degradation of these compounds in 3-methylbenzoate-grown batch cultures moculated from a chemostat culture exposed to these compounds and cultures of bacterla not previously exposed. There was very little difference In the results in both types of cultures (Table III), lndicating that adaptation processes had not influenced the results m the chemostat cultures. Thus, it appears that condltlOns m the chemostat cultures were more favourable for degradation of the tetrachlorobiphenyls than those ln batch cultures.
TABLE III DEGRADATION OF TETRACHLOROBIPHENYLS IN 3-METHYLBENZOATE-GROWN BATCH CULTURES
PCB
2,2' ,3,3'-2,2' ,5,5'-2,2' ,6,6'-
Initial conc. (~g/l )
21!.5 27.1 26.0
Hnal concentration (~g/l )
Control Exposed ll Unexposed llll
22.9 10.7 11.2 25.0 22.5 21!.2 23.3 21.1 22.8
II culture of bacterla prevlOusly exposed to tetrachloroblphenyls m a chemostat culture; lIli culture of bacterla not previously exposed to tetrachloroblphenyls.
CONCLUSIONS Cometabollc degradatlOn of tetrachlorob1phenyls takes place In chemo
stat cultures of Pseudomonas strain JB1, but not in batch cultures, whereas mono- and dichloroblphenyls are degraded ln both batch and chemostat cultures. The enhanced degradatlOn by chemostat cultures 1S not the result of adaptatlOn of the cultures, but 1S the result of more favourable conditions, probably the contmuous carbon llm1tation.
First order b1odegradation rate constants can be determlned ln cultures exposed contmuously by means of a generator column. BlodegradatlOn rate constants of three tetrachlorobiphenyls lndlcate that sterlc hmdrance of the im tial dlOxygena tion reaction determmes their blOdegradatlOn rates.
- 213-
REFERENCES 1. J.M. Wood, Environ. Sci. Technol., 16(1982)291A-297A. 2. K. Furukawa m A.M. Chakrabarty (Ed.), Biodegradation and
Detoxification of Environmental Pollutants, CRC, Boca Raton, Fla, U.S·.A., 1982, pp. 33-57.
3. J. Parsons, W. Veerkamp and O. Hutzmger, Toxicol. Envlron. Chem., 6( 1983)327-350.
4. L.H. Bopp, J. Ind. Microbiol., 1(1986)23-29. 5. D . W. Tempest ln J. R. Norrls and D. W. Ribbons (Eds.), Methods in
Microbiology, Vol. 2, Academlc Press, London, New York, 1970, p. 259. 6. P. Beltrame, P.L. Beltrame, P. Carmti and D. Pitea, Water Res.,
16(1982)429-433. 7. L.P. Moos, E.J. Kirsch, R.F. Wukasch and C.P.L. Grady, Jr.,
Water Res., 17(1983)1575-1584. 8. W. Veerkamp, R. Pel and O. Hutzinger, Chemosphere, 12(1982)1337-1343. 9. D. Liu, Bull. Environ. Con tam . TOX1COl., 29 (1982) 200-207.
- 214-
ANAEROBIC DEGRADATION, PROCESSES AND TEST METHODS
SllIIIIIIary
G. SCHRAA Department of Microbiology
Agricultural University Wageningen, Netherlands
Several aspects of the anaerobic biodegradation of organic compounds are presented in this paper. The role of molecular oxygen in the metabolism of organic carbon compounds is summarized, followed by a discuss~on of some of the pathways and the electron acceptors involved in the degradation under anaerobic conditions. Several degradation mechanisms are demonstrated for selected aliphatic and aromatic xenobiotic compounds. Examples are given of factors which will strongly influence a degradation of anthropogenic substances. Finally, existing methods for biodegradability testing, their limitations, and some suggestions for improvements are described.
1. INTRODUCTION The omnipresence of toxic organic chemicals in our environment and
their continuous build-up are a strong urge to understand the behavior and fate of these compounds in the different compartments of the environment. Specifically their susceptibility to degradation by microorganisms is of vital importance. In large parts of these compartments, e.g. deeper layers of soil, sediments and groundwater, mulecular oxygen is not available and the activity of aerobic microorganisms is minimal.
In these environments, but also in anaerobic treatment processes for various waste streams, mainly anaerobic microorganisms must bring about biodegradation of xenobiotic compounds. Although we have some knowledge of anaerobic biotransformations of anthropogenic compounds, additional research is needed to predict the ultimate fate of man-made organic compounds in anaerobic environments. Specifically whether a given compound is biodegradable under defined anaerobic conditions.
This article will focus on the potential and limitations of anaerobic biodegradation, and on methods to predict biodegradability.
2. ROLE OF OXYGEN Molecular oxygen has two different functions in the metabolism of
aerobic microorganisms which utilize organic substrates as carbon and energy source. In the oxidation of the substrate to intermediate products and in the biosynthesis of cellular components from these intermediate products, it serves in a limited number of reactions as a direct oxidant. It is also a terminal electron acceptor during respiration processes, in which reduced co-factors like NADH and FADH2 (formed during oxidation of the substrate) are reoxidized and thus regenerated. In its first function molecular oxygen is inserted directly into the substrate by highly specific enzymes (I), called oxygenases (monooxygenases and dioxygenases). In the absence of molecular oxygen, anaerobic microorganisms utilize alternative enzyme systems to oxidize the substrate. For instance, instead of using an oxygenase to obtain a hydroxyl group on an aromatic ring, they accomplish this by a combination of hydrogenation, dehydrogenation, and hydration reactions (2). Other ailaerobic reaction mechanisms are dehydration, O-demethylation, addition of water across a double bond in side chains, (de)carboxylation, hydrolysis and reductive
- 215-
dehalogenation. Dehalogenation has been observed to occur in aerobic environments, e.g. elimination of the halogen as the hydrogen halide with subsequent double-bond formation and replacement of the halide by a hydroxyl group catalyzed by either mixed function oxygenases or dioxygenases (3). In addition, several researchers in recent years have also reported dehalogenations from both alkyl (4) and aromatic (5) halides by replacement of the halide by hydrogen under anaerobic conditions. The second function of molecular oxygen, acting as a terminal electron acceptor, is fulfilled in its absence by a number of other compounds. Examples are the oxides of nitrogen and sulfur, carbon dioxide, oxidized metal ions, and some organic compounds. A disadvantage for' microorganisms to utilize these compounds as an electron acceptor lies in the fact that they have a lower redox potential compared with molecular oxygen. This will lead to a smaller energy yield during electron transport phosphory la tion ( 6 ) .
It can be concluded that during the metabolism of organic substrates by microorganisms only very few. reactions require molecular oxygen. Whether man-made organic compounds can be utilized by anaerobic microorganisms and which factors will influence this utilization will be discussed in the following sections.
3. BIODEGRADATION AND RECALCITRANCE Degradation of an organic compound is defined here as a sequence of
changes in the molecular structure which may ultimately lead to the formation of carbon dioxide, water, and various inorganic forms. Major and extensive changes in organic compounds in nature are mostly brought about via enzymatic reactions by microorganisms (7): biodegradation or biotransformation. These terms do not implie any extent of degradation. The ultimate biodegradation of an organic compound is called mineralization and results in the formation of inorganic compounds and microbial cell material. When the compound stays unchanged it is referred to as being recalcitrant; in between, many degrees of biodegradation are possible. In testing the biodegradability of an organic compound a more extensive definition of recalcitrance is required. In this paper we will use the definition of Giger and Roberts (8): a compound can be defined as being recalcitrant in a particular environment if it maintains its identity in that environment for more than an arbitrary length of time.
The recalcitrance of a compound may be attributable to two factors: (i) characteristics of that compound itself, or (ii) unsatisfactory conditions in the environment in question. Some molecules have a chemical structure, e.g. synthetic polymers like polyethylene, which makes them completely resistant to any form of microbial degradation. Environmental conditions which will directly influence biodegradation are of biological (e. g. presence of microorganisms capable of degradation), physical (temperature, water potential, accessibility of the compound), and chemical origin (presence of essential growth factors, suitable electron acceptor, pH, concentration of the compound).
If the structure of the organic chemical is responsible for its recalcitrance, it must be considered whether the manufacture of the chemical should proceed or be stopped. When optimal environmental conditions are found at which that compound is degraded, then it must be evaluated whether those conditions prevail or can be created in those environments where the chemical will eventually be present (soil, surface water, groundwater, water treatment system). If not, a further manufacture of the chemical should be a point of discussion. A strong aspect which will influence the outcome of a biodegradation test is the time
- 216-
period of testing. Biodegradation rates may be very low, especially under sub-optimal environmental conditions, and degradation may not be observed within the test period. Because recalcitrance can never really be proven (9), both a negative as well as a positive outcome of a biodegradation test must include the length of time of testing.
4. ANAEROBIC BIODEGRADATION The activity, diversity and abundancy of microorganisms in a given
habitat is largely determined by the presence of specific electron donors and electron acceptors. Organotrophic microorganisms use organic carbon as electron donor in anaerobic environments, while the oxides of nitrogen and sulfur, carbon dioxide but sometimes also the organic carbon function as electron acceptors. Dependent upon the predominant electron acceptor, we speak of fermentative, denitrifying, sulfate-reducing or methanogenic conditions (Table 1).
Table 1 Sequence of biologically mediated reductions
Process Aerobic metabolism Denitrification Fermentation Sulfate reduction Methanogenesis
Reduction reaction O2 -+ H20 N03 - -+ N2 CH20 -+ CHJOH S04 2- -+ HS CO2 -+ CH4
Redox potentiall (mV) + 810 + 750 - 180 - 220 - 250
When we consider a certain habitat, with sufficient nutrients and with molecular oxygen as the limiting factor, oxidation of the organic carbon by aerobic microorganisms will lead to depletion of the oxygen supply. Facultative microorganisms, capable of utilizing nitrate as electron acceptor, will subsequently become dominant and reaction mechanisms in which molecular oxygen plays a crucial role will disappear. From an energetic point of view, nitrate is the best electron acceptor following dioxygen (10). With a further decrease in the level of redox potential, sulfate and carbon dioxide may be reduced, together with the occurrence of fermentation reactions. Under these conditions strict anaerobic microorganisms, e.g. sulfate-reducing and methanogenic bacteria, will replace the facultative ones. Each microbial population has a different potential to degrade man-made organic compounds. The behavior of a specific compound may therefore vary from one environment to the other. This will be demonstrated with information about the biodegradation of aliphatic and aromatic compounds in different environements.
4.1. Aliphatic compounds Strong evidence for the anaerobic biodegradation of aliphatic hydro
carbons is restricted to unsaturated and halogenated ones. Some literature exists (11,12,13) in which saturated hydrocarbons are demonstrated to be degraded in the absence of oxygen but conclusive evidence is not available. Specifically the hydroxylation reaction of alkanes has not yet been demonstrated in anaerobic microorganisms (14). Once a terminal or subterminal alcohol is formed, anaerobic oxidation is feasible since dioxygen is not involved in further degradation steps.
- 217-
Because of the double bond in unsaturated aliphatic compounds, hydration may be a suitable mechanism to form an aliphatic alcohol. This has been postulated as the first step in the stoichiometric conversion of 1-hexadecene to methane and carbon dioxide by methanogenic enrichment cultures (14). We must keep in mind that, although man has created environmental pollution by release of saturated and unsaturated aliphatics in nature, they are an integral part of our environment. They have been around for millions of years and during that time microorganisms have been exposed to them. We therefore may assume that anaerobic enzyme systems exist to degrade these compounds. Other factors, like unfavorable environmental conditions, may be the reason for observed recalcitrance.
Halogenated aliphatic hydrocarbons, which are in general man-made, are of great concern. Most of the research on the biodegradation of these compounds has been with C1 and Cz hydrocarbons, which are found in many surface- and groundwaters. An overview of this research has been given by Schraa and Zehnder (15) and by Vogel et al. (16). Trichloromethane (chloroform) and tetrachloroethene (PER) will be used as examples from several biodegradation studies. In one study, the degradation of trichloromethane (60 ~g/l) was followed for eleven weeks in a batch experiment, in the presence of nitrate as electron acceptor, ethanol as primary organic substrate, and an active denitrifying bacterial culture. During this incubation no degradation took place (17). However, under methanogenic conditions biodegradation of trichloromethane (15-40 ~g/l) was observed in both batch and continuous-flow column experiments. The degradation, demonstrated by its removal and by 14COZ production from labeled trichloromethane, took place in the presence of acetate as primary organic substrate and after an acclimation period of 3 to 10 weeks (18). The initial mechanism in the transformation was thought to be the replacement of a chloride with a hydrogen atom (reductive dechlorination). A contradictory result under methanogenic conditions was reported by Shelton and Tiedje (19). In experiments, especially designed to obtain a general testing method for determining the anaerobic biodegradation of organic chemicals, no biodegradation of trichloromethane was observed at a concentration of 50 mg/l and in a time period of 8 weeks. These three results demonstrate the complexity of drawing conclusions about the biodegradation of a specific compound under anaerobic conditions.
Tetrachloroethene was found to be degraded in the same exper iments in which Bouwer and McCarty (18) observed the biodegradation of trichloromethane under methanogenic conditions. Degradation occurred after an acclimation period of up to 10 weeks and acetate was present as primary organic substrate. Because of the presence of traces of trichloroethene, reductive dechlorination was thought to be the first step in the biodegradation. This mechanism was also postulated by Parsons and Lage (ZO), who observed the formation of trichloroethene, and cis- and trans-1,Z-dichloroethene from tetrachloroethene in a reductive environment simulating underground conditions. Although the degradation was mainly biological, chemical dechlorination could also be observed. This is in agreement with other findings about abiotic reductive dechlorination of aliphatic hydrocarbons in anaerobic environments (16, Zl).
Vogel and McCarty (4) demonstrated the conversion of tetrachloroethene to trichloroethene and vinyl chloride under methanogenic conditions. In addition, 14C-labelled tetrachloroethene was at least partially mineralized to COZ. Biodegradation occurred in the presence of acetate and no mention was made of a required acclimation period. The proposed pathway is shown in Fig. 1.
The exact mechanism of mineralization remains unclear. A possibility
- 218-
would be the hydroxylation of the double bond, resulting in 2-chloroethanol. Additional oxidation steps without the involvement of molecular oxygen can eventually lead to CO2.
4.2 Aromatic compounds The studies on the anaerobic biodegradation of (halogenated) aro
matic compounds, as reviewed by Young (22), Berry et a1. (23), Tiedje et a1. (24) and Holliger et a1. (25), demonstrate that, although the information for some classes of aromatic compounds is still scarce, anaerobic
Fig. 1. Proposed pathway for the biodegradation of tetrachloroethene under methanogenic conditions. A: tetrachloroethene, B: trichloroethene, C: dichloroethene, D: vinyl chloride (4).
microorganisms are capable of metabolizing many man-made aromatics. These reviews show that most of the research has been focused on homocyclic aromatic compounds (23,25), specifically (chlorinated) benzene derivatives. Little information exists on the anaerobic biodegradation of heterocyclic and polycyclic aromatic compounds.
As has been mentioned before, aerobic microorganisms metabolize aromatic compounds via oxygenase enzymes for which molecular oxygen is required. This oxygen is inserted in the compound during hydroxylation and during ring cleavage. Anaerobic microorganisms use different enzyme catalyzed reactions. Examples will be given by describing the biodegra-
- 219-
dation of some selected compounds. Among the first aromatic compounds which were found to be degraded under anaerobic conditions were the oxygen-substituted ones like benzoate and phenol. Biodegradation has been observed in the presence of nitrate (26,27), sulfate (28,29), and under methanogenic conditions (30,31). Although the exact pathways are not all proven (different ones are proposed), degradation starts with a reduction of the ring, followed by a reductive cleavage of the ring to aliphatic acids. In between a number of hydroxylation and dehydrogenation reactions, and with benzoate a decarboxylation, may take place. This is shown in Fig. 2 for a pure culture of a Moraxella species (27).
~o (
d 1
4[H] \: •
~o (
d 2
•
Fig. 2. Proposed degradation pathway for the degradation of benzoate by Moraxella species under nitrate reducing conditions. 1: benzoate, 2: cyclohexene-1-carboxylic acid, 3: 2-hydroxycyclohexane carboxylic acid, 4: 2-oxocyclohexane carboxylic acid, 5: cyclohexanone, 6: adipic acid. Adapted from Williams and Evans (27).
It has also been demonstrated that the degradation may start with an oxidation of a methyl substituent. Bossert et a1. (32) report of a syntrophic degradation of p-cresol by two bacterial species in coculture under nitrate reducing conditions. One species metabolized p-cresol via dehydrogenation and hydration reactions to p-hydroxybenzoate, which was further metabolized by the second species (33).
Aromatic compounds without oxygen in their molecular structure, like benzene, toluene and xylene, were for a long time believed to be resistant to biodegradation without the presence of molecular oxygen. The lack of functional groups like hydroxyl or carboxyl groups would prevent hydration of double bonds. Recently, several researchers have reported fascinating results. Zeyer et al. (34) observed mineralization of toluene
- 220-
and m-xylene under denitrifying conditions in a laboratory aquifer column system. In addition, toluene, but also benzene, was found to be degraded by mixed methanogenic cultures (35). Although degradation pathways have not yet been elucidated, Grbi~Galit and Vogel (35) concluded that, based on the detected compounds, ring oxidations, ring reductions, and methyl group oxidations had occurred. Part of the benzene and toluene was degraded to CH4 and CO2, The mixed methanogenic cultures had been exposed to the tested compounds for a long period of time (over 9 months). In a study by Schink (14) however, no degradation of benzene and toluene was detected under methanogenic conditions in a test period of 14 weeks. Toxic effects of the substrates or insufficient acclimation time are possible reasons for the observed recalcitrance.
Halogenated aromatic compounds require additional degradation mechanisms for the removal of the halogen atom. Most research has been focused on chlorophenols, chlorobenzoates and chlorobenzenes (24). Reductive dechlorination was found to be the major mechanism of chlorine removal (36,37,38). This has especially been examined under methanogenic conditions. Little is known of reductive dechlorination in the presence of nitrate, while the results under conditions favoring sulfate reduction are confusing. Biodegradation of trichlorobenzenes has been found in the presence of sulfate (Bosma ~! ~l., in preparation), while other researchers have observed inhibition of the degradation of chlorophenols under similar conditions (37). The results from Shelton and Tiedje (36) show that in the degradation of 3-chlorobenzoate under methanogenic conditions a consortium of at least four microorganisms is involved, each having its specific task. The first biodegradation step is the dechlorination of 3-chlorobenzoate to benzoate by one specific microorganism. In addition, benzoate is converted to CH 4 and CO 2, It is not known yet whether this dechlorination is a specific catabolic reaction or that the chlorinated compound acts as a terminal electron acceptor. The latter means that it may have to "compete" with nitrate, sulfate or carbon dioxide for electrons.
5. TESTING BIODEGRADABILITY The results of a biodegradation study may be useful but also
limited. When a compound is found to be degraded in a system under specific conditions and the rate and extent of degradation are known, then we may be able to predict its eventual fate in our environment. The observation that degradation does not occur, does not yet mean that the compound is not biodegradable; we have only demonstrated that under a given set of conditions the microorganisms failed to bring about degradation. Therefore, characterizing a compound as being recalcitrant or biodegradable has to be combined with as many data as possible about the test conditions.
There are two major conditions which have to be met to test the susceptibility of an anthropogenic organic compound to biodegradation (9). First, microbial populations have to be present which have the genetic capability to synthesize enzymes involved in the various transformation steps. Second, environmental conditions have to be adequate for the desired reactions to proceed at a significant rate.
By adding an inoculum of microorganisms, which have never had previous exposure to the tested compound or to structural similar ones, biodegradation may not occur within the period of te.sting or at all. This makes the choice of the origin of the inoculum a very important one. Tests as developed by OECD (39) in which the aerobic biodegradation of organic compounds is followed for a period of 28 days, using sewage
- 221-
microorganisms without prior exposure to the compound, are therefore not very appropriate. Compounds which are degraded in such tests should not be of environmental concern. In many experiments in which the biodegradation of xenobiotic compounds was studied, long lag periods were required before degradation would occur. From own experience is known that the aerobic biodegradation of l,2-dichlorobenzene in a soil percolation column would not start until after a lag period of about 3 months (40). It took 2 months before l,4-dichlorobenzene was degraded in an aerobic enrichment culture (41). Eventually, a bacterium was isolated capable of mineralizing 1,4-dichlorobenzene. Under anaerobic conditions, lag periods preceding dechlorination of chlorobenzoates and chlorobenzenes were even longer, 3-6 months (42; Bosma et al., in preparation). Such a long lag period puts a serious constraint on the testmethod. An other aspect which has not been mentioned before but which may be important in the test is whether the compound is used as carbon and/or energy source or that the degradation proceeds via co-metabolism. Dechlorination for one may only supply energy.' The responsible microorganism in the degradation of 3-chlorobenzoate was found te be dependent on other microorganisms for its carbon supply (36). When a compound is tested whether it can be degraded by a microbial population via co-metabolism, an additional carbon and energy source has to be added to the test system. The degree of biodegradation is also of great importance. Just following the disappearance of the parent compound is not a realistic approach. Intermediates can be formed which may be toxic and/or less degradable. Oxygen uptake, carbon dioxide production or dissolved organic carbon removal are very limited parameters to measure biodegradation. They are feasible when high concentrations and relatively short test periods are used. However, because of the long test periods in anaerobic testing and the toxicity of many compounds, which requires concentrations in the test system in the ~g-mg/l region, other methods have to be used. Determination of the disappearance of the parent compound should be combined with the detection of intermediates, the formation of halogens in case of halogenated compounds, and the use of radiolabeled compounds (43).
The environmental conditions in the test system will also be a major factor in the biodegradation of a compound. The presence of resp. nitrate, sulfate or carbon dioxide as external electron acceptors and the respective reducing conditions will, in combination with the tested compound, determine which microbial populations will dominate in the system. The differences in biodegradation for a number of aromatic compounds under sulfate reducing and methanogenic conditions as reported by Gibson and Sulflita (37) and Boyd and Shelton (43), are a strong urge to perform biodegradability testing under different reducing conditions. An other factor which may be decisive in the outcome of the test is the concentration of the compound. Toxicity effects at higher concentrations will prevent biodegradation, while at lower concentrations the biodegradation rates may be dependent on the concentration (44). In addition, below a certain concentration (threshold value) biodegradation may not occur at all. The concentration of the compound is too low to support microbial growth (44,45). Here too, different concentrations should be tested.
6. TEST METHODS In contrast to the large number of aerobic biodegradation tests (39)
only few tests have been described in which the biodegradation of a compound is determined in the absence of oxygen and under specific anae-
- 222-
robic conditions. Owen et al. (46) were the first who developed a simple and inexpen
sive method to test the biodegradability under methanogenic conditions. Anaerobic serum bottles containing the test compounds, defined media and seed inocula were incubated at a given temperature. Degradation and also toxicity effects 'were measured by monitoring the cumulative gas production volumetrically with the syringe method of Nottingham and Hungate (47). Healy and Young (48) used this technique to test the biodegradability of several aromatic compounds. Benzoate and phenol (at 300 mg/l) were two of the compounds shown to be degraded after a lag period of 1 to 2 weeks by an unacclimated inoculum from a laboratory anaerobic digester fed primary settled sewage sludge. The technique was refined by Shelton and Tiedje (19). They proposed to standardize the test by utilizing digested sewage sludge diluted to 10%, adding 50 mg of C per liter of the test compound, using a standard anaerobic medium, and incubating at 350 C. Biodegradation was tested by the net increase in gas pressure in the bottles with the test compounds over the pressure in bottles without the test compounds. Gas production was measured by gas chromatography and by a pressure transducer. A strong drawback of this method is the high cancentration of the compound that has to be used to obtain significant gas production. As mentioned before this may cause toxicity problems. In general the method is invalid for (i) compounds which are tested at concentrations which are too low to observe gas production, (ii) compounds which are only partially degraded (e.g. dechlorination), (iii) compounds which are only degraded after a long acclimation period, and (iv) compounds which undergo very slow degradation. It remains to be seen whether this method can be used when nitrate or sulfate act as electron acceptor, and carbon dioxide plus nitrogen or hydrogen sulfide have to be measured. The same limitations as under methanogenic conditions exist.
Without developing a new anaerobic test method, I like to summarize some of the strategies and techniques which are used by a number of researchers to test the degradation of xenobiotic compounds. A first strategy is to utilize microorganisms which have previously been exposed to the test compound. This may shorten the expected lag time significantly and will increase the chance to observe degradation. A second strategy is to acclimate the microorganisms to increasing concentrations of the test compound (49,50). This procedure will avoid toxicity effects. Initial concentrations can be based on a previously performed toxicity assay (49). A third strategy is to add the test compound in combination with a number of other carbon sources. By giving several substrates both a large population as well as a certain diversity in the population will be obtained. Also, by adding selected compounds (e.g. non-chlorinated ones when a chlorinated compound is tested) processes like co-metabolism may be stimulated. At the same time these substrates may act as electron donor in reductive dechlorination reactions. On the other hand, as long as alternative carbon and energy sources are present, biodegradation of the test compound may be suppressed. This point requires further research.
Techniques which are generally used to test the biodegradation of organic compounds can be devided in batch and (semi)-continuous flow systems. The first one can be compared with the method of Shelton and Tiedje (19). However, the emphasis with this technique lies not on the measurement of gas production, but on the disappearance of the parent compound and the formation of intermediates and/or products (e.g. 20,35, 38). Analysis techniques which are required are las/liquid chromatography followed by mass spectrometry and the use of 1 C-labeled compounds. The
- 223-
principle of the second one is that after introduction of microorganisms in the system, they are retained in that system and are constantly exposed to an optimal environment by (semi)-continuous flow of the medium. Examples are fixed-film columns with microorganisms attached to glass beads (16,17,18) and percolation columns filled with sediment material (40,51). A large advantage of this system is the flexibility that the researcher has in operating it. Toxicity problems may be overcome with gradual increases of the concentration of the compound while the composition of the media can also easily be changed.
7. CONCLUDING REMARKS A large number of xenobiotic compounds has been found to be degraded
by microorganisms under anaerobic conditions. Many microorganisms have developed enzyme systems which do not require the presence of molecular oxygen. The biodegradation of some compounds, e.g. tetrachloroethene, even seems to be restricted to the absence of molecular oxygen and requires reducing conditions.
The outcome of a biodegradability test has to be judged with caution. Persistence of the compound during the test period may be caused by its chemical structure but also by the created environmental conditions. Disappearance of the compound may be brought about by biodegradation but additional analyses are required to know the extent of degradation. Whether the outcome of a biodegradability test can also be found in specific parts of our environment remains to be seen.
ACKNOWLEDGEMENTS The author is grateful to Tom Bosma for glvlng access to his data
prior to publication, Ans Broersma for typing the manuscript and Nees Slotboom for drawing the figures.
REFERENCES
(1) Dagley, S. (1975). A biochemical approach to some problems of environmental pollution. In: Rater Pollution Microbiology, P.N. Campbell and W.M. Aldridge (Eds.), Vol II, Academic Press, London, pp. 81-138.
(2) Evans, W.C. (1977). Biochemistry of the bacterial catabolism of aromatic compounds in anaerobic environments. Nature 270, 17-22.
(3) Wood, J.M. (1982). Chlorinated hydrocarbons: oxidation in the biosphere. Environ. Sci. Technol. 16, 291A-297A.
(4) Vogel, T.M. and McCarty, P.L. (1985). Biotransformation of tetrachloroethylene to trichloroethylene, dichloroethylene, vinyl chloride and carbon dioxide under methanogenic conditions. Appl. Environ. Microbiol. 49, 1080-1083.
(5) Tsuchiya, T. and Yamaha, T. (1984). Reductive dechlorination of l,2,4-trichlorobenzene by Staphylococcus epidermis isolated from intestinal contents of rats. Agric. BioI. Chem. 48, 1545-1550.
(6) Zehnder, A.J.B. (1982). The carbon cycle. In: Handbook of Environmental Che~istry, O. Hutzinger (Ed.), vol. I, Part B, SpringerVerlag, Berlin, pp. 83-110.
(7) Alexander, M. (1981). Biodegradation of chemicals of environmental concern. Science 211, 132-138.
(8) Giger, W. and Roberts, P.V. (1978). Characterization of persistent organic carbon. In: Rater Pollution Microbiology, R. Mitchell. (Ed.), Vol. 2, John Wiley & Sons, New York, pp. 135-175.
- 224-
(9) Grady Jr., C.P.L. (1985). Biodegradation: its measurement and microbiological basis. Biotechnol. and Bioeng. 27, 660-674.
(10) Zehnder, A.J.B. and Stumm, W. Geochemistry and biogeochemistry of anaerobic habi ta t s. In: Environmental Microbiology of Anaerobes, A.J.B. Zehnder (Ed.), John Wiley & Sons, New York, in press.
(11) Davis, J.B. and Yarbrough, H.F. (1966). Anaerobic oxidation ·of hydrocarbons by Desulfovibrio desulfuricans. Chem. Geol. I, 137-144.
(12) Atlas, R.M. (1981). Microbial degradation of petroleum hydrocarbons: an environmental perspective. Microbiol. Rev. 45, 180-209.
(13) Giger, W., Schaffner, C. and Wakeham, S.G. (1980). Aliphatic and olefinic hydrocarbons in recent sediments of Greifensee, Switzerland. Geochim. Cosmochim. Acta. 44, 119-129.
(14) Schink, B. (1985). Degradation of unsaturated hydrocarbons by methanogenic enrichment cultures. FEMS Microbiol. Ecol. 31, 69-77.
(15) Schraa, G. and Zehnder, A.J.B. (1986). Biodegradation of chlorinated compounds. In: Organic Micropollutants in the Aquatic Environment, A. Bjcjrseth and G. Angeletti (Eds.), D. Reidel Publishing Company, Dordrecht, Holland, pp. 278-291.
(16) Vogel, T.M., Criddle, C.S. and McCarty~ P.L. (1987). Transformations of halogenated aliphatic compounds. Environ. Sci. technol. 21,722-737.
(17) Bouwer, E.J. and McCarty, P.L. (1983). Transformations of halogenated organic compounds under denitrification conditions. Appl. Environ. Microbiol. 45, 1295-1299.
(18) Bouwer, E.J. and McCarty, P.L. (1983). Transformations of 1- and 2-carbon halogenated aliphatic compounds under methanogenic conditions. Appl. Environ. Microbiol. 45, 1286-1294.
(19) Shelton, D.R. and Tiedje, J.M. (1984). General method for determining anaerobic biodegradation potential. Appl. Environ. MicrobioI. 47, 850-857.
(20) Parsons, F. and Lage, G.B. (1985). Chlorinated organics in simulated groundwater environments. J. Am. Water Works Assoc. 77, 52-59.
(21) Macalady, D.L., Tratuyek, P.G. and Grundl, T.J. (1986). Abiotic reduction reactions of anthropogenic organic chemicals in anaerobic systems: a critical review. J. Contam. Hydrol. I, 1-23.
(22) Young, L.Y. (1984). Anaerobic degradation of aromatic compounds. In: Microbial Degradation of Organic Compounds, D.T. Gibson (Ed.), Marcel Dekker, Inc., New York, pp. 487-523.
(23) Berry, D.F., Francis, A.J. and Bollag, J.- M. (1987). Microbial metabolism of homocyclic and heterocyclic aromatic compounds under anaerobic conditions. Microbiol. Rev. 51, 43-59.
(24) Tiedje, J.M., Boyd, S.A. and Fathepure, B.Z. (1987). 13. Anaerobic degradation of chlorinated aromatic hydrocarbons. Dev. Ind. MicrobioI. 27, 117-127.
(25) Holliger, C., Stams, A.J.M. and Zehnder A.J.B. Anaerobic degradation of recalcitrant compounds (Presented at 5th Int. Symp. on anaerobic Digestion, Bologna, Italy, May 22-28, 1988).
(26) Bakker, G. (1977). Anaerobic degradation of aromatic compounds ·in the presence of nitrate. FEMS Microbiol. Lett. I, 103-108.
(27) Williams, R.J. and Evans, W.S. (1975). The metabolism of benzoate by Moraxella species through anaerobic nitrate respiration. Biochem J. 148, 1-10.
(28) Bak, F. and Widdel, F. (1986). Anaerobic degradation of phenol and phenol derivatives by Desulfobacterium phenolicum sp. nov. Arch.
- 225-
Microbiol. 146, 177-180. (29) Szewzyk, R. and Pfennig, N. (1987). Complete oxidation of catechol
by the strictly anaerobic sulfate-reducing Desulfobacterium catecholicum sp. nov. Arch. Microbiol. 147, 163-168.
(30) Horowitz, A., Shelton, D.R., Cornell, C.P. and Tiedje, J.M. (1982). Anaerobic degradation of aromatic compounds in sediments and digested sludge. Dev. Ind. Microbiol. 23, 435-444.
(31) Young, L.Y. and Rivera, M.D. (1985). Methanogenic degradation of four phenolic compounds. Water Res. 19, 1325-1332.
(32) Bossert, 1.D., Rivera, M.D. and Young, L.Y. (1986). ~-Cresol biodegradation under denitrifying conditions: isolation of a bacterial coculture. FEMS Microbiol. Ecol. 38, 313-319.
(33) Bossert, 1.D. and Young, L.Y. (1986). Anaerobic oxidation of ~cresol by a denitrifying bacterium. Appl. Environ. Microbiol. 52, 1117-1122.
(34) Zeyer, J., Kuhn, E.P. and Schwarzenbach, R.P. (1986). Rapid microbial mineralization of toluene and 1,3-dimethylbenzene in the absence of molecular oxygen. Appl. Environ. Microbiol. 52, 944-947.
(35) Grbit-GalH:, D. and Vogel, T.M. (1987). Transformation of toluene and benzene by mixed methanogenic cultures. Appl. Environ. MicrobioI. 53, 254-260.
(36) Shelton, D.R. and Tiedje, J.M. (1984). Isolation and partial characterization of bacteria in an anaerobic consortium that mineralizes 3-chlorobenzoic acid. Appl. Environ. Microbiol. 48, 840-848.
(37) Gibson, S.A. and Suflita, J.M. (1986). Extrapolation of biodegradation results to groundwater aquifers: reductive deha1ogenation of aromatic compounds. App1. Environ. Microbiol. 52, 681-688.
(38) Boyd, S.A., Shelton, D.R., Berry, D. and Tiedje, J.M. (1983). Anaerobic biodegradation of phenolic compounds in digested sludge. Appl. Environ. Microbiol. 46, 50-54.
(39) OECD Guidelines for Testing of Chemicals. (1981). Organisation for Economic Co-operation and Development, Paris, France.
(40) Meer, J.R. v.d., Roelofsen, W., Schraa, G. and Zehnder, A.J.B. Degradation of low concentrations of dichlorobenzenes and 1,2,4-trichlorobenzene by Pseudomonas sp. strain PSI in nonsterile soil columns. FEMS-Microbial Ecol., in press.
(41) Schraa, G., Boone, M.L., Jetten, M.S.M., van Neerven, A.R.W., Colberg, P.J. and Zehnder, A.J.B. (1986). Degradation of 1,4-dichlorobenzene by Alcaligenes sp. strain A175. Appl. Environ. MicrobioI. 52, 1374-1381-
(42) Horowitz, A., Suflita, J.M. and Tiedje, J.M. (1983). Reductive dehalogenations of halobenzoates by anaerobic lake sediment microorganisms. Appl. Environ. Microbiol. 45, 1459-1465.
(43) Boyd, S.A. and Shelton, D.R. (1984). Anaerobic biodegradation of chlorophenols in fresh and acclimated sludge. Appl. Environ. MicrobioI. 47, 272-277.
(44) Alexander, M. (1985). Biodegradation of organic chemicals. Environ. Sci. Technol. 18, 106-11l.
(45) McCarty, P.L., Reinhard, M., and Rittmann, B.E. (1981). Trace organics in groundwater. Environ. Sci. Technol. 15, 40-51.
(46) Owen, W.F., Stuckey, D.C., Healy Jr., J.B., Young, L.Y. and McCarty, P.L. (1979). Bioassay for monitoring biochemical methane potential and anaerobic toxicity. Water Res. 13, 485-492.
(47) Nottingham, P.M. and Hungate, R.E. (1969). Methanogenic fermentation of benzoate. J. Bact. 98, 1170-1172.
- 226-
(48) Healy, Jr., J.B. and Young, L.Y. (1979). Anaerobic biodegradation of eleven aromatic compounds to methane. Appl. Environ. Microbiol. 38, 84-89.
(49) Guthrie, M.A., Kirsch, E.J., Wukasch, R.F., and Grady Jr. C.P.L. (1984). Pentachlorophenol biodegradation-II Anaerobic. Water Res. 4, 45l-46l.
(50) Blum, D.J.W., Hergenroeder, R., Parkin, G.F., and Speece, R.E. (1986). Anaerobic treatment of coal conversion wastewater constituents: biodegradability and toxicity. J. Water Pollut. Control Fed. 58, 122-13l.
(51) Kuhn, E.P., Colberg, P.J., Schnoor, J.L., Wanner, 0., Zehnder, A.J.B., and Schwarzenbach, R.P. (1985). Microbial transformation of substituted benzenes during infiltration of river water to groundwater: Laboratory column studies. Environ. Sci. Technol. 19, 961-968.
- 227-
THE FATE OF ORGANIC COMPOUNDS IN THE ENVIRONMENT
A.H. Neilson, A.-S. Allard, C. Lindgren and M. Remberger
Swedish Environmental Research Institute
Summary
An account is given of microbiological procedures for investigating the fate of xenobiotics discharged into the aquatic environment. A brief account is given of the experimental procedures which have been developed, and their application is exemplified from the results of experiments with chloroguaiacols and related compounds. Experiments encompass both aerobic and anaerobic transformations, and emphasis is placed on environmental factors of cardinal significance in determining the outcome of laboratory experiments. An attempt is made to assess the environmental relevance of the data acquired by carrying out experiments in which natural sediment samples were spiked with the xenobiotics. Attention is directed to the significant role of binding to sediments.
INTRODUCTION
Studies on degradation including both biotic and abiotic processes occupy a central position in the environmental hazard assessment of organic chemicals. We are concerned here solely with biotic reactions, and particularly with problems associated with compounds which may be termed recalcitrant (1). In a review (8), we distinguished between biodegradation and biotransformation for pragmatic reasons.
The result of biodegradation and biotransformation depends both on the organisms carrying out the reactions, and on the substrate. For reactions in the aquatic phase, bacteria will generally be the most significant agents, and we shall restrict this discussion entirely to them. The fate of a given compound is a multicomponent function of many variables including (i) the physical environment - temperature, pH and p02 (ii) the source, nature, and number of microorganisms and (iii) the growth status of the cells.
Metabolically conclusive experiments are difficult to carry out under field conditions : we have therefore chosen to carry out experiments under laboratory conditions simulating as closely as possible natural systems. Such experiments have the advantage of both reproducibility and flexibility. An outline of the methodology which we have evolved is given in Fig. 1. The structures of the compounds encountered in these investigations are shown in Fig.2.
- 228-
STAGE I ANALYSIS OF PROBLEM
Environmental sample (sediment, water)
~ Chemical analysis of xenobiotic
Microbiological examination
I Enrichment
I Isolation of micro-organism
I Synthesis and identification of metabolites ~
Metabolic studies
Ecotoxicological studies (toxicity, bioconcentration)
STAGE 1\ VERIFICATION OF DATA
Examination of substrates and metabolites for:
• Existence in biota and sediments
• Binding to sediments and release
• Uptake and metabolism in biota
• Microbial transformation in sediment and water
• Non- microbial transport processes
Figure 1
EXPERIMENTAL PROCEDURES
Analytical considerations. Our research programme has depended critically on the availability of analytical expertise. Before beginning any experiment, methods were developed so that a large number of samples could rapidly and accurately be analysed during the course of the experiments. Such kinetic experiments reveal the dynamics of the reactions being investigated, and at the same time, increase the numerical accuracy due to the availability of such a substantial data base. Details of all of the procedures have been given in our publications, but we wish to emphasize two aspects : (i) quantification requires access to pure reference material - not only for the substrates - but also for their metabolites,and this may necessarily involve a substantial
- 229-
synthetic activity (ii) we have based identification of metabolites on either or both of the following (a) identity to the mass spectrum of an authentic reference compound together with the GC retention time of a suitable derivative (b) identity of the GC retention times of three structurally independent derivatives with those of authentic compounds.
Microbiological considerations. In all of our studies we have used the enrichment methodology in view of the cardinal significance of this choice, we summarize briefly the reasons for its choice (i) the method provides valuable insight into metabolic pathways used for degradation of the xenobiotic, including the formation of transient, and possibly toxic metabolites (ii) it provides an incisive procedure for analysis of reactions which will occur under natural conditions. Evidence supporting the view that natural populations of bacteria exposed to a xenobiotic are indeed enriched in those individuals with the capacity for its degradation may be deduced from extensive studies both in the terrestrial environment ( 10) and in the aquatic (5).Slightly different procedures have been used for aerobic and anaerobic enrichments due to the generally slower growth, and greater degree of fastidiousness of anaerobic bacteria.
OH OH
Cr°H C100H CATECHOL CI I ~ CI CI ~ CI
CI CI 3,4,5- TETRA-
OH OMe OH
CroMe Cr°H C100Me
CI I ~ CI CI I ~ CI GUAIACOL
CI ~ CI
CI CI CI 3,4,5- 4,5,6- TETRA-
OMe OMe
CroMe C100Me VERATROLE CI ~ CI CI ~ CI
CI CI 3,4,5- TETRA-
OH
MeOOOMe SYRINGOL CI ~ CI
CI 3,4,5-
Figure 2
- 230-
AEROBIC REACTIONS
As a model reaction, we chose the a-methylation of halogensubstituted phenolic compounds. This choice was based on a number of considerations (i) the widespread use and production of such compounds (ii) the products were formed by a single-step reaction and were lipophilic, rather than more polar than their precursors. In order to avoid misunderstanding, however, it should be emphasised that we regard this reaction as an alternative to biodegradation since total mineralization of both chlorophenols (refs. in 3) and chloroguaiacols (6) has been demonstrated. The organisms carrying out this reaction are clearly widespread in the aquatic environment and we have isolated bacteria belonging both to the Gram positive taxon Rhodococcus and to the Gram negative genera Acinetobacter and Pseudomonas.
We have investigated three aspects of environmental significance : (i) the effect of substrate concentration and cell density (ii) the role of concurrent metabolism (8) and (iii) the rates of the reactions. Only a rather brief summary of the salient conclusions can be presented here.
Studies on the effect of substrate concentration revealed two significant facts : (i) there was not a linear relation between the rate of a-methylation of 4,5, 6-trichloroguaiacol and the substrate concentration (2) and at low substrate concentrations ( (100 ug. L-1), the reac~ion was essentially quantitative, particularly for Gram positivE:. organisms (Table 1) . We would therefore predict that under environmental conditions where low substrate conditions prevail, this would be a quantitatively significant reaction. (ii) the nature of the metabolites could also be
Table 1. Final yields (%) of 3,4,5-trichloroveratrole formed by groups of bacteria from high ( 20 mg.L-1), and low ( 100 ug.L-1) concentrations of 4,5,6-trichloroguaiacol.
Gram positive strains
1395 1539 1571 1624 1632
Gram negative strains
1556 1557 1558 1559 1631 1637 1678
- 231-
Concentration
High
45 87
100 1.0
76
0.3 0.7 0.4 1.8 1.0 1.3 6.0
Low
100 100 100
18 100
9 30 21 56 47 71
100
affected. A striking example is provided by the investigation of 3,4,5-trichloroguaiacol at concentrations of 10.0, 1.0, and 0.1 mg. L-1 : at the lowest substrate concentration only one metabolite was formed, though with increasing concentration the metabolic pattern became successively more complex. Both of these results have a significant bearing on the design of experiments on biodegradation, quite apart from their intrinsic environmental consequence.
We have also examined the effect of cell density on the rates of O-methylation (2). Briefly, the results showed that at cell densities realistically expected in receiving waters, specific rates of O-methylation ( i.e. normalized to take into account the cell density) were not significantly lower than those at the high cell densities used in our laboratory experiments.
In natural situations, bacteria will almost never be exposed only to a xenobiotic : natural substances will be present, some of which may serve as suitable substrates for growth. We attempted to assess the significance of the structure of the growth substrate on the O-methylation reaction. Three distinct responses could be distinguished : (i) O-methylation was indifferent to the nature of the substrate and took place effectively during growth of a test strain with growth substrates as structurally diverse as betaine, succinate, gluconate and 4-hydroxybenzoate (ii) during 0-methylation using dense cell suspensions, an enhancement of the rate was observed with 3,4, 5-trichlorosyringol (7) (iii) an interesting situation arose during growth with vanillate which was structurally related to the cosubstrate (4,5,6-trichloroguaiacol). A complex sequence of reactions took place, involving de-O-methylation and successive O-methylation of 3,4, 5-trichlorocatechol to 3,4,5-trichloroguaiacol and 3,4,5-trichloroveratrole (2) •
It is therefore difficult to make generalizations about the effect of growth substrates on the O-methylation reaction although studies with cell-free extracts supported the view that 0-methylating activity was constitutive. This would suggest that previous exposure to a xenobiotic is not necessary to elicit this activity, but in a natural system where intermediates may be channelled into other catabolic pathways or into higher biota, the effects may be highly significant ecologically.
There has been an increasing interest in the rates of biodegradation / biotransformation of xenobiotics, and we took advantage of the O-methylation reaction to examine the response of two test strains - one Gram positive, the other Gram negative - to a range of halogenated phenols. The reaction was zero-order with respect to the concentration of the product so that rates were readily estimated. From the cell density of the suspension, the specific rates ( i.e. rates normalized to the cell density) were calculated. The most important conclusions from this study (3) were that: (i) the Gram negative strain was relatively insensitive to the structure of the substrate (ii) the Gram positive strain was particularly discriminate towards the substitution pattern in, for example, trichlorophenols (iii) some substrates including 2,4-dinitrophenol were not O-methylated and were apparently degraded without the formation of detectable intermediates while others such as hexachlorophene were totally unreactive. These results support our contention that, for substituted phenolic substances,
- 232-
O-methylation may be a significant alternative to biodegradation. In addition, the reaction provides a rational basis for the finding of halogenated ani soles in both biota and sediment samples ( refs. in 3) and even in atmospheric samples from remote localities (4) whose detection is consistent with the high volatility of halogenated anisoles.
ANAEROBIC REACTIONS
In th~ light of the previous investigations, it was natural to direct attention to transformations of chloroveratroles, chloroguaiacols and chlorocatechols under anaerobic conditions. This was motivated additionally by the recovery and identification of a range of chlorinated guaiacols and chlorocatechols from areas of the Baltic Sea and the Gulf of Bothnia (11,12) subjected to the discharge of bleachery effluents (Table 2). In order to make
Table 2. Concentrations ( ug per kg organic C ) of "free" and "bound" chloroguaiacols and chlorocatechols in sediment samples.
Locality
A B C
Free Bound Free Bound Free Bound
3, 4, 5-trichloroguaiacol 210 2400 <10 1870 <10 7170 tetrachloroguaiacol 180 1590 <10 550 <10 2840
3, 4, 5-trichlorocatechol 270 3700 <10 1050 14700 25800 tetrachlorocatechol <10 8250 <10 1250 7300 63200
the experiments as ecologically realistic as possible, we carried out all of the experiments at an initial substrate concentration of 100 ug.L-1. Two significant alterations were made in the procedures used for aerobic organisms (i) for experimental accessibility, all of the experiments were carried out using the concurrent metabolism methodology (8) (ii) it has not been generally possible to obtain pure cultures of the relevant organisms : advantage was therefore taken of metabolically stable consortia.
Experiments were addressed to two questions : (i) de-Omethylation reactions and (ii) dechlorination reactions. We examined consortia grown with substrates such as 3,4,5-trimethoxybenzoate, 3,4,5-trihydroxybenzoate, and 5-chlorovanillin. As illustration of the results obtained, data from the first will be briefly summarized (9) : (i) de-O-methylation of chloroveratroles and chloroguaiacols was a rapid reaction which apparently occurred concomitantly with metabolism of the growth substrate and took place in the order 4,5,6-trichloroguaiacol 3,4,5-trichloroguaiacol 3, 4, 5-trichloroveratrole tetrachloroveratrole tetrachloroguaiacol (ii) the polymethoxylated compounds ( chlo-
- 233-
roveratroles and 1,2,3-trichloro-4,5,6-trimethoxybenzene) were de-a-methylated sequentially (iii) all of the resulting chlorocatechols were stable until the cultures had exhausted the growth substrate. Selective dechlorination then took place with the formation of 3,4,6-trichlorocatechol from tetrachlorocatechol, and 3,5-dichlorocatechol from 3,4,5-trichlorocatechol. Whereas these results provide a rational basis for the predominance of chlorocatechols in natural sediment samples, they give rise to a new puzzle: the presence of chloroguaiacols. Provisionally, we can only suggest that, in some way, in the sediment phase these are shielded from the inf 1 uence of the endogenous anaerobic bacteria. Clearly, the fate of the chloroguaiacols depends critically on the oxygen tension of the system under aerobic conditions, 0-methylation to chloroveratroles may be the major reaction, whereas under anaerobic conditions de-a-methylation to chlorocatechols would be expected to dominate. Current investigations are directed to elucidating the ultimate fate of chlorocatechols, and to exploring the transformations during growth with substrates structurally more closely related to compounds which plausibly occur in the sediment phase.
VERIFICATION EXPERIMENTS
We shall only briefly refer to some experiments which have been directed to assessing the environmental significance of the experiments discussed above. Clearly, analysis of contaminated sediment samples broadly supports the conclusions from the anaerobic experiments, although interpretation of dechlorination processes cannot be carried out unequivocally since substantial amounts of catechols with only one or two chlorine atoms occur. The finding of chlorinated veratroles in samples of fish recovered from contaminated localities (7) supports the existence of these compounds in the water mass, and they must almost certainly have arisen by aerobic a-methylation of chlorinated guaiacols. Their presence in fatty tissue is consistent with the bioconcentration potential of the more highly chlorinated veratroles (7).
We have attempted to carry out experiments using natural sediment samples spiked with chloroguaiacols and chloroveratroles, and incubated under appropriate conditions of oxygen tension. The results of these experiments (11) confirmed in their entirety the conclusions which we drew from the earlier laboratory experiments using pure cultures or metabolically stable consortia. Some details such as the relative rates of the reactions, their dependence on the structure of the growth substrate, and the influence of substrate concentration are, however, clearly more readily accomplished in properly designed experiments uncomplicated by the presence of the sediment phase.
On the basis of these findings, we feel that the methodology which we have developed provides a satisfactory procedure for assessing the fate and persistence of xenobiotics discharged into the aquatic system. We should not ,however, wish to leave the impression that there are no remaining unresolved issues. Of these, we feel that at least two merit extensive investigation: (i) the bioavailability of xenobiotics - and their metabolites - bound in the sediment phase (ii) the mechanistic basis for the metabolic cohension of reactions carried out by anaerobic bacteria. These are
- 234-
among the problems currently under intensive investigation in our laboratory.
ACKNOWLEDGEMENTS
We thank the Swedish Paper and Pulp Association and the research committee of the National Swedish Environment Protection Board for partial financial support of these investigations.
REFERENCES
1. Alexander, M. 1975. Environmental and microbiological problems arising from recalcitrant molecules. Microbiol. Ecol. 2.: 17-27.
2. Allard, A.-S., M. Remberger, and A.H. Neilson. 1985. Bacterial O-methylation off chloroguaiacols effect of substrate concentration, cell density, and growth conditions. Appl. Environ. Microbiol. 49 : 279-288.
3. Allard, A.-S., M. Remberger, andA.H. Neilson. 1987. Bacterial O-methylation of halogen-. substituted phenols. Appl. Environ. Microbiol. 53 : 839-845.
4. Atlas, E., K. Sullivan, and C.S. Giam. 1986. Widespread occurrence of polyhalogenated aromatic ethers in the marine environment. Atmos. Environ. 20 : 1217-1220.
5. Heitkamp, M.A., J.A. Freeman, and C.E. Cerniglia. 1987. Naphthalene biodegradation in environmental microcosms estimates of degradation rates and characterization of metabolites. Appl. Environ. Microbiol. 53 : 129-136.
6. Haggblom, J. Apajalahti, and M. Salkinoja-Salonen. 1986. Metabolism of chloroguaiacols by Rhodococcus chlorophenolicus. Appl.-Microbiol. Biotechnol. 24 : 397-404.
7. Neilson, A.H., A.-S. Allard, S. Reiland, M. Remberger, A. Tarnholm,T. Viktor, and L. Landner. 1984. Tri- and tetrachloroveratrole, metabolites produced by bacterial O-methylation of tri- and tetrachloroguaiacol : an assessment of their bioconcentration potential and their effects on fish reproduction. Can. J. Fish. Aquat. Sci. 11 : 1502-1512.
8. Neilson, A.H., A.-S. Allard, andM. Remberger. 1985. Biodegradation and transformation of recalcitrant compounds, p. 29-86. In. O. Hutzinger (ed.) Handbook of Environmental Chemistry. Vol. 2jC. Springer-Verlag, Berlin.
9. Neilson, A.H., A.-S. Allard, C. Lindgren, and M. Remberger. 1987. Anaerobic transformations of chloroguaiacols, chloroveratroles , and chlorocatechols. Appl. Environ. Microbiol. 54 : In press.
10. Racke, K.D., and J.R. Coats. 1987. Enhanced degradation of isofenphos by soil microorganisms. J. Agric. Food Chern. ~ : 94-99.
11. Remberger, M., A.-S. Allard, andA.H. Neilson. 1986. Biotransformations of chloroguaiacols, chlorocatechols and chloroveratroles in sediments. Appl. Environ. Microbio::'. 21 : 552-558.
12. Xie, T.,K. Abrahamsson, E. Fogelqvist, andB. Josefsson. 1986. Distribution of chlorophenolics in a marine environment. Environ. Sci. Technol. 20 : 457-463.
- 235-
POSTER SESSION III TRANSFORMATION
Levels of chlorophenols in the river, ground and drinking water in the Zagreb area
LEVELS OF CHLOROPHENOLS IN THE RIVER, GROUND AND DRINKING WATER IN THE ZAGREB AREA
S. FINGLER and V. DREVENKAR Institute for Medical Research and Occupational Health,
University of Zagreb, Zagreb, Yugoslavia
S\lITI1la.YY
The presence of 4-chloro-, 2, 4-dichloro-, 2,4,5:- and ~, 4, 6-trichloro-, 2,3,4,6-tetrachloro- and pentachlorophenol was ~nvest~gated ~ samples of the Sava river water and of ground and drinking water collected in the Zagreb area. In alnDst all of the 25 river water samples higher chlorophenols were detected, with pentachlorophenol present in highest concentrations, but not exceeding 160 ng cJm-3• The levels of chlorophenols in ground water samples collected in 18 wells depended on the location and distance of the well from the Sava river bed. Their presence in the river water did not appear to have an effect on the purity of ground water sampled from wells about 2 kIn away from the river, among which ten supplying water to the Zagreb waterworks. Hewever, single chlorophenols were detected in samples of both chlorinated municipal drinking water and non-chlorinated drinking water from privately &ned wells. The occurrence of chlorophenols in drinking water samples seerred to be mainly a consequence of occasional contamination of ground waters network in the city of Zagreb, where waterworks pwnpstations and private wells are located in certain city areas often near industrial plants.
1. INTRODUCTION
The presence of chlorophenols has been detennined at highly variable concentrations in different types of surface waters as a consequence of municipal and industrial waste discharge (1-4). An investigation of organic micropollutants in the Sava river, the longest (940 kIn) Yugoslav ri ver which is reputed to have been contaminated for many years, indicated the presence of chlorophenols in river water samples collected just before and after the discharge point of municipal waste water effluent of the Zagreb city (5). Compared to the other river waters (3,4) the concentrations of tetrachlorophenol (18-26 ng cJm-3) and pentachlorophenol (35 ng cJm-3) were lower and the effect of the Zagreb waste water discharge on chlorophenol levels in the river was found to be negligible.
Before entering the Zagreb area in a length of approximately 200 kID the Sava river flews through rural and urban regions in the north-west of Yugoslavia. Directly or through tributaries it receives effluents from nurrerous industrial facilities among which from a pulp and paper mill and from a nuclear pewer plant. The ground waters in the Zagreb area, which after a treatrrent by chlorination serve as municipal drinking water supplies, are mainly recharged through the river water infiltration. Therefore the impact of the river water pollution on the purity of ground waters and consequently on the quality of drinking water is of great concern. Recently, the city supplies of drinking water have been contaminated on several occasions by uncontrolled underground industrial waste
- 238-
discharge in the city area. In this paper a screening investigation of the presence of chloro
[.henols in samples of the Sava river, ground and drinking waters collected in the Zagreb area is described. All water samples were analysed for 4-chlorophenol (p-CP), 2,4-dichlorophenol (DCP), 2,4,5- and 2,4,6-trichlorophenol (2,4,5-TCP and 2,4,6-TCP), 2,3,4,6-tetrachlorophenol (TeCP) and pentachlorophenol (PCP). The levels and species of chlorophenols in different water samples were corrpared in order to explain their presence in drinking water and to indicate their possible origin.
2. EXPERIMENTAL
2.1. Sampling
The sampling sites of the river, ground and drinking waters in the Zagreb area are presented in Fig. 1. The sampling period lasted from March 1984 to July 1986.
A total of 25 samples of the Sava river water were collected at the 712 km of the river course, where the river enters the Zagreb area (location 1), and at the 693 km of the river course, which is near the city industrial zone (location II).
Ground water samples were collected from 15 wells, 10- 32 m deep, at a distance of about 2 km from the river and from three 9-12 m deep wells, o • 2 to 0.4 km from the river bed at location II.
The presence of chlorophenols in drinking water was analysed in ten tap water samples from the municipal water supply and in nine water samples from private 8-11 m deep wells.
2.2. Procedure
Ollorophenols were determined in 500 ml water samples by rreans of an analytical method based on the procedure described by Renberg and Lindstrom (6). Four parallel samples were prepared from each water. After C18 reversed-phase adsorption enrichment (Sep-Pak C18 cartridge) the chlorophenols present in the acetonic eluate were converted in two and two duplicates to acetyl (6) and pentafluorobenzoyl (7) derivatives. All samples were analysed on two basically different gas chroIlE.tographic colunns using an electron capture detector. Only signals observedonboth colwms for both types of derivatives were evaluated. The corrpounds were identified by their retention tines as compared with knGm. standards, which were prepared by spiking the Sep-Pak C18 cartridge with knGm. anounts of chlorophenols and by subsequent treatrrent of the ace tonic eluate using the sane procedure as for the samples.
3. RESULTS AND DISCUSSION
In almost all of the 24 river water samples collected in the period March 1984 - February 1985 2 ,4,6-TCP, TeCP and PCP were determined, with PCP present in highest concentrations. l.cMer chlorophenols were not detected at all. The concentration ranges (rredian) of twelve oonthly determinations for 2,4,6-TCP, TeCP and PCP at location I were <3-39 (12), <10-21 (13) and <10-99 (58) ng clrrr3. Very similar values were determined at Location II: <3-37 (8) ng dm- 3 for 2,4,6-TCP, <10-27 «10) ng dm-3 for TeCP and <10-66 (43) ng dm-3 for PCP. No great variations in chlorophenol concentrations in respect to the tine of sampling and the river flow rate was noticed and TeCP and PCP levels were about the sane as determined at location II two years earlier (5). However, in April 1986 the concentrations of 2,4,6-TCP, TeCP and PCP in the river water at that location were
- 239.-
approxirrately twire as high as the rraximum values determined in the 1984-1985 period (Table I). In addition to these compounds DCP was also detected. In simultaneously collected ground water sanples from three wells 0.2-0.4 kIn from the river bed near Location II, in an area anticipated to become a future site of a waterworks pumpstation (Fig. 1), the levels and species of chlorophenols were almost identical to those in the river (Table I). This indicated that the compounds were not affected by any elimination processes in the grotmd. The sorption of 2,4 ,6-TCP, TeCP and PCP on the soil and two sediments in this area was checked by preliminary analy,sis. The sanples of soil and sediments taken at depths of 0.2-0.4 m, 27.5-27.8 m (clay) and 36.2-36.7 m (sand) contained 97.4,4.2 and 0.3 mg 6 1 of organic carbon, respectively. At initial chlorophenol concentratiors in water of 100 llg dm- 3 and at grotmd water soil (sediment) ratio 10: 1 a measurable sorption was notired only on the soil layer, which is in accordance with its highest organic carbon content (8). After a 24-hour equilibration at room temperature the initial concentrations of 2,4,6-TCP, TeCP and PCP decreased by 55, 60 and 40%, respectively.
A connection between chlorophenol presence in the ground waters and the location of wells and their distance from the river bed was indicated by the results of chlorophenol analysis in 15 wells between Locations I and II, about 2 krn away from the river. In none of these distant wells was the influence of chlorophenol presenre in the river water noticed. In five wells p-CP, OCP, 2,4,6-TCP and PCP concentrations were within the detection limits (2-6 ng dm- 3) and in further 10 wells, supplying water to the Zagreb waterworks, chlorophenols were not detected at all. At the same time PCP concentrations in the monitored segment of the Sava river were 49-78 ng dm-3 and the river load of PCP was 27-47 mg s-l.
Although chlorophenols were not detected in six out of ten chlorinated mtmicipal drinking water sanples from the checked waterworks wells, single corrpounds were present in concentrations of <10-70 ng dm- 3 and in one sarrple PCP concentration was as high as 123 ng dm-3 (Table II). Similar concentrations were measured in non-chlorinated drinking water sarrples from four out of nine checked privately owned wells, except the TeCP and PCP concentrations in one well of 270 and 420 ng dm- 3, respecti vely (Table II). The presence of chlorophenols in both chlorinated mtmicipal drinking water and in non-chlorinated drinking water from private wells suggests that they were of different origin and that their occurrence was not only a consequence of regular treatment of milllicipal drinking water by chlorination (9). Most of the waterworks pumpstations and the private wells· were located more than 2 krn from the Sava river but in a populated area, often near industrial plants. Occasional contamination of the grotmd waters network by illlcontrolled industrial or domestic waste discharge is therefore a serious problem, because it affects the groillld water quality in wells which especially if privately a-med, are not adequately protected.
REFERENCES
(1) Fotmtaine, J.E., Joshipura, P.B. and Keliher, P.N., Water Res.l0 (1976)185-188. --
(2) Matsumoto, G., Ishiwatari, R. and Hanya, T., Water Res.l1(1977)693--698.
(3) Wegman, R.C.C. and Hofstee, A.W.M., Water Res.13(1979)651-657. (4) Schwarzenbach, R. P., Giger, W., Hoehn, E. and Schneider, J. K. ,
Environ.Sci.Technol. 17(1983)472-479.
- 240-
(5) Ahel, M. and Giger, W., Kern. Ind. 34(1985)295-309 (in Croatian). (6) Renberg, L. and Lindstrom, K., J. Chromatogr. 214(1981)327-334. (7) Renberg, L., Chemosphere 10(1981)767-773. (8) Schellenberg, K., Leuenberger, Ch. and Schwarzenbach, R.P.,
Environ. Sci. Technol. 18(1984)652-657. (9) Chriswell, C.D., Chang,R.C. and Fritz, J .S., Anal. Chern.
47(1975)1325-1329.
Table I
O1lorophenol concentrations in the Sava river and ground water samples collected in April 1986
Sample -3 Concentration, ng dIn
(distance from the river bed, DCP 2,4,6-TCP TeCP PCP depth)
Well 1 61 23 46 151 (0.4 kIn, 9 m)
Well 2 10 13 32 117 (0.4 kIn, 9 rn)
Well 3 9 24 33 116 (0.2 kIn, 12 m)
Sava river-Location II 20 62 69 163
The results are mean values of four determinations.
Table II
O1lorophenol concentratiomin drinking water samples from the municipal water supply (A) and from private wells (B) collected in June and July 1986 in the Zagreb city area
Concentration ranges -3 Compound median), ng dIn
A (N=10) Bf'(N=9 )
p-CP < 2- 6« 2) < 2- 10 « 2)
OCP < 8- 17« 8) < 8- 30 « 8)
2,4,5-TCP < 6- 22« 6) < 6- 39 « 6)
2,4,6-TCP < 3- 9« 3) < 3- 29 « 3)
TeCP <10- 10 «10) <10-270 «10)
PCP <10-123( 13) <10-411 (<10 )
N = number of samples
,', Private wells depth: 8-11 m
- 241-
·A tA e
d n i c a d "e
I I I
r----.J I I I L---- l
Centre
703 km A I ~ _________________ L-
J( X xXx Xf f
f .t
f
Fig. 1 Sampling sites of the river, ground and drinking waters in the Zagreb area
x ground water wells
A drinking water from private wells • muniCipal drinking (tap) water
central city area
streamlets in the Zagreb area
- 242-
SESSION IV
WATER TREATMENT
Chairmen A. RIVERA-ARANDA and N. NYHOLM
Biological-chemical characterization of effluents for the evaluation of the potential impact on the aquatic environment
Test methods and strategies for environmental management purposes - environmental fate testing of chemicals and effluents
Mass fluxes of linear alkylbenzenesulphonates, nonylphenol, nonylphenol mono- and diethoxylate through a sewage treatment plant
Mutagenic compounds in chlorinated waters
The formation and removal of chemical mutagens during drinking water treatment
Application of the ozone-hydrogen peroxide combination for the removal of toxic compounds from a groundwater
BIOLOGICAL-CHEMICAL CHARACTERIZATION OF EFFLUENTS FOR THE EVALUATION OF THE POTENTIAL IMPACT ON THE AQUATIC ENVIRONMENT.
O. SVANBERG and L. RENBERG
National Environmental Protection Board, The Emission and Product Control Laboratory, S-611 82 Solna, Sweden
Summary
The approach of using integrated programs of biological and chemical tests for characterization of industrial effluents has been applied for hazard assessment by the regulatory agences in Sweden for some years now. Most cases are connected to the claim on municipal and industrial facilities to obtain discharge permits. In other cases testing programs have been used for directing process modifications or external treatment techniques towards a minimization of the environmental effects from the discharge.
The biological testing has found the widest and most efficient application for industrial plants with processes emitting complex effluents with largely unknown chemical composition and environmental impact.
The test programs are applied in a flexible manner depending on the type of industry with its process modifications, raw materials and chemicals added. The character of the receiving water body is taken into consideration when choosing the test organisms.
1. INTRODUCTION The work which has previously been devoted to reduce
negative enviromental effects from water pollution has been mainly directed towards two goals.
Firstly, measures have been taken to reduce or eliminate easily observed effects e.g. eutrophication caused by the release of nutrients, and acute fish kills due to toxic discharges.
Secondly, the identification of toxic and bioaccumulating organic compounds (e g PCB and DDT) resulted in attempts to reduce or prohibit the use of these types of compounds. In addition a considerable amount of work has been carried out to develop specific and sensitive analytical methods to be able to monitor the distribution of persistent compounds in the environment.
The great number of potentially harmful organic compounds
-~-
with a variety of chemical structures in industrial aqueous discharges makes a chemical characterization difficult.
Hazard assessment of a complex industrial effluent based solely on chemical characterization is usually unsatisfactory. Even the use of the sofisticated analytical methods cannot identify and quantify all pertinent compounds especially not those which are not water soluble or have a high molecular weight.
The use of an integrated approach in which both analytical chemical methods and biological tests are employed have shown to be highly cost effective (1,2). The primary advantages of applying analytical methods for quantifying specific compounds is that the methods can be very specific and sensitive. The main disadvantages are that the content of specific compounds in industrial waste water is usually unknown to a large extent. As previously discussed the identification techniques more or less routinely applied are relatively expensive and usually not able to identify other than relatively non-polar and low molecular compounds.
The use of biological tests allows for a more direct hazard assessment than chemical methods do, as it is an integrated test for all individual compounds including synergistic effects. Problems with biological tests include interspecies variability among the organisms and choice of pertinent systems.
The most difficult types of compounds to test and evaluate using biological screening tests are hydrophobic compounds as they usually respond after a longer period of exposure than is usually employed. These compounds are also adsorbed onto particulate matter, which makes the choice of test system complicated. On the other hand such compounds can usually be easily analyzed using standard analytical methods for micropollutants e.g. gas chromatography. A conclusion to be drawn is that chemical characterization complements biological characterization in a very cost efficient way.
As the use of chemical analytical methods for the characterization of aqueous discharges is covered elsewhere (e.g. COST 641 Working Party 1 and 4) the following discussion will be concentrated to the regulatory application of biological test systems for the prediction of the environmental impact of industrial effluents.
2. METHODS The test methods used to characterize the effluent samples
are arranged in a tiered system (Fig 1). The main strategy for Level I is to screen the discharge for its potential toxicity and for the presence of persistent and bioaccumulating substances and to support this by mainly the standard water quality parameters (BOD, COD, conductance, suspended material, pH). If a particular substance is known to occur and is suspected to be present in toxic concentrations it is also quantitatively analysed. Group specific analysis e.g. oil (IR spectrometry), AOX (adsorbable organohalogens), are applied depending on the industrial branch investigated.
Based on these results it is, in most cases, possible to make a judgement whether the discharge is likely to be harmless
-w-
Figure 1. Flow sceme of information for the regulatory use of biological-chemical characterization of effluents.
EXISTING INFORMATION
-PROCESSES
-RAW MATERIALS/ADDITIVES
-PRODUCTS/BY PRODUCTS
-CHARACTER OF
RECEIVING WATERS
-RESULTS FROM PREVIOUS
INVESTIGATIONS
COMPILATION/EVALUATION ~----l
DECISION/MEASURES
-PROCESS MODIFICATIONS
-CHANGES IN WASTE WATER
TREATMENT
-MONITORING PROGRAMS
-REPLACEMENT OF ADDITIVES
YES
- 246-
11/
II
CHARACTERIZATION TESTS
- CHEMICAL CHARACTERIZATION
- BIOLOGICAL EFFECTS
- PERSISTENCE
or potentially harmful to the environment. This might be a sufficient basis for a regulatory decision by the permit writer. If on the other hand the results are not clear-cut the investigation proceeds to Level II or Level III. This might also be the case if the economical cost to meet the decision is high or if the receiving water body is regarded as especially valuable. Data from more valid tests are often requested in such cases.
The toxicity test methods of Level I (Fig 2) are mainly the short term, standardized methods used now in many countries, and described for the application on pure substances by OECD (3). In order to meet the necessary demand however on flexibility and to move a bit closer towards the actual receiving water body, it is recommended to use test organisms normally living in the particular ecosystem.
Figure 2. Tests and analyses for characterization of effluents, Level I
Chemical characterization
COD BOD, TOC pH Susp. matter Conductivity Lipophilic compounds (TLC)
Biological characterization
Fish 96-hr LC50 Crustacean 96-hr LC50 (48) Algae 3-5 d EC50 Higher plants 3-5 d EC50
Optional
- DOC
- Known or suspected metals or organic substances
- N, P, oil, TOCl, AOX
Optional
- Activated sludge inhibition
- Microtox
The Level II testing is based on the results from the Level I. The toxicity tests are short or medium long term sublethal tests (Fig 3) and they are planned to verify the results from Level I or to establish a no observed effect concentration, (NOEC). By the use of standard biodegradation tests the bulk of the organic matter in the sample is further characterized. The chemical analyst extends the specific analysis or, depending on the case extends the knowledge on the organic matter by GC/MS analyses.
-w-
Figure 3. Examples of tests and analyses for characterization of effluents, Level II.
Chern. characterization Follow up with advanced technique GC/MS, HPLC etc.
Biodegradability Die-away or respirometric test for easy biodegradability
Toxicity Extended tests based on results from I.
Fish embryo-larval test (sebrafish) larval growth (fathead minnow) physiological effects
Crustacean reproductions test
Mussel Mytilus larvae survival
Algae "Microtitertest"
Genotoxicity Ames test
The Level III tests are launched only when there is a high demand on accuracy in the prediction of hazardness. This may be the case when the cost for taking proper measures against the pollution is high or when there are uncertanties with a risk of serious damage to large or especially valuable water bodies.
In contrast to Level I which is composed of a fairly rigid set of standardized tests, the Level III investigations are planned more as a research project. The separate tests are rather time consuming and there could only be listed a number of recommendations for these (Fig 4).
The results of the investigations are evaluated on a case by case basis. The results from the toxicity tests are considered in relation to the dilution capacity of the receiving water body. Mathematical models are used for illustrating the area likely to be influenced by toxic effects. The highest attention, however, is paid to the presence of persistent and heavily biodegraded substances in the discharge. There is a newly issued policy within the Swedish Environmental Protection Board (4) to take every opportunity to reduce the discharges of such substances.
An illustrative way to compare thetoxicity from different types of waste waters is by applying the concept of toxicity emmision factor (TEF) (5,6). The calculation is based on the LC50 or EC50 values, respectively, (i.e. the percentage of waste water which causes an effect on 50 % of the test organisms) :
TEF = 100 • (1/EC50) • water flow
- 248-
Figure 4. Examples of tests and analyses for characterization of effluents, Level III.
Chern. characterization Specific analyses with advanced technique
Bioaccumulation Analyses of fish, mussels exposed in the lab. or in cages in the receiving water or in the field
Biodegradability Ultimate biodegradability test Simulation test
Toxicity Laboratory exposure
Cage exposure
Field
Toxicity, community level
Micro/mesocosms studies
- delayed effects - physiol. effects - morphol. effects
- survival - physiol/morphol. effects - organoleptic test
- physiol/morphol. effects
- increased tolerans
- litoral communities - soft bottom communities
Thus the use of TEF-values enables the comparison of different effluents as the concept takes both the degree of toxicity and the waste water volume into consideration.
3. CASE STUDIES Municipal Sewage Treatment Plants
Samples from a number of municipal wastewater treatment plants were tested for toxic and genotoxic compounds. The aim of the investigation was to settle whether some available toxicity screening tests were sufficiently sensitive for use in effluent monitoring of existing Swedish treatment plants. If so, were there any differences in the toxicity of the treated waters as a result of different treatment efficiency or to the quality of the incoming wastewater.
The tests applied in the first stage were:
- Daphnia 48-hr LC50 - Sebrafish 96-hr LC50 - Selenastrum 5-d EC50 (green algae) - Ames test (mutagenicity)
Samples of wastewater before and after treatment were taken from 7 plants, brought to the testlaboratories and immediately deep frozen.
-~-
The results from the tests show that the treated waste water from some of the plants were still slightly acute toxic to sebrafish and algae. Daphnia were less sensitive showing almost no mortality in the undiluted samples. There was a tendency to higher toxicity in the samples from plants receiving industrially polluted wastewaters compared to those with only domestic wastes.
The effects of the treatment processes seen as reduced toxicity is not obvious for all plants. With still another series of tests, however, selecting the tested dilutions on the present results there should have been a better resolution showing more acurate the effect of the treatment in the plants.
The toxicity in the wastewaters can not be correlated to heavy metals analyzed in the samples. Zinc and copper were present in fairly high concentrations reported to be toxic when tested as pure metal salts. The content of organic matter is obviously inhibiting their toxicity by complex formation.
The Boras treatment plant was also the subject of testing with more sensitive tests on sublethal effects. In these tests (Table 1) Daphnia were more sensitive than sebrafish. Rainbow trout exposed to the wastewater developed symptoms of physiological effects. The liver mixed function oxidase enzymes (MFO) were significantly induced at 100 times dilution of the incoming water and at 10 times dilution of the treated water. This indicates the presence of compounds in the wastewater with the same effect as polyaromatic hydrocarbons.
Table 1 Toxicity testing of waste water from Boras municipal waste water treatment plant
Acute tests
Sebrafish 96-hr LC50 Daphnia 48-hr EC50 Selenastrum 5-d EC50
Chronic tests
Sebrafish, embro/larvae survival hatching of eggs
Daphnia, young survival 22-d LC50
Induction of MFO
Genotoxicity
Ames test
untreated (vol. %)
75-100 48 50-100
> 25 50
> 3,1 7,2
< 1
++
treated (vol. %)
» 100 > 100 > 100
> 50 > 100 > 25
79 < 10
+
The untreated as well as the treated water also contained compounds with slight mutagenic effects as shown by Ames test.
-~-
The experience from this investigation was that the applied test methods were able to rank the waste waters according to toxicity. Industrial discharges rendered the waste waters a higher toxicity. The results verify the present policy of encouraging internal measures by the industry to reduce the discharges of toxic chemicals.
Chemical plant A chemical plant producing a great variety of chemical
products made an application for a permit to discharge effluents from the process. The receiving water body is a wide bay of the Bothnian Bay. The water body is badly polluted since many years, also by severeal other plants e.g. pulp and paper mills, an aluminia smelter and by domestic discharges. Results from monitoring the water quality in the sea outside the plant will therefore be difficult to interpret in terms of "who contributes to which effect?".
The company was asked to provide results from ecotoxicological testing of the total effluent to guide the evaluation of the potential hazardness of the discharge.
The effluent was sampled from a lagoon separated from the sea by a barrier of coarse stones. To characterize the potential toxicity the following screening tests were applied on the samples:
- Microtox l5-min EC50 - Sebrafish 96-hr LC50 - Scenedesmus (algae) 3- and 5-day EC50 - Ames test (mutagenicity) - Activated sludge respiration inhibition
As it was soon realized that the water in the lagoon was acutely toxic also a Sebrafish egg/embryo test was run to further strengthen the indication of hazardness.
The content of organic substances were recorded as COD and TOC and to further characterize this, the amount of hexanextractable substances was analysed. A 28-d test on readily biodegradability was run in which DOC was analysed to trace the degradation rate.
One single compound in the effluent was known to be toxic. This compound Dinoseb (2-sec-butyl-4,6-dinitrophenol) was therefore specifically measured.
The results (Table 2) was interpreted in terms of "a rather toxic effluent containing heavily biodegraded compounds". The Environmental Protection Board considered this as a not acceptable discharge and claimed for measures to be taken.
To treat the whole effluent was not economically and technically possible. The company therefore agreed to investigate the various subeffluents from the processes to find out if the hazardous substances were released from a certain source. The total effluent comprise eight subeffluents and the toxicity of each of these were screened by the Microtox assay complemented by COD analyses.
- 251-
Table 2 Characterization of the total effluent from a chemical plant.
Activated sludge resp. inhib.
Microtox ECSO, lS min
Sebrafish, LCSO, 96-h
_11- larvae, survival
Algae, Scenedesmus ECSO, 3-d
S-d
Biodegradability, DOC 28-d
Mutagenicity
COD TOC Dinoseb Extractable matter
Effect concentration (vol%)
O.S
40 ( 3 • 2)
24
10
1.4
0.72
S4 %
300 mg/l 67 mg/l 14 ~g/l 17 mg/l
The ECSO values varied from 0.4 to 31. 4 vol. % for .the samples (Table 3). There were no correlation between toxicity and COD-content. The test results in terms of ECSO however only consideres the concentration of a toxicant in a water sample. In order to visualize also the total amount of toxicants the toxicity emission factors (TEF) were calculated. The TEF value indicates the volume of water needed to dilute the discharged effluent to a concentration of ECSO.
By this procedure the subeffluents were divided into three separate categories: TEF around 20, TEF around 200 and with TEF 2-6000. And with this as a basis the Environmental Protection Board argued infront of the permit writing authority that the subeffluents with TEF 2-6000 should be treated with priority.
The fact that the total effluent was more toxic than was expected from the toxicity of the subeffluents is probably due to one of the alternative reasons:
more than additive toxicity when combining the subeffluents (synergistic effects) all subeffluents were not identified and ineluded in the study
- 252-
Table 3 Characteristics of subeffluents from a chemical plant
Flow (m 3 /d)
A 46
D 28
C 8
D 5
E 67
F 41
G 41
H 900
Total 1445
'l'OC (kg/d)
55
44
1
1
41
20
21
166
Microtox TEF (EC50, 15 min) (1/EC50·100·flow)
0.9 5100
0.5 5600
0.4 2000
23.1 22
31.4 213
1.4 2900
1.4 2900
15.6 5800
3.8 38000
The Franchise Board in its decision prescribed the company to change the processes or to treat the toxic subeffluents. After the measures have been taken another biological-chemical characterization should be carried out on the waste water in the lagoon.
Pulp mill effluents Bleachery effluents from pulp mill production comprise the
largest environmental impact compared to any other type of Swedish industrial aqueous discharges. This fact is due to mainly two factors.
Firstly, bleachery effluents exert a variety of toxic effects against many types of aqueous organisms including acute and sublethal effects on fish, crustaceans and algae.
The extensive flow rate (usually 0.5-1 m3 /sek) of the outgoing waste water from a pulp mill in combination with the toxic properties have resulted in that massive measures (internal and/or external) have been or are being taken. The broad span of potential toxic effect originating from chemical plants and pulp mills, respectively, can be illustrated using the TEF concept. Acute toxicity tests with Sebra fish shows a TEF-value of 250 m3 /h for the chemical plant previously dicussed. The corresponding TEF-value for a bleached kraft mill is approximate 9 000.
Secondly, the effluents contain organic chlorinated compounds, known or suspected to be persistent, which are released in large amounts (typically 2- 10 metric tons per day
- 253-
Several approaches have been investigated how to reduce both the effect and the amount of chlorinated compounds released. One of the most successful ways has been to modify the bleaching process. Both the introduction of oxygen bleaching and replacement of elemental chlorine with other chemicals (mainly chlorine dioxide) result in a reduction of the toxicity and the load of organochlorines.
The reduction of the use of elemental chlorine used for bleaching soft wood pulp is reflected in both the toxicity and amount of organic chlorinated compounds. One of the most important groups of such compounds emitted from pulp mills are chlorinated phenols. The reduction of elemental chlorine used in the first bleaching step with approximate 40 % has shown to reduce the amounts of released polychlorinated phenols, guaicols and catechols from 41 kg/day down to 22 kg/day (Table 4).
Biological and chemical characterizations have exensivly been used as a tool to reduce the environmental impact of pulp mill effluents. The applications of these methods (7) for evaluating different bleaching sequences and external methods have been of great importance.
Table 4 Chlorinated phenolic compounds in bleachery effluent (~g/l)
HIGH = 85 % chlorine in the first bleaching stage LOW = 50 % chlorine in the first bleaching stage
2,4,6-TCP 2,3,4,6-TeCP 3,4,S-TCG 4,S,6-TCG TeCG 3,4,S-TCC TeCC
TOTAL
T = tri Te = tetra CP = chlorophenol CG = chloroguaicol CC = chlorocatechol
HIGH (kg/day)
1.0 2.0 9.8 0.8 4.S
14.8 7.7
40.6
LOW (kg/day)
0.7 0.4 6.S 1.2 1.0
11.3 1.3
22.3
4 CONCLUSION The frequent use of biological tests together with
chemical analysis to obtain a base for hazard assessment of industrial and domestic waste water effluents has provided the regulatory agencies with a resource saving tool for better decisions (8).
Experience from the passed years together with technical
-~-
and scientic progress is now brought into a revised and updated guidance document which will be issued within short.
REFERENCES
(1) BENGTSSON, B-E and RENBERG, L The use of chemical and biological parameters to characterize complex industrial effluents. Regulatory Toxicol Pharmacol ~ (1986) 238.
(2) Organization for Economic Co-operation and Development. The use of biological tests for water pollution assessment and control. ENV/WAT/86.1 (1986).
(3) Organization for Economic Co-operation and Development. OECD Guidelines for Testing of Chemicals. Paris 1981.
(4) National Swedish Environmental Protection Board. Action Plan for Marine Pollution. May 1987.
(5) SPRAGUE, J.B. Measurement of pollutant toxicity to fish. Water Research i (1970) 3.
(6) WONG, A. et al. Toxicity, BOD and color of effluents from novel bleaching processess. Pulp and Paper Canada 79 (1978) 41.
(7) SODERGREN, A. BENGTSSON, B-E. JONSSON, P. LAGERGREN, S. LARSSON, A. OLSSON, M. and RENBERG, L. Summary of results from the Swedish Environment/Cellulose. Symposium, Forest Industry Wastewaters. Tampere, Finland, June 2-12 (1987).
(8) National Swedish Environmental Protection Board Characterization of industrial effluents. SNV Meddelande 6/82 (1982) (in Swedish).
-lli-
TEST METHODS AND STRATEGIES FOR ENVIRONMENTAL MANAGEMENT PURPOSES - ENVIRONMENTAL FATE TESTING OF CHEMICALS AND EFFLUENTS
Surmnary
N. NYHOLM Water Quality Institute
Denmark
A survey is presented of practical management oriented methods and strategies for environment fate testing of chemicals and effluents with main emphasis on biodegradability. The major framework of the paper is the EEC legislative tests and current international cooperative work on test development and test harmonization. It is pointed out that test methods and strategies developed for use in connection with notification and general regulation of chemicals can be used in modified form for testing of specific effluents.
1. INTRODUCTION Practical management oriented test methods for predicting the
environmental fate of chemical substances (including constituents in effluents) are currently undergoing a rapid development under international auspecies, and particular attention is given to methods for assessment of biodegradation, which process is the major elimination route in aquatic environments for most chemicals.
In general terms, the prediction of environmental chemical fate under actual or hypothetical scenarios is the basis for determining the exposure (actual or potential) of aquatic organisms or of humans, and consequently for making any assessment of effects or risks. Chemical fate predictions are also needed for more general environmental management of chemicals, as increasing chemical contamination of the environment is regarded as a long term concern, even if effects are absent.
In some situations qualitative assessments may be sufficient, like for example "readily degradable and not bioaccumulable" or even "persistent", while in others, estimates of kinetic rates of biodegradation and more precise information on the environmental partitioning are required. Quantitative assessments of biodegradation rates in particular must necessarily relate to specific environmental scenarios, as biodegradability is a combined property of the chemical and its environment, which fact has been stressed so often. To this end it may be of special importance to account for adaptation phenomena, as degradation rates may differ widely between non-adapted and adapted systems, respectively. Detailed biochemical or microbiological studies, on the other hand, on metabolic pathways etc. are normally not required for practical management purposes, although such studies can be very desirable and can be an important basis for generalization of results. It should be mentioned that from a practical viewpoint any stable chemical intermediate can be regarded and handled as an individual compound.
- 256-
There is an interplay between the general regulation of production, trading and uses of chemicals and the envirorunental management of specific chemical discharges, like for example discharges of industrial effluents. First, a lot of common information on the environmental properties of the chemicals of concern is needed for both uses, including in particular properties related to envirorunental fate (physical/chemical parameters, biodegradability in wastewater treatment plants, in surface water and perhaps in sediment). Second, for management of industrial effluents, a number of factors have to be considered, among which external treatment and effluent quality are just some. The effluent may also be made more envirorunentally acceptable by internal measures, such as substitution of process or auxillary chemicals or by modification of the process or even the end product. Such measures interplay with the general regulation of chemicals, or will at least do so in the future. An illustration of the chemical fate aspects to be considered for the management of industial wastewater discharges is presented in Figure 1. They include among others: 1) comparisons of the properties of concern chemicals with the properties of possible chemical substitutes (part of evaluation of feasible internal measures); 2) evaluation of the treatability of the effluent in a possible treatment plant of specific design (pretreatment or final treatment); 3) evaluation of the fate of concern chemicals in a municipal treatment plant; 4) evaluation of the dilution and the subsequent physical transport and mixing in the receiving water; and 5) evaluation of the chemical fate in the receiving water: (1) partitioning between water, sediments, suspended particulates and biota; (2) degradation (first of all biodegradation); and (3) volatilization into the atmosphere.
DISCHARGE OF CHEMICALS IN
INDUSTRIAL EFFLUENTS
Receiving water
Partitioning water!
sediment/biota
Degradation
Volat ilization
- 257-
Figure 1. Factors to be taken into account for the environmental management of chemicals discharged with industrial effluents. Chemical fate aspects.
Usable test methods for determining the relevant environmental fate properties of pure chemicals are available today and are undergoing rapid further developnent and international harmonization. Most work is related to the general legislations on chemicals, however, and with regards to testing complex effluents, no similar international cooperative endavours take place, unfortunately. National practices for testing effluents relate to national laws on environmental protection, and practices may differ widely, even within the same country. .
In the following, a brief survey is given on chemical fate test methods for chemicals and for effluents, as related to either general regulations of chemicals or to management of specific discharges. Main emphasis is put on biodegradation.
2. GENERAL STRATEGY A simplified general strategy for cost effective biodegradabili
ty testing of either chemicals or effluents for environmental management purposes can be outlined as follows, and is based in part on the philosophy which emerged from the OECO Chemicals Testing Programme (1) :
1) Screening tests for ready biodegradability. Such tests should be simple and cheap, and they are normally stringent, meaning that a positive test result indicates ready biodegradability in the environments that the test system models, e.g. river water or sewage treatment plants. Due to the stringency of the test, a negative test result does not necessarily mean that the test material is not biodegradable under environmental conditions, but indicates that more work will be necessary to establish this.
2) Tests for inherent or potential biodegradability or treatabigty. Such tests possess a high potential for biodegradation compared to screening tests as well as to most environmental conditions. The tests may still be simple (e.g. the Zahn Wellens test), but some are definitely more costly and labour consuming than the screening tests (e.g. the SCAS test). (See later for details on both methods). The philosophy behind the inherent tests is that if the test result is negative, it is not very likely that the test material will degrade under normal environmental conditions, and this means that in a number of situations further testing may not be worth while. A positive test result, on the other hand, does not guarantee environmental biodegradability, and consequently more testing is required to investigate this.
3) Simulation tests. Such tests aim at giving information about biodegradability under more realistic environmental conditions. Which may imply for example testing at low concentrations in surface water or testing in continuous flow through sewage treatment laboratory model systems like the OECO confirmatory test (see later). The information derived at this level of testing is quantitative and can be expressed in terms of rates and kinetics. Both site specific test systems and more general or standardized model systems are included in the category of simulation tests.
Simulation tests may be relatively complicated and costly, like for example treatability or degradabili ty studies in continuously operated laboratory sewage treatment plants, not to mention pilot plant studies with actual industrial effluents. Simulation tests need not be expensive, however, just because they are designed to give information on kinetics. Die-away tests in natural surface water performed with chemicalS in low concentrations using l4-C-
- 258-
labelled substances are relatively cheap and simple, provided appropriate general facilities for handling radio-labelled materials are at disposal - and provided that radio-labelled test material is available. If synthesis of 14-C-1abe11ed test material is required, the total costs of a study may of course mount up.
A simplified outline of the official EEC strategy for chemical fate testing of new chemicals is presented in Figure 2. The extent of testing required is here related to produced t9nnages as a first criterion. Thus, if the annual or the accumulated production exceed certain quantities, requirements for testing at a higher level in the test hierarchy is automatically triggered. Additional criteria are special concerns related to use and disposal patterns and of course to the properties of the chemical - both environmental fate properties and other properties, such as toxicity or suspected carcinogenicity.
EEC STRATEGY FOR ENVIRONMENTAL FATE TESTING OF NEW CHEMICALS
Level a (> 1 t/yr)
Physical chemical parameters
Screening tests for ready biodegradability
Levell (> 100 t/yr)
Bioaccumulation in fish
Further studies on biodegradability (inherent biodegradability tests)
Level 2 (> 1000 t/yr)
Further studies on accumulation, degradation and mobility (simulation tests)
Figure 2. Simplified outline of the EEC strategy for testing chemical fate properties of new chemicals.
A quite analogous test strategy can be used for effluents, although the test methods and the relative emphasis of criteria for requiring more extensive testing may differ. Thus, it is obvious that in relation to the problem of discharging perSistent chemicals, the discharged quanti ties should be an important criterion for deciding the extent of testing.
3. TEST METHODS FOR CHEMICALS Test methods for chemicals in use in Europe inClude in particu
lar the EEC legislative methods (2) inCluded in Annexes VII and VIII to the Council Directive 79/83l/EEC (the so-called 6th ammendment to the 67-Directive on claSSification, packaging and labelling of dangerous substances). The EEC methods are presented as rigorous test protocols, and most of them are based on similar but more 100sly drafted OECO guidelines (3). The methods have been adopted for notification of new chemicals. It needs mentioning that a major goal of the EEC legislation and of the EEC and OECO harmonization of test
- 259-
methods and test strategies has been the prevention of technical trade barriers for chemicals. The resulting test strategy, for biodegradation in particular, presents no doubt an important innovation, but the test methods on which the strategy is based are unfortunately not yet harmonized to an extent that is satisfactory from a technical/scientific viewpoint. Most test methods seem to have been adopted for historic reasons and are slight modifications of previous national methods.
Screening tests for ready biodegradability. Characteristics of the various EEC level 0 screening tests are presented in Table 1, which includes for comparison also the Zahn Wellens test for inherent biodegradability and some newly proposed seawater tests (6, 10). The different level 0 tests listed are all intended as models for investigating aerobic degradability in freshwater and sewage treatment plants, but they differ greatly with respect to their biodegradation potential, owing to large differences in the quantities of microbial inoculum used, and to differences in the applied concentrat:i,on of test material, and for some tests also owing to a large freedom with regard to the source of inoculum. (In one test, even the use of soil extract as inoculum is allowed).
Table 1. List of the current EEC level 0 screening tests for ready biodegradability in order of increasing general biodegradation potential, and for comparison the Zahn Wellens test for inherent biodegradability and three newly proposed seawater biodegradation tests. Estimated cell densities are from (6-9).
Inoculum Test Approx. cell substance
density cone. Analytical
Name of test Source (viable counts/ml) (mgjl) Earameter
Closed bottle secondary effluent 0.25 102 2-10 O2
MITI sludge grown on (0.2-1) 106 100 CO and peptone/glucose or 30 mg SS/l fifial DOC
Modified OIlCD secondary effluent 0.5-2.5 102 10-80 DOC river water or soil extract
Modified AFNOR secondary effluent (5:!:.3 ) 105 80 DOC (filtered and re-suspended)
Sturm sludge supernatant 104 _ 105 5-20 CO2
UK-MITI sludge (0.2-1) 106 100 CO2 and (30 mg SS/l) final DOC
Zahn Wellens sludge (0.6-3) 107 100-800 DOC or COD (1000 mg SS/l)
Seawater - seawater 104 _105 10-80 DOC or shake flask or 0.1-1 specific
analysis
Seawater - seawater 104_10 5 2-10 O2 or closed bottle or 0.1-1 specific
analysis
Seawater seawater 104 _105 10-3 -0.1 14-C-simulation activity
- 260-
Interlaboratory comparison programmes or ring-tests of the various screening methods, organized by the EEC. ( 4) and the OECO (1), did in fact reveal an unanticipated large scatter in results, demonstrating the decisive role of the character and quantity of inoculum used. Subsequent ring-tests of methods with less variability allowed in the inoculum have yielded more consistent results (5, 6), although same chemicals still behaved erratically, as they passed the tests in only same laboratories. One model chemical, 4-nitrophenol has shown a particular aberrant behaviour in screening tests, see e.g. Nyholm et al. (7).
The EEC level 0 screening tests (Table 1) are die-away tests of a maximal duration of normally 28 days, and are performed with the test material added as the only carbon source to a natural consortium of aerobic microorganisms (the inoculum) not preexposed to the test substance. The following unspecific summary parameters are used to measure the degradation: DOC (dissolved organic carbon), O2 ( dissolved oxygen), or evolved CO • The applied concentration of test material is chosen for anal;tical reasons or to satisfy test technical needs, and range fram 2-100 mg/I.
Following the test strategy described above, it is possible to live with the variability between tests and laboratories, as a negative test result only implies that more testing should 'be done. However, the tests can easily be harmonized and improved to give more consistent results. Following recoounendations by the ECETOC (European Chemical Industry Ecology & Toxicology Centre, (8», the OECO has commenced work to update the tests, and the ISO is very active for the time being in standardizing test methods for biodegradability, which work is resulting in significantly improved and well described methods. (The first ISO standard was a DOC-based shake flask test (11». The most important change suggested by the ECETOC is a general prescription of the use of pre-conditioned activated sludge as the inoculum in concentrations of either 15-30 mg/l or 5-10 mg/I. Also the data interpretation, as practised today, might be improved towards a semiquantitative kinetic description, see Figure 3.
100
80
60
% removol
I 1
>-- tL -*-c.ISo---1 " \eg phos@"
I 60 days
II
- 261-
Figure 3. Interpretation of a screening test biodegradation curve.
The use of pre-adapted inocula (7, 12) is also under consideration by both the EEC and the ISO. This may be a way of circumventing the problem of using artificially high test substance concentrations for analytical reasons. There are problems, however, with the classification of such tests somewhere between the screening tests and the inherent tests (see below).
Inherent tests. There are two different standardized inherent tests in current use: the Zahn Wellens test and the SCAS test. The Zahn Wellens test (13) has been used particularly in the Federal Republic of Germany and is a simple batch test with prewashed activated sludge in a concentration of initially 1 gil of suspended solids and the test substance added in a relatively high concentration (50-400 mg DOC/I). The test has a high potential for biodegradation due to the high biomass and the batch-wise mode of operation. It should be emphasized that carbon substrates other than the test substance is continuously being released from the sludge in small concentrations, which fact probably also favours biodegradation.
The SCAS test (Semicontinuous Activated Sludge Test) is based upon a different principle, as it, as the name suggests, is operated semicontinuously with a I-day fill and draw cycle, which inVOlves fresh inoculation with sewage each day, and a period of substrate sufficient conditions followed by a period of sludge starvation with the test substance left as the unused substrate residue. The test may be run for several months and offers great opportunity for adaptation and consequently for degradation.
Simulation tests. The only simulation test officially adopted so far is the Activated Sludge Simulation Test (15), which is a modification of the OECO confirmatory test (16) used for detergents for a number of years. The test is carried out at room temperature with a laboratory scale activated sludge unit fed with synthetic (peptone) sewage, to which the test substance is added in a relatively high concentration. The unit is operated at a hydraulic retention time of 3 or 6 hours and a 6-days sludge retention time (or about that). The method has its limitations - data are obtained for only one set of operational conditions; synthetic and not real sewage is used; and the concentration of test chemical is high, at least compared to the micropollutant levels normally found in municipal sewage.
Nevertheless has the test proved useful in practice, especially for anionic detergents (which, however, enter sewage treatment plants in ppm concentrations).
For degradation in surface water, no official EEC/OECO simulation method has been adopted, but die-away tests with natural waters and low concentrations (1l9/1 level) of l4-C-labelled substances have been in use for some time and can easily be standardized as official methods. Usually Simple 1st order degradation kinetics prevail over extended periods of time, which make elaboration of the kinetics an easy matter (17). Testing of actual low environmental concentrations to estimate (hopefully) "real world" degradation rates has especially been advocated by Martin Alexander in a number of publications, for review see (18).
Tests for degradation in aerobic sediments using suspended sediments or sediment cores have also been described, and there is an increasing interest in investigating biodegradation under anoxic and anaerobic conditions in groundwater, in sediments, and in sewage treatment plants.
It will be a challenge to work out general test principles and protocols for harmonized and technically feasible tests that can
- 262-
make it possible to transfer methods from use at the specialized research level to more general use in environmental management.
4. TEST METHODS FOR EFFLUENTS The state of the art for testing complex effluents obviously
lags behind that for individual chemicals. Traditionally, studies have been restricted to measurements of BOD/OOD ratios and chemical analyses for certain compounds. For investigating treatability, specific studies have been made frequently using only 000 and BOD reductions as quality criteria. However, the test strategies and the test methods developed for chemicals are partly being transferred (in modified form) also to effluents, at least in some countries.
For examination of effluents discharged to surface waters, a stepwise strategy containing some of the following elements might be feasible:
1) Screening of biodegradability by means of a long term BOD test of duration 28 days, carried out in a manometric BOD apparatus with 30 mg sludge/l (UK-MITI test).
2) Shake flask die-away test in natural water using DOC measurements to follow the degradation, and, if relevant and technically feasible, supplemented with measurements of TOX and specific compounds (start and final concentrations).
3) Stabilization study, which involves incubation of the effluent with surface water under aerobic conditions (aeration) for up to several months, and characterization of the residue by biotests and a~propriate chemical analyses. The course of degradation can be followed by DOC analyses. By selecting an appropriate (low) dilution factor of the effluent in surface water (e.g. 1:3), a residue can be obtained which is sufficiently concentrated for biotests and chemical analyses to be carried out. The dilution must be great enough to prevent severe toxic effects, though. Because toxic effects can seldomly be ruled out entirely, the degree of conversion (DOC reduction) should be compared to the extent of degradation observed in more dilute systems, such as the shake flask study described above. The method has been used by the water Quality Institute for about six years on a number of effluents. An example of a study described in a publicly available report is (19).
4) Simulation tests of individual concern compounds for estimation of kinetic rates and for investigating biodegradability under environmentally realistic conditions. Concern compounds may be those compounds identified in the residue after stabilization which are present in significant quantities, or which have particular undesirable properties (toxic, bioaccumulable, etc.). The simulation tests selected should be either surface water or sediment tests or both, depending on the environmental partitioning of the compound in the receiving environment. For an evaluation of the partitioning of the substances, relevant physical/chemical properties (such as the K ) must be known. If not available in the literature, however, these~roperties can often be estimated with sufficient accuracy from chemical structures.
If resources are available, tests on the effluent should be complemented with chemical monitoring in the near field around the discharge. A monitoring progr8llllle may comprise: 1) analyses of cri tical compounds in caged mussels or fish exposed at various distances from the outlet, as well as, of course, 2) analyses of surface sediments. A method of the future is perhaps to use physical/chemical concentrators (e. g. ion exchange resins) for estimation of integrated
- 263-
mass fluxes. For effluents discharged to treatment plants, test methods are
in one way more developed, as various treatability tests have been in use for a number of years. As has been pointed out, however, the traditional characterization by largely BOD and COD must be extended to include appropriate characterization of the residue (including sometimes even simulation biodegradability tests for individual compounds) in order to satisfy the environmental management needs of today.
REFERENCES
( 1 ) OECO, Environmental Committee. Chemicals group. ( 1981 ) • OECO Chemicals Testing Progranune. Expert group degradation/accumulation. Final report. Umwelt Bundesamt, Bundesrepublik Deutschland. Government of Japan. Berlin & Tokyo.
(2) Offical Journal of the European Conununities No. L 251, Vol. 27, 19. September 1984. Commission Directive of 25 April 1984 adopting to technical progress for the sixth time Council Directive 67/548/EEC on the apparoximation of laws, regulations and administrative provisions relating to the classification, packaging and labelling of dangerous substances.
(3) OECO (1981). OECO's guidelines for testing of chemicals. OECO, Paris.
( 4 ) EEC (1979). Intercomparison Progranune on Detecting BiodegradabiIi ty of Chemicals in Water. CEC Environmental and Consumer Protection Service, prepared by Gesellschaft fUr Strahlen- und Umweltforschung MBH, MUnich. (H. Rohleder).
( 5 ) EEC (1985). Conunission of the European Conununi ty, DegradationaccumUlation subgroup. Ring Test Progranune 1983-84. Assessment of biodegradability of chemicals in water by manometric respirometry. DG XI -283/82, Rev. 5. (Painter, H.A. and King, A.F.).
( 6 ) EEC (1987). Conunission of the European Conununi ties. DegradationaccumUlation subgroup. Ring Test Progranune 1984-1985 .. Screening test methods for assessment of biodegradability of chemical substances in sea water. Final report, prepared by the Water Quality Institute, Denmark. (Nyholm, N. and Kristensen, P.).
(7) NYHOLM, N., LINDGAARD-J0RGENSEN, P. and HANSEN, N. (1984). Biodegradation of 4-ni tropheno1 in standardized aquatic degradation tests. Ecotoxicoloty and Environmental Safety 8, 451-470.
(8) BLOK, J., DE MORSIER, A., GERlKE, P., REYNOLDS, L. and WELLENS, H. (1985). Harmonization of ready biodegradability tests. Chenosphere 14, 1805-1820.
(9) PAINTER, H.A. and KING, E.F. (1980). A mathematical model of biodegradabili ty screening tests as an aid of interpretation of observed results. Proceedings from the EEC symposium on Environment and Quality of Life. Sophia, Antipolis Valbonne, France.
(10) DAMBORG, A. and NYHOLM, N. (1986). Proposed method for determining the degradation of chemical substances in relatively low concentrations in surface water ("simulation test method"). Working document submitted to the European Communi ties. Degradation-accumulation subgroup. (DOC/87/263/XI).,
- 264-
(11) International Standard. ISO 7827. water quality - Evaluation in an aqueous medium of the "ultimate" aerobic biodegradability of organic compounds - Method by analysis of dissolved organic carbon (DOC). First edition 1984.10.15.
(12) ISO/TC 147/SC 5/WG 4. N97. PAINTER, H.A. (1987). Degradation with adapted inocula: A survey. water Research Centre, U.K.
(13) ZAHN, R. and WELLENS, H. (1974). Ein einfaches Verfahren zur PrUfung der biologischen Abbaubarkeit von Produktion und Abwasserinhaltstoffen. Chern. ztg. 98, 228-232.
(14) American Soap and Detergent Association (1965). Biodegradation Subccmni ttee. Procedure and standards for. determination of biodegradability of ASS and LAS. JAOCS 42, 986-993.
(15) EEC Directive 79/831 Annex V (OG XI/605/82) Part C: Methods for determination of ecotoxicity - LevelL C(L1 )8: Biodegradation. Activated sludge simulation tests.
(16) OECO (1976). Pollution by detergents. OECO, Paris. (17) NYHOLM, N. and DAMBORG, ~. (1987). Biodegradation kinetics in
aqueous systems - An overview. Commission of the European Communities. Water pollution research report 3. Behaviour and transformation of organic pollutants in groundwater treatment. EUR 11094.
( 18) ALEXANDER, M. ( 1985) • Biodegradation of organic chemicals. Environ. Sci. Technol. 18, 106-111.
( 19) Naturvardsverket (Swedish Environmental Protection Board). SNV Rer-rt 3037. Undersokninger av bionedbrytbarhet av industriavloppsvatten fran Stenungsund (April 1985). Prepared by the Water Quality Institute, Denmark. (LINDGAARD-J(llRGENSEN, P. , NYHOLM, N., KUSK, K.O. and LUND, U.).
- 265-
MASS FLUXES OF LINEAR ALKYLBENZENESULPHONA TES, NONYLPHENOL, NONYLPHENOL MONO- AND DIETHOXYLA TE
THROUGH A SEWAGE TREATMENT PLANT
A. MARCOMINI*, S. CAPRI**, P.H. BRUNNER*** and W. GIGER***
* Department of Environmental Sciences, University of Venice, 1-30123 Venezia, Italia
** Water Research Institute (lRSA), 1-00198 Roma, Italia *** Swiss Federal Institute for Water Resources and Water Pollution Control
(EA WAG), CH-8600 Diibendorf, Schweiz
Summary Dissolved and particulate LAS, NP, NIPEO and NP2EO were determined in raw sewage, primary effluent, secondary effluent, primary and secondary sludge of the sewage treatment plant Ziirich-Glatt on two consecutive sampling days. The results were .combined with sewage flow rate and dry matter load measurements in order to obtain the respective mass fluxes. The latter provided information on the major removal processes affecting these detergent derived chemicals, i. e. biodegradation during the activated sludge treatment and incorporation into sludge. The overall removals of LAS over the two sampling days were 99.3% and 98.7% (w/w), of which 13 % and 16%, respectively, were found in raw sludge and 9% and 13% ,respectively, in anaerobically digested sludge. Both NP1EO and NP2EO were poorly removed in the secondary effluent and were educts for NP in sludge. Nonylphenol loads in the anaerobically digested sludge were 4.7 and 6.4 times larger than in the raw sewage.
1. INTRODUCflON Linear alkylbenzenesulphonates (LAS) and alkylphenol polyethoxylates (APnEO) are the most important aromatic synthetic surfactants. LAS are mainly contained in household detergents, whereas APEO are mainly employed in industrial and institutional detergent products [1]. Both are complex mixtures of homologs, isomers and oligomers; the APEO most widely used in Western Europe are the nonylphenol polyethoxylates (NPnEO, n=I-20). Because of their direct use in water, these surfactants are discharged into municipal and industrial wastewaters and enter sewage treatment plants (STP) at mgIL concentration levels. Since the introduction of LAS in the late sixties to replace the less biodegradable, highly branched alkylbenzenesulphonates (ABS), many investigations on the occurrence and the biodegradation behaviour of LAS in influents and effluents of STP, as well as in natural waters, were carried out employing predominantly non-specific methylene-blue based procedures [2] . Selective methods, determining LAS by gas-chromatography after adequate alteration or derivatisation were only rarely used [3,4]. The accumulation of LAS as a result of their sorptive properties has been only recently systematically examined [5-7], although concentrations of 0.8 - 1.3 % dry matter of anionic surfactants in sludges were reported already in 1966 [8]. In contrast to LAS which are more persistent and toxic than their biodegradation products, the phenolic nonionic surfactants of the NPnEO type have been found to be rather easily transformed to more persistent metabolites during sewage and sewage sludge treatment, namely nonylphenoxy acids [9], nonylpolyethoxylates with one (NPIEO) or two (NP2EO) ethoxy units and nonylphenol [10]. The latter accumulate in aerobically and anaerobically stabilized sludge
- 266-
to an average of 1.3 and 0.3% dry matter, respectively [11]. Few systematic studies on the behaviour of LAS and APEO during wastewater and sewage treatment, other than the examination of the raw sewage (influent) and of the treated wastewater (effluent), were so far published [12-15]. This islargely due to the lack of suitable analytical methods for routine determinations of these chemicals. Following the development of HPLC based analytical procedures capable to simultaneously enrich and determine from aqueous and solid matrices both LAS and NP, NPnEO, we report here on an investigation aiming at mass-balancing these chemicals through a STP.
2. EXPERTIMENTAL 2.1 Sewage treatment plant (STP) The Zurich-Glatt STP has a population equivalent capacity of 240'000, with a sewage flow rate of 55-60'000 m3 /day. After mechanical treatment of the raw wastewater (influent) which is of mixed, residential and industrial origin, the primary effluent undergoes an activated sludge treatment. The effluent from the secondary claryfier is discharged into the receiving water which is the Glatt River. Primary and secondary sludges are combined and treated by anaerobic digestion.
2.2 Sampling Twenty-four hours composite samples of raw sewage, primary and secondary effluents and of activated sludge were collected on May 13 and 14, 1986, with automatic, flow proportional, sampling devices. Raw and digested sewage sludges were taken from open ducts during sludge transfer [16]. Sludge and wastewater samples were approximately 2.5 L. Immediately after collection the samples were preserved by the addition of 1 % (v/v) formaldehyde. The samples were stored either as received in the dark at 4°C, or as follows: After vigorous shaking, aliquots of 500 mL were dried at 60 °C to a constant weight to determine the "dry matter content" (d.m.). The dry residue was pulverised to a particle size smaller than 300 11m and stored at 4°C. The particulate matter ("suspended solids") was determined by fIltration through 111m fiber fIlters.
Table I: Analytical methods
Compound Sample
LAS influents
NP NP1EO NP2EO
effluents sludge liquors
sludges
influents effluents sludge liquors
sludges
Enrichment SeparationnDetection
Cls-SiCh reversed-phase HPLC/ cartridges UV -fluorescence
Soxhlet
Cls-SiCh normal-phase HPLC/ cartridges UV -fluorescence
Soxhlet
HPLC: high-performance liquid chromatography; UV: ultraviolet.
- 267-
Ref.
16
15
10,16
10,15
2.3 Analytical methods The analytical methods used for the specific determination of LAS, NP, NP1EO and NP2EO are described in detail elsewhere. An overview and reference are given in Table I.
3. RESULTS AND DISCUSSION The procedure aiming at mass-balancing consisted in determining: a) sewage flow rates and loads of particulates (dry matter); b)concentrations of LAS, NP, NPlEO and NP2EO in filtered water and particulates Combination of a) and b) provided the mass-fluxes of the examined pollutants. Such an approach to the study of individual chemicals through the STP necessitates concerted and composite sampling of influent, effluent from primary sedimentation, activated sludge and effluent from secondary clarification, as well as influents and effluents from sewage sludge digestion. However, the sludge residence time in the digester (ca. 20 days) is much longer than in the primary clarifier which does not allow a direct comparison between samples of raw sludge taken on two sampling days with those of digested sludge. The concentrations of dissolved LAS, NP, NPIEO and NP2EO were determined for all matrices; additionally, the concentrations of the same compounds on particulate were determined for influent and activated sludge, whereas for the sludges the concentrations into the dry matter was measured. The results are reported in the Table II and III. The fluxes were obtained according to the following equation:
Q F = cd'Q + cp·d.m: 1000
where F is the mass-flux (kg/day), Cd (kg/m3) is the concentration in the dissolved phase; Q is the sewage flow rate (m3/day); cp (g/kg d.m.) is the concentration on particulate and d.m. is the dry matter (kg/m3). The mass-fluxes are shown in the Figs. 1-5. Accuracy and precision of the data obtained from the chemical analysis were reported elsewhere [10,15,16] in each of the examined matrix by spiking and reproducibility experiments. Percentage recovery and relative standard deviation values were in the ranges 88-102% and 2-8%, respectively. Due to their chemical-physical properties, biodegradation and adsorption are the only processes affecting LAS, NP, NP1EO and NP2EO during the wastewater and sewage treatment. The aqueous solubility of commercial LAS mixtures, below their critical micell concentration, varies from 30 to 100 mg/L [18], depending on the exact average alkyl chain length and on the phenyl isomer composition. Adsorption partition coefficent to bacteria, measured with activated sludge solids, was 3'000 mUg [18]; octanol-water partition coefficients, Kow of 4-500 [19] and 91 [20] were reported. Aqueous solubilities of NP, NPlEO and NP2EO were determined as 5.2,2.9 and 3.1 mg/L, respectively [21]; sorption constants for different activated sludges were reported to be 10'000-26'000 ml/g for NP, 3'900-11'000 ml/g for NPIEO and 1'300-6'900 ml/g for NP2EO [21]. A logKow value of 4.2 for NP was determined by HPLC [22]. The concentration ratios dissolved to particulate deducible from Tables II and III show that at low suspended solids concentrations of 1 to 100 mg/L (e.g. raw sewage, primary and secondary effluent), more than 90 % LAS are dissolved in water. When the particulate concentration increases to 1'000-2'000 mg/L, due to biomass production and flocculation during activated sludge treatment, the fraction of dissolved LAS decreases to 50% and at suspended solid concentrations above 10'000 (e.g. raw and digested sludge), more than 90% of the LAS are concentrated on the particulate matter. According to its lower solubilities and higher Kow, partitioning of NP between sewage and sludge, compared with that of LAS, is shifted toward lower values of suspended solid concentrations, which causes a stronger accumulation of NP on particulate matter. The removals through sorption and biodegradation are effectively visualized by the massfluxes (Figs. 1-5). With regard to the dry matter (Fig. 1), despite the higher amounts of
- 268-
particulate in the influent of May 13, the amounts removed by raw and digested sludge and those released through the secondary effluent were remarkably similar for both sampling days. Influent LAS are efficiently removed by the STP to the receiving waters to an extent of 98.7% (May 13] and 99.3% [May 14] (w/w) (Fig. 2). This is in agreement with previous investigation employing selective methods for determining LAS in influents and effluents [12]. From the mass balances, the most important process determining the fate of LAS in sewage treatment is biodegradation by the activated sludge process, which was responsible of 86% and 83% (w/w) removal on May 13 and 14, respectively. The second most relevant transfer path for LAS is the incorporation into sewage sludge. Of the influent LAS, 13% (May 13) and 16% (May 14) (w/w) were transferrred to the raw sludge. During the anaerobic mesophilic sludge digestion 4% (May 13) and 3% (May 14) (w/w) were eliminated, which means that 31 % (May 13) and 18% (May 14) of the LAS in raw sludge were metabolized during sludge stabilisation. As above mentioned, however, the last two figures have to be considered with caution because of the longer sludge residence time in the digester than in the primary clarifier. Although return activated sludge was not analyzed, its contribution to the accumulation of LAS in the raw sludge can be neglected because of the removal efficiency of both adsorbed and dissolved LAS during the activated sludge treatment (Tables II, III and ref. 23). Contrary to LAS, good closures on the mass balances could not be obtained for NP, NPlEO and NP2EO due to their feature of metabolic refractory products. Figure 3 confirms the hypothesis that, during anaerobic sludge treatment, 4-nonylphenol is being accumulated in sewage sludge [14,24], and that this NP is quantitatively (99%) associated with sludge particulate matter. Compared with the input load, the NP enrichment factor in anaerobically digested sludge was 4.7 on May 13 and 6.4 on May 14. Figures 4 and5 show an overall poor removal from the STP via secondary effluent of NP1EO and NP2EO, according to the results of Giger et al. [14]. Both are partially degraded during aerobic and anaerobic sewage and sludge treatment and are educts for NP in sludge. The fact that the load of NPlEO and NP2EO in the raw sludge can be higher than in the raw sewage infers that NPnEO (n=3-20) supply NPlEO and NP2EO when they are degraded during aerobic wastewater treatment. Nonylphenolcarboxylates, which are also formed by the degradation of NPnEO [9] and which can be further transformed to NP, were not determined in this work.
4. CONCLUSION The approach to the study of LAS, NP, NPlEO, NP2EO through STP based on massfluxes provided quantitative informations on distribution of these pollutants between sewage and sludge and on removals from the individual stages of the treatment. The aromatic surfactants LAS and NPnEO are not fully degraded during sewage treatment. Approximately 1 % (w/w) of the LAS entering the plant left in the secondary effluent. Over the two sampling days, 13% and 16% (w/w), respectively, were transferred to raw sludge and only partly degraded during anaerobic sludge treatment. Nonylphenol mono- and diethoxylate are supplied by higher NPnEO during aerobic wastewater treatment and are educts for NP especially in anaerobically digested sludge. Partitioning of LAS and NP is controlled by the suspended solids concentration; according to its higher lipophilicity, NP accumulated more strongly on particulate matter. Based on the literature and on this work, it would be anticipated that the amounts of LAS leaving the STP adsorbed to sludges would only slowly biodegrade as long as anaerobic conditions prevail.
- 269-
REFERENCES [1] Layman, P.L., Chern. Eng. News, Jan. 20,21 (1986). [2J Swisher, RD., "Surfactants Biodegradation", 2nd ed., M. Dekker, New York,
1987,751 and references therein. [3J Waters, J. and Garrigan, J.T., Water Res., 17, 1549 (1983). [4] Osburn, Q.W., J. Am. Oil Chern. Soc., 63,257 (1986). [5] McEvoy, J. and Giger, W., Environ. Sci. Technol., 20, 376 (1986). [6] Motschi, H. and McEvoy, J., Naturwissenschaften, 72, 654 (1985). [7J De Henau, H., Mathijs, E. and Hopping, W.D., Int. J. Environ. Anal. Chern.,
26, 279 (1986). [8] Bruce, A.M., Swanwick, J.D. and Ownsworth, R.A., J. Proc. Inst.Purif., 427
(1966). [9] Ahel, M., Conrad, T. and Giger, W., Environ. Sci. Technol., 21, 697 (1987).
[10] Ahel, M. and Giger, W., Anal. Chern., 57, 1577 (1985). [11] Giger, W., Brunner, P.H. and Schaffner, C., Science, 225, 623 (1984). [12] Sedlak, RI. and Booman, K.A., Chim. Oggi,9,21 (1986). [13J Brown, D., de Henau, H., Garrigan, J.T., Gerike, P., Holt, M., Kunkel, E.,
Mathijs, E., Waters, J. and Watkinson, R.J., Tenside, 1,24 (1987). [14] Giger, W., Ahel, M. and Koch, M., Wat. Sci. Tech.,19, 449 (1987). [15] Marcomini, A. and Giger, W., Anal. Chern., 59,1709 (1987). [16] Marcomini, A., Capri, S. and Giger, W., J. Chromatogr., 403, 243 (1987). [17] CEC, Commission of the European Communities: COST Project 68 Sewage
Sludge Processing, Final Report of the Management Committee, Brussels, Belgium (1985).
[18] Games, L.M., "Field validation of exposure analysis modeling system(EXAMS) in a flowing stream", in "Modeling the Fate of Chemicals in the Aquatic Environment", K.L. Dickson, A.W. Maki and J. Cairns, Jr., Eds., Ann Arbor, Michigan, 1982, pp.325-346.
[19] Kimerle, R.A., Swisher, RD. and Schroeder-Comotto, RM., "Structureactivity correlations in studies of toxicity and bioconcentration with aquatic organism", G.D. Veith and D.E. Konasewich Eds., U.C.R.A.B., Burlington, 1975, pp.2-55.
[20] Holysh, M., Paterson, S., Mackay, D. and Bandurraga, M.M., Chemosphere, 1,3 (1986).
[21] Ahel, M., Giger, W. and Koch, M., Proceedings of the 4th European Symposium on "Organic MicropoIlutants in the Aquatic Environment", A. Bjoerseth and G. Angeletti Eds., D. Reidel, Dordrecht, 1985, pp. 414-428.
[22] McLeese, D.W., Zitko, V., Sergeant, D.B., Burridge, L. and Metcalfe, C.D., Chemosphere, 10, 723 (1981).
[23] Yoshimura, K. and Nakae, A., Jap. J. Wat. PoIlut. Res., 63, 5 (1982). [24] Tschui, M. and Brunner, P.H., Vom Wasser, 65,9 (1985).
- 270-
;:l
Tab
le I
I C
once
ntra
tion
s o
f LA
S, N
P,N
PlE
O a
nd N
P2E
O in
sew
age
and
slud
ge f
rom
the
sew
age
trea
tmen
t pla
nt Z
Uri
ch-G
latt
o
n M
ay 1
3,19
86.
Q
dry
LAS
NP
N
PIE
O
NP
2EO
m
atte
r
tota
l pa
rt
diss
ol
tota
l pa
rt
diss
ol
tota
l pa
rt
diss
ol
tota
l pa
rt.
diss
ol
m3/
d W
L Il
g/L
m
glkg
Il
g/L
Il
g/L
m
glkg
Il
g/L
Il
g/L
m
glkg
Il
g/L
Il
g/L
m
glkg
Il
g/L
raw
sew
age
5550
0 0.
58
45
10
20
8 42
50
50
8.
8 24
75
17
33
65
14
21
effl
uent
from
pr
imar
y 55
500
0.46
3
86
0
134
3800
39
6.
7 15
71
20
24
55
10
19
cl
arif
icat
ion
activ
ated
n.
d.
2.3
640
147
290
88
81
1 30
0 31
0 6.
5 70
52
3.
5 sl
udge
effl
uent
from
se
cond
ary
5510
0 0.
43
60
b.d.
60
5.
6 b.
d.
2.5
41
b.d.
17
31
b.
d.
13
clar
ific
atio
n
raw
slu
dge
550
13
58
50
0
41
80
42
00
2990
23
0 5
4290
33
0 16
36
40
280
15
dige
sted
slu
dge
100
48 ~2
1 00
0 4
55
0
2700
75
360
1570
9.
5 67
20
140
1.1
2400
5
0
0.1
Q =
55
500
m3 /
d ;
num
bers
in it
alic
wer
e ca
lcul
ated
; b.
d. :
bel
ow d
etec
tion
lim
it;
n.d
: n
ot d
eten
nine
d.
I ij
I
Tab
le ill
Con
cent
rati
ons
of L
AS
, NP
,NP
IEO
and
NP
2EO
in s
ewag
e an
d sl
udge
fro
m t
he s
ewag
e tr
eatm
ent p
lant
ZU
rich
-Gla
tt
on M
ay 1
4,19
86.
Q
dry
LAS
NP
N
PIE
O
NP
2EO
m
atte
r
tota
l pa
rt
diss
ol
tota
l pa
rt
diss
ol
tota
l pa
rt
diss
ol
tota
l pa
rt.
m3/
d gI
L
IlgI
L
mgl
kg
Ilg/
L
Ilg/
L
mgl
kg
Ilg/
L
Ilg/
L
mgl
kg
Ilg/
L
IlgI
L
mgl
kg
raw
sew
age,
5
69
00
0.
46
30
30
68
3
00
0
21
2.3
20
30
8.
3 26
16
2.
3
effl
uent
from
pr
imar
y 5
69
00
0.
43
35
20
39
3
50
0
15
2.9
14
30
9.6
26
17
4.
8 cl
arif
icat
ion
activ
ated
n.
d.
1.9
490
133
230
150
74
0.8
530
300
5 11
0 44
sl
udge
effl
uent
from
se
cond
ary
56
80
0
0.37
18
b.
d.
18
2.7
b.d.
2.
7 19
b.
d.
19
15
b.d.
cl
arif
icat
ion
raw
slu
dge
480
15
57
00
0
35
50
3
80
0
28
50
19
0 3
39
00
26
0 13
3
50
0
230
dige
sted
slu
dge
100
52
2300
00
43
50
2
70
0
78
00
0
15
00
9
67
00
13
0 0.
65
18
20
35
Q =
55
500
m3 /
d ;
num
bers
in it
alic
wer
e ca
lcul
ated
; b.
d. :
bel
ow d
etec
tion
lim
it;
n.d.
: n
ot d
eter
min
ed.
diss
ol
I
Ilg/
L 15
15
2.8 15
12
0.0
4
Figure 1 Mass fluxes of dry matter ("total solids") through primary clarification (PC), activated sludge treatment (AT), secondary sludge treatment (ST) and anaerobic digestion (AD) in the sewage treatment plant Ziirich-Glatt. Sampling days: May 13 (A) and May 14 (B) 1986.
STP GLATT May 13 th 86
A
32 100%
DRY MATTER FLUX in 103 kg day-I
2 3°10
f.D part icu late d. m.
c:::J dissolved d. m.
1104 kg day
23 purified 71% sewage
biodegradation
I~==.~ms!l!lmimlm~ ~O% sewage sludge
2 6%
anaerobic degradalion
STP GLATT May 14th 86
B
26 «::::::} 100% ««:: raw sewage
DRY MATTER FLUX in 103 kg da(1
,---- ? ------
?
tm\iI particulate d.m,
1:-:·:·:-:1 dissolved d. m.
1104 kg day
21 puri fied 81% sewage
~:jm!i!Iilmmmll§D ~% sewage sludge
2 7%
anaerobic degradation
- 273-
Figure 2. Mass fluxes of LAS through primary clarification (PC), activated sludge treatment (An, secondary sludge treatment (Sn and anaerobic digestion (AD) in the sewage treatment plant Ztirich-GlattSampling days: May 13 (A) and May 14 (B) 1986.
A
245 100%
B
LAS - FLUX in kg day-l
~---?----..
210 86%
22 1~=l_="':===<mD 9%
10 4%
LAS - FLUX in kg da(1
~---.? - - -"
5 3%
- 274-
Figure 3 Mass fluxes of NP through primary clarification (PC), activated sludge treatment (AT), secondary sludge treatment (ST) and anaerobic digestion (AD) in the sewage treatment plant ZUrich-Glatt.Sampling days: May 13 (A) and May 14 (B) 1986.
A
NP - FLUX in kg da(l
12kg day-l
1.6 r:::::J~=---;;~:::;~~~-rr--=-=---Y 0.13 100%L;; 8%
?
B
?
NP - FLUX in kg da(l
- 275-
?
Figure 4 Mass fluxes of NP1EO through primary clarification (PC), activated sludge treatment (AT), secondary sludge treatment (ST) and anaerobic digestion (AD) in the sewage treatment plant ZUrich-Glatt.Sampling days: May 13 (A) and May 14 (B) 1986.
A
NP1EO - FL UX in kg day-l
"""""~--:-?~::""';;~:-:':""--:':""7":-7'=::::""';h-"::::::::"""~...,..,.. 0.94
?
B
1.76 lOa % I.ni.c<ii=~
?
NPI EO - FLUX in kg day-l
- 276 -
__ -\-,-,-'--'J 39 %
?
I lkg da(l
1.1 -+I----=:-=----+-'--'-~ 65 %
?
Figure 5 Mass fluxes of NP2EO through primary clarification (PC), activated sludge treat-ment (AT), secondary sludge treatment (ST) and anaerobic digestion (AD) in the sewage treatment plant ZUrich-Glatt. Sampling days: May 13 (A) and May 14 (B) 1986.
A
?
B
NP2 EO - FLUX in kg day- l
?
I I k g da(l
<..::.....::::p;=====~ 0.24 - 62%
NP2EO-FLUX in 9 day-l
0.93 F:7":-::-:--:-:-I"---~~~,...,if~-~~-+~~ 100%t.:.;.....-:..:..4-..
?
- 277-
?
sc
0.18 19%
0.90 96%
Summary
MUTAGENIC COMPOUNDS IN CHLORINATED WATERS
Bjarne Holmbom* and Leif Kronberg** Abo Akademi
*Laboratory of Forest Products Chemistry ** Department of Organic Chemistry
SF-20500 Turku/Abo, Finland
Chlorination of lignin and humic material in water produces the strong bacterical mutagen 3-chloro-4-(dichloromethyl)-5-hydroxy-2(5H)-furanone or its acyclic form Z-2-chloro-3-(dichloromethyl)-4-oxo-butenoic acid ("MX"), and its geometric E-isomer ("E-MX"). Although MX is a trace component in the mixture of chlorination products, it is the main single contributor to the Ames mutagenicity in wood pulp chlorination liquors as well as in chlorinated drinking waters. Higher mutagenicity and higher amounts of MX and E-MX are produced during chlorination under acidic conditions than under neutral conditions. Substitution of chlorine by chlorine dioxide decreases the formation of mutagens. Although E-MX has a much lower mutagenicity than MX the compound is of concern since it can be isomerised to MX. This conversion occurs particularly under acidic conditions. MX reacts with sulfite and sulfide. Sulfite treatment may be a feasible process to remove chlorinated mutagens from both bleaching waters and drinking waters.
1. INTRODUCTION It has been known for about ten years that chlorination stage liquors
of wood pulp bleaching and chlorinated surface waters are mutagenic in the Ames assay (1-3). Much research work has been devoted to find the active mutagens formed during chlorination of lignin and humic material. A number of chlorinated mutagens have been identified in drinking waters (4) and in pulp bleaching liquors (5,6) but these compounds account for only a small part of the total mutagenic activity (4).
An extremely strong mutagen was isolated from pulp bleaching liquors in 1980 (7), and was identified as 3-chloro-4-(dichloromethyl )-5-hydroxy-2(5H)-furanone (Fig. 1). This new compound, coded MX, was recently synthesised (8), and its structure as well as its extreme mutagenicity was confirmed.
Studies on chlorinated humic and drinking waters in Finland indicated that the mutagenic activity is due mainly to nonvolatile organic acids of intermediate polarity (9-11).
Recently we succeeded in detecting the MX compound both in chlorinated humic waters and in drinking waters (12). Al though the concentration of MX in drinking waters was only at the level of 20-50 ng/L the compound contributed to a substantial part of the Ames mutagenicity of the
- 278-
waters. The next interesting discovery was the geometric isomer of MX, E-2-chloro-3-dichloromethyl-4-oxo-butenoic acid, coded E-MX (Fig. 1). It was first isolated from chlorinated humic water (13) and subsequently detected also in drinking waters and pulp bleaching liquors (14).
This paper summarises results of recent research regarding the formation'and properties of this "supermutagen" and its geometric isomer.
"}~jo H G G 12 G 1 HGGI 2 GOOH - ~ - H - GHO GOOH - GHO GI
0
'- -.r= ./
'MX' 'E-M x'
Fig. 1. Tautomeric and isomeric forms of "MX".
2. EXPERIMENTAL Natural humic water (20 mg TOC/L) from a marsh lake in SW Finland was
treated with chlorine (20 mg/L) and chlorine dioxide and ozone at pH 7.0 as described previously (15). In an other experiment the humic lake water was treated with different doses of chlorine at pH 3.0 and 7.0.
Analysis of MX and E-MX in chlorinated waters was performed by GC/MS with selected ion monitoring (SIM) of methylated XAD-4/8 extracts (12). E-MX was determined using the specific ions produced by loss of Cl and OCH3'
Ames mutagenicity tests were carried out by the plate incorporation assay without metabolic activation with duplicate or triplicate plates and at three or more dose levels (9).
Total organically bound chlorine (TOCl) was determined by concentration on XAD-resin and chlorine determination by the PIXE-technique (16). Total organic halide (TOX) was determined using the Dohrmann DX 20 system with adsorption to activated carbon, pyrolysis and microcoloumetric titration.
The stability of MX and E-MX was determined by HPLC analysis of water solutions (20 ~g/mL) stored in the dark at various temperatures and pH conditions. The reaction of MX with sulfite and sulfide ions was studied in a similar manner.
3. RESULTS AND DISCUSSION
3.1. Occurrence of MX and E-MX in chlorinated waters Analysis of a number of chlorination stage liquors from bleaching of
softwood kraft pulp revealed the presence of MX and E-MX in all liquors, although in quite different concentrations, particularly for E-MX (Table I). MX gave a varying, but significant contribution to the mutagenicity of the bleaching liquor extracts.
Drinking waters from more than 30 municipalities in Finland have been surveyed regarding mutagenicity and presence of mutagens (17). Mutagenicity and presence of MX and E-MX was consistently found in waters produced from surface waters using disinfection with chlorine. Some typical results are given in Table II. MX accounted for a major proportion of the mutagenic activity.
MX has now been detected in chlorinated drinking waters also in the USA (18) and in the UK (19), and its substantial contribution to the Ames mutagenicity has been verified.
- 279-
Table I. Mutagenicity and concentrations of MX and E-MX in extracts of chlorination stage liquors from bleaching of softwood pulp in kraft mills.
Mill Net Revertants/mL MX E-MX TOCI MX Contribution To TA 100, -S9 )1g/L )1g/L mg/L Mutagenicitya TOCI
% 22m
A 1780 7.4 29.5 155 11 23 B 460 10.8 0.8 NAb 60 C 460 13.9 1.2 NA 78 D 770 11. 5 4.2 NA 38 EC 700 3.9 10.0 238 14 7
a Based on 5600 revertants/nmole for MX; b NA: not analyzed; c Extracted with ethyl acetate
Table II. Mutagenicity and concentrations of MX and E-MX in Finnish drinking waters prepared from surface waters employing postdisinfection with chlorine.
Drinking TOC Chlorine MutageniCity MX E-MX MX Contribution Water Dose TA 100, -S9 to Mutagenicity Plant mg/L mg/L Net Rev./L ng/L ng/L %
A 8.6 1.0 600 5 6 22 B 4.6 1.0 900 14 12 40 C 5.6 1.0 4100 46 45 29 D 5.3 1.3 3200 39 23 32 E 6.9 1.9 1300 14 11 28 F 6.9 2.2 2300 39 17 44
It is evident that MX and E-MX are ubiquitous reaction products from chlorination of lignin and humic substances and that MX is a key compound in the mutagenicity of chlorinated waters.
3.2. Formation of MX and E-MX Chlorination under acidic conditions produces more mutagenicity and
higher amounts of MX (Fig. 2). A higher chlorine dose does not appear to increase the amount of mutagens formed when chlorination is carried out at pH 7.
Substitution of chlorine by chlorine dioxide decreases the mutagenicity and the concentrations of MX and E-MX (Fig. 3). However, mutagenicity and MX was observed also when mere chlorine dioxide was applied. Pretreatment with ozone (10 mg/L) before chlorination (10 mg/L) decreased only slightly the concentration of mutagens (Fig. 3).
Although it is obvious that MX and E-MX are formed from a variety of humic substances and from lignin, there is still only little knowledge on the structures of the precursors involved and the mechanisms of formation. MX and E-MX can be produced during chlorination of various phenols (14) including the amino acid tyrosine (19), and MX is the major mutagen in these chlorination reaction mixtures.
- 280-
TA 100 NET REV . /L
20000
10000
pH 3 . 0
0.5 1 1.5 WIW CHLORINE : TOC
MX 430 720 880 E - MX 1090 2030 2220
pH 7 . 0
MX CONTR I BUTION (THEOR . )
190 200 700 320
ngll 190
Fig. 2. Effect of chlorination pH and chlorine dose on mutagenicity and concentrations of MX and E-MX in chlorinated humic lake water.
NET REV./L TA 100
10000
8000
6000
4000
2000
lAX E-MX TOX
10010
193
393 2 . 38
75125 50/50
100 47
55 4.2 2 . 34 1 . 63
MX CONTRIBUTION
25175 01100 ell' ICIO,. WT, % CI,
20 20 194 128 ng/L 1.3 243 219
0 . 98 0.98 1.50 1.44 mg/l
Fig. 3. Effect of increased proportion of chlorine dioxide and of preozonation on mutagenicity and concentrations of MX and E-MX in chlorinated humic lake water.
100
UNREACTED MX
50
SULFITE .pH7 . 0
o~ = .:::: g.:: _ 0 .......
" " ' 0, " ~} E - M X , , ' 0
1:0
1:1
0~ __ ~ ______ ~ ____ ~ ____ ' __ ~1~: ~5_
10mln 4 24h
TI M E
SULFIDE .p H 9 . 0
0 - - -0 __
0 0 _ , ...... 0 ,
'" .... \ ,
\ ° , \ - 0 ,
\ ,
- 0 \
\ {< 1:5
10min 4 24h
TI ME
Fig. 4. Re~tion of MX with various molar ratios of sulfite (HS03-' S03 -) and sulfide (HS-) at pH 7.0 and pH 9.0.
- 281-
1:0
1:1
3.3. Isomerisation and degradation of MX and E-MX MX is stable in water at pH 2 but is gradually degraded at pH 4 and
above (20,21). MX undergoes ring opening around pH 6-8 (18,20). At pH 7 a minor part of the MX loss is due to isomerisation to E-MX (Fig. 4). Also E-MX undergoes degradation at neutral or alkaline pH. Under acid conditions E-MX will slowly be converted to MX (20). At pH 2 all E-MX will convert to MX.
It is known that the mutagenicity of chlorinated waters can be destroyed by treatment with sulfite (22-25). Therefore we studied the reaction of MX with sulfite. MX did react with sulfite (Fig. 4). The reaction rate was fairly independent of pH in the range 5-9. The strong nucleophile ion hydrosulfide reacted as expected even faster with MX.
Sulfite treatment may be a practical method for destruction of MX in drinking water and pulp chlorination liquors. The conditions needed for complete removal are still to be determined.
Combined mill effluents from kraft mills contain only a minor portion of the mutagenicity and the MX present in the chlorination liquors (14). The mutagenicity loss may be due to reaction with sulfide originating from the cooking liquor when the various mill process streams are combined.
3.4. Genotoxic and other properties of MX MX exhibits an extremely high mutagenic activity in the Ames assay
with the tester strains TA 100 and TA 98. With values of 3000-13000 revertants/ nmole on TA 100 and 180-800 revertants/nmole on TA 98 (8,18,20) MX is among the strongest bacterical mutagens tested so far.
It was recently reported that MX induces chromosomal aberrations in Chinese hamster ovary cells in culture, but did not induce micronuclei in mouse bone marrow cells when administered up to toxic doses (18).
MX is fairly hydrophilic (Kow = 11.9) and one would thus not expect it to have a high bioaccumulation propensity (14).
There is still not much information on the toxicological properties of MX. In view of the extreme mutagenicity of MX and the chronic exposure to man and the environment via drinking water and pulp mill effluents more studies on MX, and E-MX, are urgently needed.
4. CONCLUDING REMARKS The identification of a major Single mutagen in chlorinated waters
has opened new possibilities to achieve fundamental chemical understanding on the formation and removal of mutagenic activity in drinking water preparation and in wood pulp bleaching. This finding has also provided a new basis for the research to assess the risks to man and the environment of exposure to chlorinated waters.
REFERENCES
(1) ANDER, P., K.-E. ERIKSSON, M.-C. KOLAR, K. KRINGSTAD, U. RANNUG, and C. RAMEL, Sv. Papperstidn. 80: 454-459 (1977).
(2) GLATZ, B.A., C.D. CHRISWELL, M.D. ARGUELLO, H.J. SVEC, J.S. FRITZ, S.M. GRIMM, and M.A. THOMSON, J. Am. Water Works Assoc. 70:465-468 (1978).
(3) LOPER, J .C., D.R. LANG, R.S. SCHOENY, R.B. RICHMOND, P.M. GALLAGHER,
and C.C. SMITH, J. Toxicol. Environ. Health, 4:919-938 (1978). (4) MEIER, J.R., H.P. RINGHAND, W.E. COLEMAN, J.W. MUNCH, R.P. STREICHER,
W.H. KAYLOR, and K.M. SCHENK, Mutat. Res. 157:111-122 (1985). (5) KRINGSTAD, K.P. and K. LINDSTROM, Environ. Sci. Technol. 18:236A-248A
( 1984) •
- 282-
(6) KRINGSTAD, K.P., P.O. LJUNGQUIST, F. DE SOUSA, and L.M. STROMBERG, Environ. Sci. Technol. 15:562-567 (1981).
(7) HOLMBOM, B.R., R.H. VOSS, R.D. MORTIMER, and A. WONG, Tappi 64(3): 172-174 (1981).
(8) PADMAPRIYA, A.A., G. JUST, and N.G. LEWIS, Can. J. Chem. 63:828-832 (1985).
(9) KRONBERG, L., B. HOLMBOM, and L. TIKKANEN, vatten 41:106-109 (1985). (10) VARTIAINEN, T., and A. LllMATAINEN, Mutat. Res. 167:29-34 (1986). (11) KRONBERG, L., B. HOLMBOM, and L. TIKKANEN, Sci. Total Environ. 47:
343-347 (1985). (12) HEMMING, J., B. HOLMBOM, M. REUNANEN, and L. KRONBERG, Chemosphere
15:549-556 (1986). (13) KRONBERG, L., B. HOLMBOM, J. HEMMING, and L. TIKKANEN, in Proc. XVlth
Annual Meeting of the European Environmental Mutagen Society, pp. 104-109 (Brussels, Aug. 25-30, 1986).
(14) HOLMBOM, B., L. KRONBERG" P. BACKLUND, J. HEMMING, R. REUNANEN, A. SMEDS, and V. LANGVIK, Paper presented at the Sixth Conference on Water Chlorination, Oak Ridge, TN, May 3-8, 1987.
(15) BACKLUND, P., L. KRONBERG, G. PENSAR, and L. TIKKANEN, Sci. Total Environ. 47:257-264 (1985).
(16) HEMMING, J., B. HOLMBOM, S. JARNSTROM, and K. VUORINEN, Chemosphere 13:513-520 (1984).
(17) KRONBERG, L. and T. VARTIAINEN, manuscript in preparation. (18) MEIER, J.R., R.B. KNOHL, B.A. MERRICK, and C.L. SMALLWOOD, Paper
presented at the Sixth Conference on Water Chlorination, Oak Ridge, TN, May 3-8,1987.
(19) HORTH, H., M. FIELDING, H.A. JAMES, M.J. THOMAS, T. GIBSON, and P. WILCOX, Ibid.
(20) KRONBERG, L., B. HOLMBOM, and L. TIKKANEN, Ibid. (21) HOLMBOM, B.R., R.H. VOSS, R.D. MORTIMER, and A. WONG, Environ. Sci.
Technol. 18:333-337 (1984). (22) ERIKSSON, K.-E., K. KRINGSTAD, F. DE SOUSA, and L. STROMBERG, Sv.
Papperstidn. 85(9):R73-R76 (1982). (23) DONNINI, G.P., Pulp Paper Can. 84(4):T93-T98 (1983). (24) CHEH, A.M., J. SKOCHDOPOLE, P. KOSKI, and L. COLE, Science 207:90
(1980) • (25) WILCOX, P., and S. DENNY, in Water Chlorination - Chemistry, Environ
mental Impact and Health Effects, Vol. 5, Eds. R.L. Jolley, R.J. Bull, W.P. Davis, S. Katz, M.H. Roberts, Jr., and V.A. Jacobs, Lewis Publishers Inc., Chelsea, MI, 1985, pp. 1341-1353.
- 283-
THE FORMATION AND REMOVAL OF CHEMICAL MUTAGENS DURING DRINKING WATER TREATMENT
M. FIELDING and H. HORTH WRc Environment, Medmenham Laboratory, Henley Road,
Medmenham, POBox 16, Marlow, Buckinghamshire. SL7 2HD
Summary The application of bacterial mutagenicity tests to drinking water has shown the presence of mutagens. Invariably, mutagens generated during water treatment chlorination account for much of the mutagenicity detected in most surface-water-derived drinking waters. These mutagens are formed from widespread, naturally occurring precursors, such as humic substances, although other substances, especially amino acids may be involved. Most of the chlorination-derived mutagens identified account for little of the mutagenicity of drinking water. However, one mutagen, MX, is highly potent and may have a significant contribution. Its occurrence and toxicity need full evaluation. Other oxidants/disinfectants, such as ozone and chlorine dioxide, can also generate mutagens but these are likely to differ from those produced by chlorination. Granular activated carbon is effective in removing chlorination-derived mutagens but less effective for precursors. Dechlorinating agents, such as sulphur dioxide, can eliminate some of the mutagens produced by chlorination. The bacterial mutagenicity test cannot give information on the actual risks to health posed by the mutagens. Such information could arise from follow up work on identified major mutagens and from application of other bioassays more indicative of effects in man. Consequently, at the present time it is difficult to justify major changes to treatment practice aimed solely at controlling mutagenicity.
1. INTRODUCTION The presence of low concentrations of a wide range of organic
chemicals in drinking waters has aroused considerable concern about any significance to health. Some of these chemicals have been identified and, where sufficient toxicological data is available, their significance to health evaluated. However, many compounds are known to be present which cannot be identified. In order to avoid the arduous task of identifying all the chemicals present, tests were sought which would enable resources to be focused on substances of high potential concern.
One such test that has become widely applied is the bacterial mutagenicity test, often referred to as the Ames test. This test when applied to water can indicate the presence of chemical mutagens, substances which have a high probability of being carcinogenic. The test, however, only gives an indication of a qualitative risk. Thus, a positive mutagenicity result obtained with drinking water cannot be translated into a quantitative risk to health that would be associated with consumption of the water, unless either the active chemicals are identified and their toxicity assessed or the drinking water is tested in other bioassays more indicative of effects in man.
Application of bacterial mutagenicity tests to drinking water in recent years has indicated the presence of mutagens (potential carcinogens) in a variety of drinking waters. Most surprising is the finding that chlorination of natural waters generates such substances and these chlorination-derived mutagens'account for much of the mutagenicity in most surface water-derived drinking waters. The generation of mutagens by other disinfectants, such as ozone and chlorine dioxide, has received less study
-284-
and, consequently, a comparison of the merits of alternative disinfectants is not easy to make. The formation of mutagens during water treatment has led to studies on their elimination.
This paper focuses on the formation of chlorination-derived mutagens, their nature and their removal, or avoidance using alternative oxidants.
2. DETECTION OF MUTAGENICITY IN WATER AND THE FORMATION OF CHLORINATION-DERIVED MUTAGENS
2.1 Detection of mutagens in drinking water The Salmonella/microsome mutagenicity assay was developed by Ames and
coworkers (1). The assay uses strains of Salmonella typhimurium which contain mutations in genes that code for enzymes involved in the biosynthesis of the amino acid, histidine. Consequently, these stains cannot synthesise histidine and only grow if histidine is supplied in their growth medium. Spontaneous reversion of the mutant site in the gene occurs naturally, but infrequently, and this results in the regeneration of the normal gene and the cell regaining its ability to synthesise histidine. Such reversions can be detected by the ability of the cells to multiply and produce visible colonies on histidine-free medium. The action of mutagens on these strains causes an increase in the frequency of reverse mutations and the presence of a mutagen can be inferred from a significant increase in the number of colonies growing on histidine-free medium. Different strains can be ~sed for testing. Such strains are reverted by different types of mutagens, for example strain TA98 responds mainly to frameshift mutagens while strain TA100 is most sensitive to base-pair substitution mutagens. An additional feature of the mutagenicity test is the use of rat liver homogenate (termed S9) to simulate mammalian metabolism. Many chemicals are only mutagenic after conversion to mutagenic metabolites in the body while some mutagenic chemicals may be metabolically inactivated. The usual Salmonella/ microsome assay involves mixing several components, such as the bacterial strain, test chemical and S9 (if necessary) with molten salt agar and overlaying onto an agar plate.
Some research groups, including ourselves, use a modification of this method known as the fluctuation test. In this test exposure to the chemical takes place in a liquid medium in a large number of replicate cultures. Growth may occur in any of these replicates only if a reverse mutation is induced in one or more of the cells. If the added chemical is mutagenic then it is detected by an increase in the number of positive cultures (where growth has occurred) over the untreated controls. Growth is usually detected by a change in the colour of an indicator in the liquid medium due to the pH change. The fluctuation test is often more sensitive than methods involving agar plates.
It is difficult to detect, and consequently study, the mutagenicity of drinking water directly because of the low concentrations of organic chemicals involved (2). Consequently, mutagens have to be isolated and concentrated from water prior to mutagenicity testing. A variety of concentration techniques have been used for studying mutagenicity in drinking water (3). Each concentration technique is selective and the research carried out to date indicates that different techniques recover, in part, different mutagens. Studies that have included concentration of mutagens from water under different pH conditions have also demonstrated the existence of different mutagens. Consequently, comparison of data from different studies is difficult because one is dealing with mixtures of unknown mutagens that are detected by their overall mutagenicity. In the absence of the identities of the mutagenic components, techniques cannot be calibrated in the normal sense by the substances being measured. As stated already, there is no quantitative relationship between mutagenicity and risk to health and, consequently, caution is needed when comparing mutagenicity data.
-w-
2.2 Formation during treatment The application of bacterial mutagenicity assays to concentrated
extracts of water has demonstrated the presence of mutagens (4-8). Yhile raw waters occasionally show mutagenicity, which is usually weak, chlorinated, treated surface waters consistently exhibit mutagenicity. Many studies have demonstrated the generation of mutagenicity during chlorination of water supplies. The sporadic detection of mutagenicity in raw waters and the generation of mutagenicity during chlorination, found in our work, is shown in Tables I and II. Surface-derived waters showed little evidence of mutagenicity in TA100(-S9) but in the presence of S9 a few sites exhibited activity. Chlorination produced a positive response in TA100(-S9) in all surface water samples. Chlorinated river waters generally exhibited slightly higher levels of TA100(-S9) activity than upland waters. S9 reduced the level of TA100 activity in all cases. All chlorinated samples showed activity with TA98(-S9) but generally it was noticeably weaker than with strain TA100(-S9) (Table II). The effect of S9 on TA98 mutagenicity was variable. Several sites showed weak TA98(+S9) mutagenicity before chlorination while one sample gave a clear positive (significant at the 0.1% level) with TA98(-S9). No correlation was found between mutagenicity and the presence or absence of waste water in the raw water. Follow up studies (9) have confirmed that the production of mutagenicity during chlorination is not related to contamination of raw waters and that ubiquitous, naturally occurring precursors are involved.
The general conclusions on the effects of chlorination on mutagenicity, i.e. a marked increase in TA100(-S9) activity, a much smaller increase in TA98(-S9) mutagenicity and often a decrease in both brought about by the addition of S9 and the likelihood of natural precursors, have been found in several other studies.
3. THE NATURE OF CHLORINATION-DERIVED MUTAGENS Since mutagenicity results from reaction of chlorine with organic
compounds in drinking water, attempts to identify mutagens have focused on the byproducts of chlorination. Similarities were observed between chlorinated drinking water, chlorinated wood pulp and chlorinated aqueous solutions of humic, fulvic and amino acids, with respect to mutagenicity and chlorination products. These chlorination studies have contributed much to the identification of mutagens and other byproducts in drinking water.
3.1 Products of reaction between chlorine and or anic com ounds. Since t e identification of c loroform by Rook (10) in rln ing water,
several hundred other organic micropollutants have been identified (11). Many of the chlorinated compounds, e.g. chloroform, chloral, chlorophenols and dichloroacetonitrile were also identified in chlorinated aqueous solutions of isolated, extracted humic and fulvic acids (12,13,14) and amino acids (15). A variety of chlorinated, saturated and unsaturated alkanoic acids, summarised by Christman and others (12), appear to be major products of the aqueous chlorination of humic material, extracted from water. De Leer and others (14) identified over 100 products in chlorinated aqueous solutions of humic acids extracted from soils. Many of these were identical to those previously identified by Christman, but included additional products such as chlorinated carboxylic acids, w-cyanoalkanoic acids, and trichloromethylhydroxy acids and trichloromethylketo acids (chloroform precursors). De Leer and others (16) also identified chlorinated hydroxyamoyl compounds which appeared to be intermediate products in the formation of dichloroacetonitrile, dichloroacetic acid and dichlorosuccinic acid, and he concluded that nitrogenous compounds played an important role in the formation of chlorination byproducts.
-m-
Table I Mutagenic activity of freeze-dried extracts prepared from raw and laboratory chlorinated water samples in Salmonella typhimurium TA100
Slope values (level of significance) Sample Raw water Chlorinated water
TA100 - S9 TA100 + S9 TA100 - S9 TA100 + S9
Lowland rivers 120 0.04 (NS) 2.95 (+++) 11. 75 (+++ ) 3.49 (+++) 121 0.96 (NS) 1.09 (NS) 17.30 (+++ ) 1.60 (+++) 122 0.38 (NS) Toxic 16.17 (+++) 1.42 (+++) 123 0.66 (NS) 6.73 (+++) 7.21 (+++) 1.16 (NS) 124 1.40 (NS) 5.23 (+++) 19.87 (+++) 0.04 (NS)
Upland reservoirs 125 0.07 (NS) Toxic 12.25 (+++) 0.28 (NS) 126 < 0 (NS) Toxic 9.77 (+++) Toxic 127 < 0 (NS) Toxic 4.41 (+++ ) Toxic
Groundwater 128 < 0 (NS) 3.19 (+++) < 0 (NS) 2.17 (+++)
Slope values were derived from dose response plots with each sample in the fluctuation assay. NS Not significant
+ Significant at 5% level ++ Significant at 1% level
+++ Significant at 0.1% level
Table II Mutagenic activity of freeze-dried extracts prepared from raw and laboratory chlorinated water samples in Salmonella typhimurium TA98
Slope values (level of significance) Sample Raw water Chlorinated water
TA98 - S9 TA98 + S9 TA98 - S9 TA98 + S9
Lowland rivers 120 1. 31 (+) 2.16 (+++) 7.54 (+++) 7.92 (+++) 121 2.03 (NS) 2.38 (++) 10.40 (+++) 7.20 (+++) 122 0.86 (NS) 2.72 (NS) 5.62 (+++) 3.08 (NS) 123 5.50 (+++) 1.45 (NS) 5.57 (+++) < 0 (NS) 124 0.85 (NS) 6.10 (+++) 5.34 (+++) 5.57 (+++ )
Upland reservoirs 125 0.23 (NS) 3.06 (++) 3.76 (+++) 3.28 (+++) 126 < 0 (NS) 2.63 ( ++) 0.80 (NS) 4.31 ( +++) 127 < 0 (NS) 4.04 (+++) 1. 88 (+) 5.45 (+++)
Groundwater 128 0.38 (NS) 1.38 (NS) 2.09 (+) 2.88 ( ++)
Key: See Table I
3.2 of muta ens in drinkin water. ucts are known not to be mutagenic in
bacterial assays, while many others have yet to be tested for mutagenicity. Clearly, the task of identifying all the chlorination products and obtaining toxicological information is impractical. For this reason, work was restricted to toxicologically relevant compounds by combining fractionation techniques with mutagenicity testing and analysis of mutagenic fractions. The aim was to isolate mutagenic compounds for subsequent identification. The technique has been applied to chlorinated wood pulp (17), chlorinated aqueous solutions of humic acids (18) and
- 287-
drinking water (19,20). The use of model compounds that are precursors of mutagenicity (e.g. amino acids), has also been applied to simplify the task of identifying mutagens (15,20,21).
3.3 Precursors of mutagenicity Mutagenicity of similar characteristics (i.e. predominantly TA100
activity, inactivated by S9 metabolic activation, partly extractable by XAD adsorption at neutral pH and partly at low pH) is produced when chlorinating treated water, wood pulp, aqueous solutions of isolated humic and fulvic acids and certain amino acids. This may indicate that the same mutagens are formed because the above materials have common structural components. There are structural similarities between wood pulp, which is mainly composed of lignins, and structural components of aquatic humic material, which is partly derived from lignin (22). Amino acids could represent structural components of humic substances or produce the same intermediates on chlorination as humic material, and this could lead to the formation of the same mutagens.
Table III lists organic materials and individual chemicals which produce mutagenic activity on aqueous chlorination. Although a large variety of compounds can be precursors of mutagenicity, it is difficult to assess precisely their relevance to drinking water treatment. Undoubtedly, humic and fulvic acids in water react with chlorine to produce mutagens. Free amino acids occur in most waters, but usually at low levels, and their contribution to the formation of mutagens is likely to be small. However, amino acids and many of the other acids and phenolic compounds listed in Table III represent structural components of humic material, and could therefore, be involved in the formation of mutagens.
In our own experiments (15,20,21) we used substrate concentrations (in terms of total organic carbon) and pH conditions (buffered to pH 6.2) similar to a drinking water treatment plant. In addition we used XAD adsorption followed by diethyl ether elution to concentrate the chlorination products and compared the mutagenic activity with an extract of drinking water. ~e found similar levels of mutagenic activity in chlorinated drinking water, chlorinated aqueous solutions of humic acids from different sources (extracted from peat bogs, lake water, and commercial humic acid probably extracted from soil) and from a chlorinated aqueous solution of a mixture of 21 amino acids. Not all amino acids were precursors of equally potent mutagenic activity, the major precursors among those tested being methionine, tyrosine, phenylalanine, tryptophan, histidine and proline (15). The mutagenic activity produced from bound amino acids (albumin and short-chain peptides, see Table III) was insignificant compared with that produced from amino acids (21).
The precursors of mutagenicity reported by Rapson and others (25) (see Table III) were, wood pulp, lignin, and many lignin model compounds and intermediate products of wood pulp chlorination. Their experiments were carried out under conditions appropriate for wood pulp bleaching, i.e. relatively high concentration of substrate (4-8 mMol.l-l) and probably low pH. However, Holmbom and others (26) chlorinated some of the same phenolic compounds at lower concentration (about 0.3 mMol.l-l) at pH 7 (acidified subsequently for extraction) and detected mutagenic activity (Table III). The experiments with phenolic compounds reported by Onodera and others (27) were carried out at pH 7 and at 0.5 mMol.l-l substrate concentration.
3.4 Mutagenic products of chlorination Table
identified and amino were not necessarily differences techniques,
IV lists mutagenic products of chlorination which have been in chlorinated drinking water, chlorinated aqueous humic, fulvic acids and in chlorinated wood pUlp. The fact that some mutagens
identified in particular chlorinated substrates does not mean that they were not present. It may be due to quantitative or to the application of different extraction and analytical
which failed to detect them.
-288-
Table III. Precursors of mutagenicity in aqueous chlorination
Precursor Reference Precursor Reference
humic acid (28) phenol (25,26) fulvic acid (23) catechol (25,26) lignin (25) resorcinol (25) amino acids (15) 3-methylcatechol (25)
methionine (15,29) 4-methylphenol (27) phenylalanine (15,29) hydroquinone (25) tyrosine (15,.29) phloroglucinol (25) tryptophan (15) pyrogallol (25) histidine (15) guaiacol (25,26) proline (15) veratrole (25) hydroxyproline (15) p-hydroxybenzoic acid (25,26) cysteine (15,29) 3,4-dihydroxybenzoic acid (25) cystine (15) ferulic acid (26) glycine (29) o-vanillic acid (25)
protein/peptides (21) p-vanillic acid (25) albumin (bovine serum) (21) eugenol (25) glycyl-phenylalanine (21) fumaric acid (25) phenylalanyl-glycine (21) muconic acid (25)
acetovanillone (25) 4-phenylphenol (27) acetoveratrone (25) 2,4,6-trichlorophenol (25) m-hydroxyacetophenone (25) 4,5-dichlorophenol (25) p-hydroxyacetophenone (25) 2,5-dichloro-p-benzoquinone (25) m-methoxyacetophenone (25) 2,6-dichloro-p-benzoquinone (25) p-methoxyacetophenone (25)
Most of the mutagens in Table IV are direct-acting TA100 mutagens. Haloforms, halo-alkanes and halo-alkenes, chlorinated ketones, aldehydes, nitriles and furanones are the predominant groups of mutagens identified. Probably the presence of brominated products of chlorination is due to the presence of bromide ions in the raw water; the bromide is converted to hypobromous acid, which reacts faster than hypochlorous acid.
Although the number of identified mutagens is considerable, most do not account for much of the detected mutagenic activity. For example, we tested a mixture of 14 identified mutagens, mainly haloforms, halo-alkanes and halo-alkenes, and halo-nitriles. Ve also tested the mixture spiked into a pH 6.2/XAD extract of drinking water (15). Ve estimated that the contribution of these mutagens together, at the low ~g concentrations normally found in drinking water, accounted for less than 10% of the activity in the extract. Similar results were reported by Meier and others (31) who estimated that several chlorinated ketones and aldehydes, identified in low pH extracts of chlorinated, aqueous solution of humic acid, accounted in total for less than 10% of the activity of the low pH extract.
MX is the most interesting mutagen identified to date (see Table IV). It is a highly potent, direct-acting TA100 mutagen and is sufficiently potent to account for a significant proportion of the mutagenicity of drinking water, even if present at low ng.l- 1 concentration. It was first identified in chlorinated wood pulp (17), and later in humic-rich water (24), drinking water (24) and chlorinated aqueous solutions of isolated humic acid (34) and tyrosine (21). MX was detected at concentrations ranging from 5 to 46 ng.l- 1 in pH 2/XAD extracts of drinking waters, and was estimated to contribute to between 22% and 46% of the mutagenic activity of the same extracts (26). Ve obtained similar results. Figure 1 shows the mutagenicity of MX and an extract (pH 2/XAD-2) of treated chlorinated upland
- 289-
Table IV. Bacterial mutagens produced by aqueous chlorination
bromoform bromochloromethane bromodichloromethane dibromomethane dibromochloromethane bromoethane 1,2-dichloroethane iodoethane I-bromo propane I-bromobutane trichloroethylene tetrachloroethylene tetrachloropropene pentachloropropene 1,3-dichloropropanone 1,I-dichloropropanone 1,1,I-trichloropropanone 1,1,3-trichloropropanone 1,1,3,3-tetrachloropropanone hexachloropropanone 3,S,S-trichloropent-4-ene-2-one chloral chloroacetaldehyde 2-chloropropenal 3,3-dichloropropenal 2,3,3-trichloropropenal 2-Phenyl-2,2-dichloroethanal chloropicrin(a) bromochloroacetonitrile dichloroacetonitrile trichloro-l,2,3-trthydroxybenzene bromo-p-cymene dichloro-p-cymene E2-chloro-3-(dichloromethyl)-4-oxo-butenoic
A(IS) A(1S) A(1S) A(lS) A(1S) A( IS) A(1S) A(1S) A(1S) A(1S)
B(31) A(31) B(31) A(31) B(31)
B(31) B(31)
A(11) B(14) C(21)
B(31) B(31) B(31)
C(21) A(1S) A(1S) A(IS) B(31) C(21)
0(30) 0(30) 0(30)
0(30) 0(30) 0(30) 0(30) 0(30)
0(30) 0(30) 0(30)
0(30) 0(30)
0(30) 0(30) 0(30)
acid (E-MX) 3-chloro-4-(dichloromethyl)-S-hydroxy-2(SH)
furanone (MX) 3,4-dichloro-S-(dichloromethyl)-S-hydroxy
furanone
A(32) B(32) C(21) 0(26)
A in chlorinated drinking water B in chlorinated humic substances C in chlorinated amino acids
A(24) B(24) C(21) D(17)
D in chlorinated wood pulp o Reference
0(33)
(a) TAI00 with S9 metabolic activation
water. The level of MX in the extract (about 23 ng 1-1 accounted for about 60% of the mutagenic activity. E-MX, first identified by Kronberg and others (3S), is an isomer of the open form of MX and may be converted to MX under certain conditions. Its mutagenic activity has not been fully evaluated. It is less potent than MX, but may, nevertheless, be the second most important mutagen so far identified in drinking water.
The other furanone in Table IV is much less potent than MX and is likely to be less significant than MX, if it is present in drinking water.
It is interesting to nOte that tyrosine (a phenolic amino acid) has been identified (21) as a precursor of MX (Table V). In addition, Holmbom and others (26) detected MX in chlorinated aqueous solutions of phenolic compounds. Norwood and others (36) have shown that phenolic components of humic material react with aqueous chlorine. Although these components comprise a small proportion of the humic material, they may be important in
- 290-
the formation of MX, since MX is an extremely minor chlorination product, but highly significant in terms of its contribution to the mutagenicity of such mixtures.
5 TA100 (No 88)
0 e • • 4 CI.
: c • -.. 3 e > e .. '0
• 2 ~
E ,. C
'V
! 1 • E
• w
0.01 0.02 0.04
Watar equl"alent. (lltre./ml)
i
0 0.75 1.5
Chlorofuranone (UX) ngJlII1
Figure 1. Mutagenicity of Treated Upland Water (pH2/XAD-2 extract) alMl of MX.
Table V. Precursors of MX
wood pulp humic acid tyrosine guaiacol phenol p-hydroxybenzoic acid catechol ferulic acid
4. REMOVAL OF CHLORINATION-DERIVED MUTAGENS
Reference (17) (34) (21) (26) (26) (26) (26) (26)
For reasons already stated, the significance to health of the mutagenicity generated by chlorination is unclear. Consequently, it is difficult to justify major changes to treatment practice intended solely to remove such mutagens, and as far as can be ascertained, no treatment plants have been modified and operated specifically to eliminate chlorinationderived (or any other) mutagens. However, several groups have investigated
- 291-
removal or avoidance of chlorination-derived mutagens but, usually, as just one factor in addition to others, such as organoleptic characteristics, adsorbable organohalogens and trihalomethanes (THMs). Since conventional filtration methods are not usually effective, this paper reviews the possibility of avoiding or removing chlorination-derived mutagens by using alternative oxidants/disinfectants, granular activated carbon (GAC) and dechlorination.
4.1. Do other oxidants produce mutagens? ~hen it was realised that water treatment chlorination generated
mutagenicity by reaction of chlorine with organic matter in raw waters, attention focused on other oxidants/disinfectants, such as ozone and chlorine dioxide, and whether they also produced mutagenicity when added to raw waters. Consequently, several studies have investigated this question, particularly in the case of ozone.
4.1.1. Ozone Koor--and Van Kreijl (6) demonstrated that river water, which produced
marked increases in mutagenicity during breakpoint chlorination, generally produced no significant mutagenicity upon ozonation (2 mg ozone 1-1 at 10 min contact time). In fact, the mutagenicity in raw water was reduced slightly on ozonation. However, at one plant a small increase in TA98(-S9) mutagenicity was encountered.
Overall, mutagenicity in the river Rhine water (settled, coagulated dual-media-filtered) was reduced slightly by ozonation (2.7 mg/l ozone, 7 min contact time) whereas chlorination produced a dramatic increase in TA100(-59) and TA98(-59) mutagenicity (5).
Backlund and others (37) found that humic water (water rich in humic matter) and flocculated humic water generated no mutagenicity on ozonation (10 and 2.9 mg 1-1 respectively). There are indications that generation of mutagenicity during application of ozone may depend more (compared to chlorination-derived mutagenicity) on treatment conditions, especially ozone dose, and on the quality of water being treated. Cognet and others (38) found that the relationship between ozone dose and mutagenicity was complex. At one site (groundwater - possibly recharged with treated river water, biological nitrification) ozone reduced overall mutagenicity when applied for both short and long periods but increased it at intermediate contact times. At another plant treating sand-filtered river water, an increase in the mutagenicity of the water was encountered occasionally. Usually any mutagenicity generated by ozone was detected by TA98 (-59). A similar finding was encountered by Lykens and others (39) who investigated Mississippi river water (clarified, pressure sand-filtered). Ozone treatment (about 10 mg 1-1 , 30 min contact time) produced no detectable mutagenicity whereas chlorination produced mutagenicity detectable with both TA98(-59) and TA100(-59).
Van Hoof and others (8) also found that ozonation can produce mutagens under certain conditions. River Meuse water, after storage, was ozonated with different doses of ozone prior to coagulation and dual filtration and showed generation of mutagenicity, mainly TA98(-59). However, the mutagenicity produced was variable, strong in some samples but not others. Little activity due to TA100 mutagens was encountered, which contrasted with earlier studies. The authors suggested that the absence of TA100 mutagens, was due to a different point of application, i.e. preozonation -before coagulation and filtration. They concluded that, whereas chlorination produces basically the same type of mutagenicity, regardless of the point of chlorine application in the treatment process, the mutagenicity produced by ozone was to some extent dependent on its point of application. An ozone dose of 1 mg 1-1 appeared to produced more mutagenicity than higher doses although at each dose level results were variable.
-m-
4.1.2. Chlorine dioxide Chlorine dioxide regularly produced mutagenicity, TA98(-S9), when
applied to clarified, pressure-sand-filtered Mississippi river water (39). Unlike chlorine, no TA100 mutagenicity was encountered.
High doses of chlorine dioxide (up to 15 mg 1-1) applied to stored river water (6) also increased TA98(-S9) mutagenicity. Application of less than 1 mg 1-1 chlorine dioxide produced no increase in mutagenicity (6). However, Backlund and others (37) found no production of mutagenicity, TA100(-S9) and TA98(-S9), when humic'water and flocculated humic water were treated with chlorine dioxide (21 mg 1-1 and 6.5 mg 1-1 respectively). In their more recent work chlorine dioxide was found to generate mutagenicity under, apparently, very similar conditions (26).
4.1.3. Monochloramine Little work on the ability of monochloramine to produce mutagenicity
has been reported. Lykens and others (39) applied monochloramine to clarified, pressure-sand-filtered Mississippi river water and this consistently produced marked increases in TA100(-S9) and TA98(-S9) mutagenicity. The action of monochloramine was similar to that of chlorine but the level of mutagenicity was less.
4.2. Removal of chlorination-derived mutagenicity by granular activated carbon (GAC) GAC is a means of reducing the level of organic micropollutants in
drinking water. Consequently, it is not surprising that several groups have examined the ability of GAC to reduce the levels of mutagenicity produced by chlorination.
Monarco and others (40) studied the effectiveness of GAC for removing mutagenicity from Ohio river water (after 3 days storage, chlorination, coagulation/flocculation, sedimentation and rapid-sand filcration). With fresh GAC removal of mutagenicity was high, especially for TA100(-S9) mutagens. When the GAC was 3 months old the overall removal had dropped to about 87% (based on numbers of revertants in extracts) although removal of total organic carbon (TOC) had fallen to about 34%, which indicates selective removal of mutagens. However, the mutagens which were emerging from the GAC after this time were found in acid fractions, in the case of TA100(-S9) and in both neutral and acid fractions in the case of TA98(-S9). This suggests that in addition to selective removal of mutagens (compared to organic matter in general) there may be selective removal of different types of mutagens.
The above study was extended by Loper and others (7) who looked at the same treatment plant over a longer time (up to 35 weeks) but with different sample extraction techniques applied at neutral pH. The efficiency of GAC for removal of mutagens produced by chlorination was confirmed. Removal of TOC was only around 35%. In general, no mutagenicity was detected after GAC which differed from the earlier study, where a small breakthrough of mutagenicity was encountered at 3 months. Possibly this was due to the application of different extraction techniques.
Kool and Van Kreijl (6) also showed that mutagenicity (most of which produced by chlorination) can be reduced substantially by filtration through GAC. At a plant treating river Meuse water, mutagenicity, TA98(-S9) and TA100(-S9), was not detectable after filtration over GAC which had been in service for less than 1 year. Removal of TOC was about 32%, which again shows selective removal of mutagens.
Van der Gaag and others (5) looked specifically at the effectiveness of GAC filtration in relation to mutagens generated by chlorination, over long periods of time (up to 83 weeks). No mutagenicity broke through the GAC in this time.
Lykens and others (39) demonstrated the efficient removal of mutagens produced by chlorination of clarified, pressure sand-filtered Mississippi
river water. TA100(-S9) and TA98(-S9) mutagenicity was effectively removed for up to 14 months, at which time breakthrough of TA98(-S9) mutagenicity occurred.
4.2.1. Final chlorination of GAC treated water Although GAC filtration can produce water in which mutagenicity cannot
be detected, breakthrough of the precursors of mutagenicity (i.e. residual organic substances from the raw water, or produced during treatment, that react with chlorine to produce mutagens) is a possibility. These precursors could then react during final disinfection to produce mutagenicity. Several groups have reported on this aspect, Van der Gaag and others (5) demonstrated that breakthrough of such precursors after GAC treatment of chlorinated water occurred relatively early in the life of a GAC bed. Increases in TA100(-S9) activity occurred after about 6 weeks in the case of acidic fractions and after about 8 weeks in the case of neutral fractions. Yith GAC treatment of ozone-treated water, a similar breakthrough, but delayed and of lower intensity, was encountered.
Lykens and others (39) found that, by 6 months, GAC was only partly removing precursors of mutagenicity, TA100(-S9) and TA98(-S9), from chlorinated, treated river water.
Figures 2 and 3 show data from work at VRc which demonstrates several of the main features of removal of chlorination-derived mutagens by GAC, i.e. significant reduction of mutagenicity, early breakthrough of precursors and later breakthrough of mutagens.
4.3. Effect of dechlorinating agents on chlorination-derived mutagens Partial dechlorination is frequently used as a final treatment stage
to reduce the level of chlorine to more acceptable levels prior to distribution. Although in some situations sulphite and metabisulphite are used, sulphur dioxide is the most common dechlorinating agent employed. The ability of such agents to destroy some of the mutagenicity in water samples was indicated by Cheh and others (41) but a more detailed investigation was carried out by Wilcox and Denny (42). Using chlorination and dechlorination conditions as close as possible to those used in water treatment plants, the following conclusions were reached;
total dechlorination of chlorinated drinking waters with various dechlorinating agents substantially reduced chlorination-derived mutagenicity, TA100(-S9) and TA98(-S9), partial dechlorination had little effect unless the free residual was less than 0.5 mg 1-1 ,
low level rechlorination (up to 0.5 mg 1-1 ) of totally dechlorinated waters restored some TA100(-S9) mutagenicity, chloramination resulted in less 'restored activity'.
Dechlorination did not eliminate all activity which suggests the presence of less labile chlorinated mutagens.
5. CONCLUSION Chlorination of raw waters consistently produces mutagenic substances,
mainly from reaction with widespread, naturally occurring organic constituents of water. Yhat effect, if any, such mutagenic substances have on human health is unknown but any effects are probably small. To assess such potential effects, either the nature of the mutagens needs to be established, which will enable a variety of follow up evaluations to be undertaken, or other tests nned to be applied that better indicate effects in man. As others have already stated (6), while mutagenicity tests give an indication of a qualitative risk, they cannot be used to estimate risks in a quantitative sense and, consequently, should not be used as a guide to water quality or as water quality standards at the present time.
Many mutagens produced by chlorination have been identified but to date only the compound MX, and to a lesser extent E-MX have sufficient
-m-
• :0 .. > • Q. 0 iii
• :0
• >
• ~ .2 .,
ao.o Ra •• at.r v •• for. GAC (c"lorlnat.~) 1 Aft.r GAC
25.0 0 Contact tank
20.0
15.0
10.0
5.0
0 0 1.0 2.0 3.0 4.0 5.0 8.0 7.0 8.0 1.0
Ag. of b.d (lIIonlh.)
Figura 2. Effact of GAC on Mutagenicity (TAI8-SI).
180.0 x Raw.ater v Befor. GAC (chlorlnat.d)
180.0 1 After GAC 0 Contact tank
140.0
120.0
100.0
ao.O
80.0
40.0
20.0
0±0~S;~'~.0~::~2~.0;:~~~====~:=::~~:::::;;=====:======:====~ 3.0 4.0 5.0 8.0 7.0 8.0 1.0
Age of bed (montha)
Figure 3. Eff.ct of GAC on Mutagenicity (T A 1 OO-SI).
- 295-
potency to account for much of the mutagenicity detected in drinking water. The occurrence, fate and toxicity of these substances need to be established. It would seem wise to determine the effects of treatment processes 'on MX and rela ted subs tances.
Other oxidants/disinfectants can form mutagens and, in the case of ozone and chlorine dioxide, the mutagens are likely to differ from those produced by chlorination. Although the overall mutagenicity detected may be less or barely detectable in the case of ozone and chlorine dioxide, this does not necessarily mean less risk to health if different mutagens are involved. While more still needs to be known about chlorination-derived mutagens, much more information is needed on mutagens (and byproducts in general) generated by other oxidants/disinfectants -particularly ozone.
GAC is effective in removing mutagens produced by chlorination and, to a lesser extent, the precursors of chlorination-derived mutagens. However, at the present time it is difficult to justify major changes to treatment practice solely to control mutagenicity. Clearly it is prudent to know the behaviour of mutagenicity when evaluating the effects of treatment modifications on other parameters (such as THM levels, organoleptic properties, colour etc.) If mutagenicity is reduced by treatment, without substituting different mutagens, then it is conceivable that any risks to health have also been reduced, although the value of such a reduction cannot be determined at present. Thus, it may be possible in some situations to reduce mutagenicity simply by reducing the level of chlorination used. However, if such changes are undertaken, then disinfection must not be compromised.
ACKNOWLEDGEMENT The assistance of staff in the Mutagenicity, Toxicology and Mass Spectrometry sections of WRc is gratefully acknowledged.
REFERENCES (1) AMES, ~.N. (1971). The detection of mutagens with enteric bacteria.
A. Hollander (ed), Chemical Mutagens, Principles and Methods of Detection, Vol.1, Plenum Press, 267-282.
(2) FORSTER R. et al. (1983). Use of the fluctuation test to detect mutagenic activity in unconcentrated samples of drinking water in the U.K. R. L. Jolley et al. (eds.), Water Chlorination, Vol. 4, Ann Arbor Scientific Publ. 1189-1197.
(3) WILCOX P. et al. (1986). Isolation and chacterisation of mutagens from drinking water. A. Leonard and M. Kirsch-VoIders (eds.), Proceedings of the XVIth Annual Meeting, European Environmental Mutagen Society, 92-103.
(4) FAWELL, J.K. et al. (1986). Health aspects of organics in drinking water. Technical Report TR 231, Water Research Centre, Marlow, Bucks., U.K.
(5) Van der GAAG, M.A. et al. (1985). The influence of water treatment processes on the presence of organic surrogates and mutagenic compounds in water. Sci. Total Environ. 47, 137-153.
(6) KOOL, H.J. and Van KREIJL, C.F. (1984). Formation and removal of mutagenic activity during drinking water preparation. Water Res. 18, 1001-1016.
(7) LOPER, J.C. et al. (1985). Continuous removal of both mutagens and mutagen-forming potential by an experimental full-scale granular activated carbon-treatment system. Environ. Sci. Technol. 19, 333-339.
(8) Van HOOF, F. et al. (1985). Formation of mutagenic activity during surface water pre-ozonation and its removal in drinking water treatment. Chemosphere 14, 501-509.
- 296-
(9) BAKER, A. and ~ILCOX, P. (1987). An investigation on the orlgln of mutagens in water samples collected from three different sites along a lowland river. Report PRD 1468-M, ~ater Research Centre, Marlow, Bucks., U.K.
(10) ROOK, J.J. (1974). Formation of haloforms during chlorination of natural waters. ~ater Treat. Exam. 23, 234-243.
(11) FIELDING, M. et al. (1981). Organic micro-pollutants in drinking water. Technical Report TR 159, ~ater Research Centre, Marlow, Bucks., U.K.
(12) CHRISTMAN, R.F. et al. (1984). Oxidative degradation of aquatic humic material. Second International Humic Substances Conference, Birmingham.
(13) RECKHO~, D.A. and SINGER, P.C. (1985). Mechanisms of organic halide formation during fulvic acid chlorination and implications with respect to pre-ozonation. R.L. Jolley et al. (eds.), ~ater Chlorination, Vol.5, Lewis Publ. 1229-1257. ------
(14) De LEER, E.~.B. et al. (1985). Identification of intermediates leading to chloroform and C-4 diacids in the chlorination of humic acid. Environ. Sci. Technol. 19, 512-522.
(15) FIELDING, M. and HORTH, H. (1986). Formation of mutagens and chemicals during water treatment chlorination. ~at. Supply, 4, 103-126.
(16) De LEER, E.~.B. et al. (1986). Chlorination of w-cyanoalkanoic acids in aqueous medium. Environ. Sci. Technol. 20, 1218-1223.
(17) HOLMBOM, B. et al. (1981). Fractionation, isolation and identification o~ Ames-mutagenic compound in Kraft chlorination effluents. Tappi 64 (3), 172-174.
(18) MEIER, J.R. et al. (1986). Mutagenic by-products from chlorination of humic acid. Environ. Health Perspect. 69, 101-107.
(19) TABOR, M.~. and LOPER, J.C. (1980). Separation of mutagens from drinking water using coupled bioassay/analytical fractionation. Intern. J. Environ. Anal. Chem. 8, 197-215.
(20) HORTH, H. et al (1987). Techniques for the fractionation and identification of mutagens produced by water treatment chlorination. I.H. Suffet and M. Malaiyandi (eds.). Organic Polutants in ~ater, Advances in Chemistry Series 214. American Chemical Society.
(21) HORTH, H. et al. (1987). The production of organic chemicals and mutagens during chlorination of amino acids in water. 6th Conference on ~ater Chlorination, Oak Ridge, Tennessee.
(22) ERTEL, J.R. et al. (1986). Dissolved humic substances of the Amazon river system. Limnol. Oceanogr. 31, 739-754.
(23) MARUOKA, S. (1986). Analysis of mutagenic by-products produced by chlorination of humic substances by thin layer chromatography and high-performance liquid chromatography. Sci. Total Environ. 54, 195-205.
(24) HEMMING, J. et al. (1986). Determination of the strong mutagen 3-chloro-4(dichloro-methyl)-5-hydroxy-2(5H)-furanone in chlorinated drinking and humic waters. Chemosphere 15, 549-556.
(25) RAPSON, ~.H. et al. (19aO). Mutagenicity produced by aqueous chlorination of organic compounds. Bull. Environm. Contam. Toxicol. 24, 590-596.
(26) HOLMBOM, B. et al. (1987). Formation and properties of 3-chloro-4-(dichloro-methyl)-5-hydroxy-2(5H)-furanone, a potent mutagen in chlorinated waters. 6th Conference on ~ater Chlorination, Oak Ridge, Tennessee.
(27) ONODERA, S. et al. (1986) Chemical changes of organic compounds in chlorinated water. J. Chromatography 360, 137-150.
(28) MEIER, J.R. et al. (1983). Formation of mutagens following chlorination of humic acid; a model for mutagen formation during drinking water treatment. Mutation Res. 118, 25-41.
-m-
(29)
(30)
(31)
(32)
(33)
(34)
(35)
(36)
(37)
(38)
(39)
(40)
(41)
(42)
SUESSMUTH, R. (1982). Genetic effects of amino acids after chlorination. Mutation Res. 105, 23-28. RAPSON, Y.H. et al. (1985), Mutagenicity produced by aqueous chlorination or---tyrosine. R.L.Jolley et al (eds.), Yater Chlorination, Vol.5, Lewis Publ., 237-249. Meier, J.R. et al (1985). Identification of mutagenic compounds formed during--cliIorination of humic acid. Mutation Res. 157, 111-122. KRONBERG, L. et al (1987). Identification of the strong mutagen 3-chloro-4-(dichloromethyl)-5-hydroxy-2(5H)-furanone and of its geometric isomer E-2-chloro-3-(dichloromethyl)-4-oxo-butenoic acid in mutagenic fractions of chlorine treated humic water and in drinking water. 6th Conference on Yater Chlorination, Oak Ridge, Tennessee. STROEMBERG, L.M. et al (1987). An abundant chlorinated furanone in the spent chlorination liquor from pulp bleaching. Envi~on. Sci. Technol. 21, 754-756. MEIER, J.R. et al. Studies on the potent bacterial mutagen, 3-chloro-4-(dichloromethyl)-5-hydroxY-2(5H)-furanone: aqueous stability, XAD recovery and analytical determination in drinking water .and in chlorinated humic acid solutions. Mutation Res., in press. KRONBERG, L. et al (1986). Determination of the strong mutagen, 3-chloro-4-(dichloromethyl)-5-hydroxy-2(5H)-furanone (MX) and identification of other chlorinated compounds present in mutagenic fractions of chlorinated humic water. A. Leonard and M. Kirsch-VoIders (eds.), Proceedings of the XVlth Annual Meeting, European Environmental Mutagen Society, 104-107. NORWOOD, D.L. et al (1987). Structural characterisation of aquatic humic material. Environ. Sci. Technol., 21, 791-798. BACKLUND P et al (1985). Mutagenic activity in humic water treated with alternative disinfectants. H.A.M. de Kruijf and H.J. Kool (eds.), Organic Micropollutants in Drinking Water and Health. Elsevier, 257-264. COGNET, L. et al (1986). Mutagenic activity of disinfection by-products. Environ. Health Persp. 69, 165-175. LYKENS, B.W. et al (1986). Chemical products and toxicologic effects of disinfection. Jour. Amer. Wat. Yks. 78, 66-75. MONARCO, S. et al (1983). Removal of mutagens from drinking water by granular activated carbon. Water Res. 9, 1015-1026. CHEH, A.M. et al (1980). Nonvolatile mutagens in drinking water. Science 20, 9~ WILCOX, P. and DENNY, S. (1985). Effect of dechlorinating agents on the mutagenic activity of chlorinated water samples. R.L.Jolley et al (eds.), Yater Chlorination, Vol.5, Lewis Publ. 1341-1353.
-m-
APPLlCATI~ OF 'lHE OZCJ.lE-BYDROGEN PEROXIDE camINATI~ ¥ 'lHE REXNAL OF TOXIC CCJoIPOONDS FRCft A GlUHW.'1'ER
J.P. Duguet, C. Anselme, P. Mazounie and J. Mallevialle
Centre de Recherche Lyonnaise des Eaux-Degremont 38 rue du President Wilson, 78230 Le Pecq, France
SUMMARY
Nitro and chloro-benzenic compounds, which are widely used in dye industries, have been associated recently with groundwater contamination. Because of their potential toxicity and for taste and odor considerations, three main actions were decided to solve the problem. First, to follow the advance of pollution toward the wells, samples were collected automatically and analyzed using GC-MS. Results indicate that o-chloronitrobenzene was the main pollutant in concentrations ranging from 10 to 20,000 pg/l. Second, to monitor the drinking water quality, an on-line spectrophotometer was used to measure the optical density at 254 om at the inlet and outlet of the plant. Third, the feasibility of using the 0 /H 0 combination was determined at a 450 l/h pilot plant. Reduction of chfor6nttrobenzenes from 1.9 mgll to less than 20 pg/l could be reached by the application of 8 g o 1m3 and 3 g H 0 1m3 with a 20-minute contact time. To avoid an eventual ~cterial regr~ in the network due to biodegradability of the oxidation by-products, sand and GAC filtration were tested after oxidation. An evaluation of the costs of these different treatments is also presented.
1. INTlUXJCTI~ Organoleptic problems have always concerned the water supply industry.
New techniques, such as flavor profile analysis and closed loop stripping analysis (1), provide a better understanding of taste and odor causing compounds as well as mechanisms and origins. The use of such techniques for water quality control has allowed detection of pollution caused by stored wastes. These wastes have infiltrated the soil and polluted groundwater near wells used for drinking water production.
Chloro and nitrobenzenic compounds which are the main pollutants are widely used in chemical industries for dye synthesis. These types of compounds have been associated recently with water contamination (2,3). Because of their possible toxicity (4) and their effect on taste and odor (5), the use of such water for the production of drinking water requires a sophisticated treatment in order to remove these compounds.
To solve this pollution problem, three main courses of action were undertaken. First, to follow the path of the advance of pollutants toward the wells, samples were collected from different sites by use of a continuous liquid-liquid extraction device, after which GC/MS analysis on the extract was performed in the laboratory. Second, to monitor the quality of the drinking water, an on-line spectrophotometer was used. The optical density at 254 om (which is representative of the aromatic compounds) was measured continuously on water at the inflow and outflow of the drinking water treatment plant. Third, the feasibility of a complete treatment line comprising an 0 /H 0 combination for the detoxification of the groundwater was determined. 3Th@ Sim of this article is to present results concerning
- 299-
the optimization of such a treatment line and to give cost evaluation.
2. MATERIALS AND ME'lBDS pilot Plant The study is performed on a 450-l/h pilot. This pilot comprises: two
ozone contactors (150 mm in-diameter, 4 m in height), an ozone generator (Degremont, France) which produces up to 10 g/h, a dosage system for hydrogen peroxide which permdts the injection of a diluted solution in the input of the two columns, and two filters running in parallel (sand and GAC) after the oxidation step to reduce biodegradable oxidation by-products.
Analytical Methods Ozone concentration in the gas and dissolved ozone in water are
measured respectively by the KI and indigo methods (6). The concentration of H207._in the stock solution and the residual in
water are measured using an iOdometric method. Identification and dosage of the benzene derivatives were performed
using 2 types of liquid-liquid extraction. For concentrations above 10 pg/l, 1 liter of water adjusted to pH 2 was extracted by 100 ml of methylene chloride added in three times. Extract was concentrated to 1 ml using Dufton distillation columns. Two microliters were injected "on column" in a gas chromatograph (GC) (Varian 3500, USA) on a fused silica 50 m capillary column (Chrompack OV 1701). Detection and quantification were realized with a flame ionization detector. Identification was done by the use of a mass spectrometer (MS) (Finnigan ITDS, USA). For concentrations lower than 10 pg/l, a second technique was used. 1 liter of water was extracted by small volumes of benzene (2 to 5 ml). The extract was injected in GC equiped with an electron capture detector to reach sensitivity limits of 0.1 to 0.05 pg/l.
Sensory analyses used to evaluate taste and odor characteristics were threshold odor number technique (roN) and flavor profile analysis (FPA) (7,8) which gives a better description of organoleptic properties present in the sample.
The optical density at 254 nm is traced using an automatic spectrophotometer developed by the Lyonnaise des Eaux Laboratory (9).
3. RAW WATER OJARACTERISTICS The pilot plant is fed by groundwater from a well. The main
characteristics of the water are reported in Table I. The presence of low concentrations of nitrite, iron, and manganese, as well as the moderate alkalinity, are noted. The bicarbonate concentration is one of the main parameters that affect the oxidation efficiency (10).
Table II gives the evolution of pollutant concentrations in the groundwater from January to March 1987 :
- the total concentration of benzenic compounds is in the range of 1400-2500 pg/l, with an average value equal to 2100 pg/l ;
- the major pollutant (representing 70% of the pollution) is o-chloronitrobenzene (1000-1800 pg/l - average concentration: 1500 pg/l) ;
- some other pollutants, such as chlorobenzene and nitrophenol, are present in the groundwater at variable concentrations depending on the sampling period.
During this period roNs in the fed water were above 20.
4. FEASIBILITY OF DE'roXIFICATI~ BY AN O~~2 <DJBINATI~ IntrOdUction Ozone has been used for many years in drinking water treatment to
remove color, control tastes and odors, inactivate microorganisms, and for various other purposes. Because of its high oxidation potential, ozone is also able to oxidize organic micropollutants in water. But depending of on
- 3()()-
their nature, oxidation kinetics can vary widely. Boigne and Bader (10) report the reaction rate constants of ozone with different substituted benzenes, from which it can be concluded that the nitro and chloro substituents reduce ozone reactivity. Therefore, the high concentrations of chloronitrobenzenic compounds contained in the groundwater seem to be more refractory to ozone. To detoxify such a groundwater, the very reactive and oxidizing hydroxyl radical (08°) can be used. This radical can be produced in particular by combining ozone with hydrogen-peroxide (11) or with ultraviolet light (12). The first combination, which is tested in this study, has been found to be an attractive treatment for the removal of chlorinated solvents from groundwater (9). For a given raw water, the parameters which are responsible for the oxidation efficiency are the rate of radical production and the contact time. So the parameters which need to be optimized are the ozone and hydrogen peroxide dosages and the contact time. For plant design, two other parameters must be taken into account the maximal concentration of pollutants and the alkalinity of the water. The hydroxyl radicals (OBO) are scavenged by the bicarbonate ions, therefore the oxidation efficiency depends on the pollutant concentrations/alkalinity ratio.
Results Two types of tests are performed to verify the efficiency of ozone
alone and ozone combined with hydrogen peroxide. Ozonation The results obtained by ozonation alone are summarized in Figures 1 and
2. The tests show that for a contact time of 30 minutes, the decrease of the optical density at 254 nm (which corresponds to the decrease of aromatic compounds concentrations) is a function of the increasing applied ozone dose. At an ozone dose of 16 g/m3, the removal of chloronitrobenzene reaches 85%. With a constant ozone dose, the reduction of the optical density at 254 nm increases when the contact time is longer, up to 30 minutes. The application of 16 9 ~l/m3 with a 40-min contact time allows a 88% reduction of chloronitrobenzene corresponding to a residual concentration of 150 pg (o-chloronitrobenzene)!l in the ozonated water.
Oxidation by ozone alone, even for large doses of ozone and long contact times, does not allow a sufficient reduction of the chloronitrobenzene concentration, which remains higher than the desired amount of 20-30 pg/l in the ozonated water.
Combination of ozone with hydrogen peroxide The 03/H7.02 combination is performed in two columns placed in series.
Operating parameters are the same in each column. Previous experiments reported in Figure 3 show that the best o-chloronitrobenzene removal is obtained for a hydrogen peroxide/ozone mass ratio of 0.4. The application of ozone and hydrogen peroxide dosages of 8 g/m3 and 3 g/m3 , respectively, during a 20-minute contact time permitted a 99% elimination of benzenic compounds (Figure 4). An increase in the contact time to 30 minutes does not improve removal efficiency.
5. EFFICIENCY OF "mE CCJIIPLETE TRFA'H'tENl' LINE It is well-known that during the oxidation process biodegradable
compounds are formed which can induce bacterial regrowth in the distribution network. To avoid this, the oxidation process must be followed by a biological treatment such as sand or GAC filtration. The latter technique has a distinct advantage in that it removes micropollutants by adsorption. For the study presented, fixed beds GAC and sand columns with 15 min hydraulic residence times are used. All of the treatment lines run for 3 months. Samples are collected weekly at different treatment steps. The mean concentrations of benzenic compounds and o-chloronitrobenzene in the water after the oxidation step are 22 and 16 pg/l respectively, corresponding to a 99% reduction. TOC removal obtained by oxidation is
- 301-
about 0.3 mg/l; this corresponds to the mineralisation of 30% of the benzenic compounds (Table III). Identification of oxidation by-products (e.g aldehydes, carboxylic acids) will follow at a later stage of the study. After GAC filtration, the treated water has a benzenic compounds concentration lower than the detection limit (1 pg/l). From raw water to oxidized water, the threshold odor number (TON) is reduced from 20 to 3, and there is no taste and odor in the GAC filtered water after 3 months of operating time. Therefore, this treatment line is very efficient and results in a high-quality drinking water. If the technical feasibility of such a treatment line can be demonstrated, investments and the operating costs of a new plant can be evaluated.
6. INVES'l1IIENTS AND OPERATIKi OOSTS As previously shown the oxidation efficiency is dependent upon the
organic matter concentration and alkalinity. The drinking water treatment plant uses different waters from the wells. These waters have different alkalinities and pollutant contents (120 to 230 mg caoo3/l). Therefore, it is a necessity to evaluate the increase of ozone anO hydrogen peroxide dosages and resulting changements in costs for various concentrations of benzene derivatives and alkalinity. In Table IV, both investment and operating costs are calculated in the case of a 30,000 m3/day plant, for a removal of 500 to 5000 pg/l of ortho- chloronitrobenzene (o-CNB). The investment costs for oxidation process vary from 9,5 Millions FF to 33 Millions FF, with 12 Millions FF for a 5 m/h GAC filtration, total investment cost is between 21,5 and 45 Millions FF. on the other hand, operating costs for the oxidation process and GAC filtration are varying from 0.09 to 0.49 FF. In these cases, ozone and hydrogen peroxide dosages are between 3.5 to 32 glm3 , and 1,5 to 13 g/m3 , respectively. Table V gives an evaluation of the ozone dose as a function of the initial o-CNB concentration and alkalinity. Results show that two-fold increases in applied ozone dose are needed when the alkalinity varies from 120 to 230 mg/l as caco3.
7. CCH:LU5ICN Taste and odor control can be used to detect pollution of water by
industrial waste. For example, the use of the flavor profile analysis combined with the close loop stripping analysis has permitted to detect the pollution of a groundwater by chloro- and nitrobenzenic compounds in a range of 10 to 20,000 pg/l. The use of such a groundwater for the production of potable water requires a detoxification step. A promising way for the removal of such compounds is oxidation using ozone. Tests performed using ozone alone showed that even large ozone dosages do not allow to reach a satisfactory level of removal.
on the other hand the use of powerful oxidant species such as hydroxyl radicals which are formed by the combination of both ozone and hydrogen peroxide has permitted to obtain a water containing a concentration in benzene derivatives lower than 20 pg/l. This concentration corresponds to the odor threshold concentration of the main pollutant, the o-chloronitrobenzene. Furthermore this treatment reduces drastically the TON measured on oxidized water. The feasibility of such a treatment is demonstrated using ozone and hydrogen peroxide dosages of 8gjm3 and 3 g/m3, respectively, with a 20-minute contact time. An evaluation of the costs of such an effective treatment for the specific case studied showed, that investment and operational costs were of the same order than construction of new wells at another location and transportation of the water.
This oxidation step can be economically used to reach a satisfactory oxidation at reduced ozone dosages followed by an artificial groundwater recharge which can lead to a biological removal of pollutants and their by-products. Moreover, the oxidant dosages can be adjusted as a function of
- 302-
the degree of pollution. Monitored for example by a UV spectrophotometer, this can lead to very flexible operating costs. In the case of a large decrease of the pollution to be treated, hydrogen peroxide injection can be suppressed, resulting in a classical treatment and a minimization of operating costs.
8. REFERENCES
(1) C. Anselme, K. N'Guyen, J. Mallevialle, J.P. Bordet. Influence des traitements de desinfection et d'oxydation sur les qualites organoleptiques de l'eau : cas de l'usine de Morsang/Seine. Paru dans les proceedings de la 38eme Journee Internationale du CEBEDEAU, Bruxelles, 10-12 juin 1985. .
(2) J. Mallevialle, A. Bruchet, E. Schmitt. Nitrogenous Organic Compounds: Identification and Significance in Several French water Treatment plants. Presented to the WQTC, Norfolk, Virginia, Dec. 4-7, 1983.
(3) W. KUhn, D. Clifford. Experience with Specific Organic Analysis for water Quality Control in west Germany. Published in AWWA Technology Conference Proceedings: Advances in water Analysis and Treatment. Presented at the AWWA water Quality Technology Conference. Houston, Texas, Dec. 8-11, 1985.
(4) E.J. Fairchild. Suspected Carcinogens : a Sourcebook of the Toxic Effects of Chemical Substances. Castle House Publications Limited, Londres.
(5) L.J, Van Gernert, A.H. Nettenbreijer. Compilation of Odor Threshold Values in Air and water. Nat. Inst. Font. wt. Supply, Voorburg, Netherland and Centro Inst. for Nutr. & Food Res. TNO, Zeist. Netherland, 1977.
(6) H. Bader, J. Hoigne. Determination of Ozone in water by the Indigo Method. A Submitted Standard Method. Ozone Sci. Eng. 4, 169-176, 1982.
(7) S.W. Krasner, M.J Mc Guire; U.B Ferguson. Application of the Flavor Profile Method for Taste and Odor Problems in Drinking water. Presented at the WQTC, Norfolk, Virginia, Dec. 6, 1983.
(8) J.P. ouguet, E. Brodard, M. Roustan, J. Mallevialle. The Development of an Automated procedure and the Applicability of this Procedure for Monitoring the Effectiveness of Ozone. Ozone Science and Engineering, 8, n0 4, 321-338. To be published in Analysis of Ozone in water and wastewater Treatment Manual, R.G. Rice, Editor (1986).
(9) W.H. Glaze, J.W. Kang, M. Aieta. Ozone-Hydrogen Peroxide Systems for Control of Organics in Municipal water Supplies. Presented at the 2nd International Ozone Conference, Edmonton, Canada, April 27-29, 1987.
(10) H. Hoigne, H. Bader. Rate Constants of reactions of Ozone with Organic and Inorganic Compounds in Water-1, Non-Dissociating Organic Compounds. water Research, 17, 173-183, 1983.
(11) J.P ouguet, E. Brodard, B. oussert, J. Mallevialle. Improvment in the Effectiveness of Ozonation of Drinking water Through the Use of Hydrogen Peroxide. Ozone Sci. Eng. 7, 241-258, 1985.
- 303-
(12) JP Duguet, C. Brossard, E. Brodard, J. Mallevialle, M. Roustan. Development in Ozonation Techniques using combination of Ozone and Ultraviolet Light. Presented at the 8th Ozone World Congress, Zurich, Switzerland, September 15-18, 1987.
TABLE I. Main Characteristics of Ground water
Resistivity (S2/cm) 1695
pH 6.7
KMn04 oxidability (mg/l) as oxygen
1.05
TOC (mg/l) 1.9
Alkalinity (mg/l as caC03) 225
Chloride (mg/l) 88
Sulfate (mg/l) 58
Nitrate (mg/l as N03) 23
Nitrite (mg/ as N02) 0.30
Iron (.ug/l) 50
Manganese (.ug/l) 30
Magnesium (mg/l) 28
Calcium (mgjl) 71
- 304-
TABL
E II
. E
vo
luti
on
of
Ben
zen
e D
eriv
ati
ves
Con
cen
trat
ion
(ll
g/l
.) in
Gro
undw
ater
-
(Jar
ruar
y-M
arch
198
7)
COM
POUN
DS
1/8
1
;22
1
/27
2
/3
2/4
2
;20
3
/3
3/1
1
3/1
7
3/2
4
3/3
1
Ch1
orob
enze
ne
165
235
150
640
540
34
67
85
85
66
145
o-D
ich
1o
rob
enze
ne
< 1
2
2 2
3 3
4
Ben
zen
ic c
ompo
und
13
5
25
21
21
24
18
26
24
27
31
Nit
rob
enze
ne
7 9
7 5
10
5 7
7 3
12
Nit
rop
hen
ol
280
300
17
96
98
187
o-C
hlo
ran
i lin
e
< 1
11
23
18
4
4 7
4 2
o-N
itro
tolu
en
e
145
152
136
120
132
120
90
165
135
125
134
w ~
m-N
itro
tolu
ene
14
15
15
12
13
14
9
17
17
16
19
I
p-N
itro
tolu
en
e
20
11
12
8
3 8
7 10
14
m-C
hlo
ron
itro
ben
zen
e 11
1 10
0 10
3 72
73
81
80
10
6 15
0 17
9 11
3
p-C
hlo
ron
itro
ben
zen
e 12
3 5
42
39
o-C
hlo
ron
itro
ben
zen
e 16
10
1580
15
60
1300
14
60
1690
97
0 18
28
1490
13
20
1402
Dic
hlo
roan
ilin
e
6
Met
ho
xy
nit
rob
enze
ne
7 7
6 5
5 6
5 5
8
Dic
hlo
roan
ilin
e
6 6
6 5
12
3 4
2 3
Tri
ch
loro
an
ilin
e
5
Din
itro
tolu
en
e
3
'TO
TA
L 25
01
2420
20
36
2200
22
72
2023
13
97
2247
20
35
1750
21
21
TABLE III. Evolution of 'lOC, Benzenic CoIJIIOunds, crchl.oroni trobenzene aJXl 'lhreshold Odor Number
at Different Treatment steps
RUNNING TIME (Days) 0 4 11 19 25 32 39
RAW WATER roc ( mg/1 ) 1.9 1.9 1.95 1.75 1.9 Benzenic compounds (pg/l) 2023 1397 2249 2031 1750 2121 o-ch1oronitrobenzene " 1690 970 1830 1490 1320 1402 TON > 20
OXIDIZED WATER roc ( mg/l) 1.7 1.5 1.75 1.6 1.3 1.45 -Benzenic compounds (pg/l) 52 6 16 21 18 19 23 o-Ch1oronitrobenzene " 41 4 13 14 14 14 14 TON = 3 to 5
SAND-FILTERED WATER roc ( mg/l ) 1.4 1.55 1.45 1.55 1.15 1 Benzenic compounds (pg/1) 17 6 11 35 17 11 14 o-Ch1oronitrobenzene " 15 3 8 28 11 8 9 TON=2to3
GAC-FILTERED WATER roc ( mg/l ) 0.9 0.40 0.40 0.35 0.45 -Benzenic compounds (pg/1) < 1 < 1 < 1 < 1 < 1 < 1 o-Ch1oronitrobenzene " < 1 < 1 < 1 < 1 < 1 < 1 TON = 1
TABLE IV. Evaluation of the Investment aJXl the Operating Costs as a Function of the Initial o-Chloronitrobenzene Concentration
Orthochloronitrobenzene to be removed 500 1000 1500 3000 5000
(pg/1)
INVESTMENTS (30000 m'/j)
1) Oxidation Ozone gjm' 3.5 7 10 20 32 H ° gjm' 1.5 3 4 8 13 c~neact time min 20 20 20 20 20
cost 106 FF 9.5 12 15 23.5 33
2) GAC Filtration (5m/h)
Cost 106 FF 12 12 12 12 12
TOTAL 106 FF 21.5 24 27 35.5 45
Ol'ERA'rJNi COSTS
Ozone FF/m' 0.035 0.07 0.10 0.20 0.32 H20Z 0.017 0.035 0.04 0.09 0.14
GAC 2 years 0.035 0.035 0.03 0.03 0.03
TOTAL 0.09 0.14 0.17 0.32 0.49
- 306-
TABLE v. Evaluation of the Ozone Doses as a Function of the initial o-Chloroni trobenzene Coocentration and Alkalinity
120
160
250
OCNB pg/l
500 1000 1500 3000 5000
2.7
3.5
5.5
5.5
7
10
- 307-
8
10
15
16
20
29
27
32
43
80
fP. 75
iii > 0 E 70 Q) ... ..r LO N 65 c
0
60 ~.
OZONE OXIDATION
Fig.1. In·fluence of the ozone dosage (Contact time: 30min)'
1 (o-CNB) c 1300J,lg/l
Raw water (p-CNB) : 27J,1g/1
80
fP. 75
iii > 0 E 70 Q) ... ~ LO N 65 c
0
~ ~ V ./ VO l o-CNB = 208J,1g/ i
Y p-CNB : 8~g/ l
V I
60 I---
8 10 12 14 16
03 (91m3 )
OZONE OXIDATION
Fig.2. Influence of contact time - 0 3 = 16 g/m 3
\ (o-CNB) = 1300 ~g / t Raw water I (p-CNB) = 27 J,lg / l
v---b---
/ ~-CNB=182J,19 / 1
-CNB = 7~g/t
7 /
10 20 30
Contact time (min)
- 308-
lo-CNB: 150~g / l rp-CNB= 5~g / t
40
90 '200
~ 85 150 'iO
~
> '-0 Cl
E ;;)
41 100 ... III q- 0 Z Il'l ()
N I
0 0
0 50
• o-CNB 70 0
0 0.4 O.B 1.2 H202 / 03(W/ W)
Fig.3. Influence of H202 dosage on the oxidation efficiency (ozone dosage, 12g/m3 jcontact time,20min)
85 300
~ \ ... °°254
'-
reNBfDD2S' iii 75 , ~/ 200 Cl
\ I .3 > 0 I t E CII
, I x ... , I q- '4 Il'l Contact time N ~ g 65 --- 2x 1 Omin
1\ - 2X'15min 100
I , \~\:NB I
I -, I I .
55~--------~--------~----------J0 o 5 10 15
03 (91 m3)
Fi9.4. Removal of 00254 and o-eNS by a double 03/H 20 2 injection. Influence of the ozone
dose and contact time (H202/03 : 0.4 W/W
[o-CNBJ o = 1800~g/l)
- 309-
III z ()
I 0
POSTER SESSION IV WATER TREATMENT
Possibilities and limitations of the combined use of ozone and hydrogen peroxide in drinking water preparation from surface water
Presence of polycyclic aromatic hydrocarbons in surface waters used for the production of drinking water
Research and behaviour of organic micropollutants from waste distillery wine in anaerobic treatment
Mass spectrometric identification of halogenated surfactants in Barcelona's water treatment plant
Effects of chlorine dioxide preoxidation on organic halide formation potentials
NMR study of kraft pulp mill waste and natural humic substances
Identification of bioaccumulable compounds in kraft bleaching effluents
Influence of humic water substances on the degradation of PAH during water chlorination
Influence of waste water disinfection treatments on some genotoxic chemical micropollutants
POSSIBILITIES AND LIMITATIONS OF THE COMBINED USE OF OZONE AND HYDROGEN PEROXIDE IN DRINKING WATER PREPARATION
FROM SURFACE WATER
Summary
---_._--F. VAN HOOF, J. JANSSENS and E. PLUYS
Study Center for Water c/o Antwerpse Waterwerken
Mechelsesteenweg 6~ 2018 Antwerp - Belgium
The feasability of the combined use of ozone and hydrogen peroxide in surface water preozonation and in the ozonation of pretreated water was evaluated in relation to H202 dosage, time of H202 addition, 03 dosage TOC, UV25~ absorption, formation of oxidation by-products, trihalomethane precursor removal and adsorption on activated carbon in comparison with the use of ozone as such. Although in all cases a small but consistently better UV25~ removal was observed though the combined use of ozone and hydrogen peroxide, the beneficial effects for other parameters were limited or non existent, especially within the range of ozone doses applied during surface water preozonation « 2 mg/L). Therefore, the combined application of ozone and hydrogen peroxide seems to have little advantage over the use of ozone alone in the treatment of this type of surface waters.
Introduction
The reactions of ozone with impurities in water are twofold : 1) a direct reaction which is relatively slow and selective. 2) decomposition of ozone, catalysed by OH-ions, proceeding more rapidly with increasing pH and which is accelerated by a radical-type chain reaction in which free radicals (e.g. OHo and H02°) act as carriers. This type of reaction shows a more aselective character. (1)
A t several occasions it has been tried to take advantage of the strong oxidative character of the free radicals to attack species which are restistant to ozone as such. The combination of ozone with UV-radiation and ozone with hydrogen peroxide are the best known approaches for the induction of the second pathway. The reaction between ozone and hydrogen peroxide is strongly pH dependent, being much faster at higher pH values with a rate increase of one order of magnitude per pH unit. The reaction proceeds as follows :
H02- + 03 -+ OHO + 02'- + 02 The radicals formed react with organic matter, but also with ozone :
OHO + 03 -+ H02° + 02 or are scavenged by carbonate, bicarbonate or phosphate, which do not act as chain carriers (2).
The applicability in drinking water treatment of the ozone-hydrogen peroxide combination has to some extent been evaluated, mainly in relation to ozone doses applied which are very high (3,~). One of the
- 312-
aims of this study was to find out how the ozone-hydrogen peroxide combination performed when applied at more realistic ozone doses on surface water having bicarbonate levels of approximately 2mmole 1-1.
Experimental
Ozonation of different water was performed in batch experiments. The reactor consisted of a modified gas washing bottle equiped with a sintered glass dispersion floor delivering ozone to the bottom of the reaction ¥essel. Excess ozone passed out through the top of the reactor into a gas washing bottle containing 2 % KI solution buffered at pH 7.0
Most experiments were carried out on river Meuse derived surface water after impoundment and to a lesser extent on treated water from the same origin (after prechlorination, coagulation and double layer filtration).
The TOC levels of the surface and pretreated water were 3-6 and 3-4 mg C CI respectively. The bicarbonate levels were 2mmole (1 approximately.
H202 concentrations were assayed spectrophotometrically as described by Masschelein et al. (5). Linear aldehydes were quantified through HPLC with UV detection at 365 mm after derivalisation with 2,4 dinitrophenylhydrazine (6).
Trihalomethane formation potentials (THMFP) were measured after 72 hours contact times at 25°C and pH 8.
Adsorption isotherms were determined through the bottle point method with varying amounts of activated carbon (Chemviron F400) and contact times of eight days.
Results
1. Changes in UV 254 absorption
The addition of hydrogen peroxide alone does not alter UV254 absorption. The efficiency of the ozone - H202 combination was monitored through the evolution of the UV254 in comparison with ozone application as such. Favourable changes induced by ozone - H202 were seen only after application of H202 doses up to 0,25 mgtl. Higher doses did not results in greater decreases un UV254 absorption than ozone application alone. (Figure 1).
The influence of the time of injection of H202 was investigated, whether H202was applied before ozonation or after five minutes of ozonation did not induce significant differences in UV 254 absorbance after 15-20 minutes of ozonation (Figure 2).
In all cases investigated H202 consumption was lower than ten percent of the initial H202 dose applied.
2. Formation of oxidation by products
In spite of the very limited H202 consumption different oxidative pathways seem to be followed by the combination 03 + H202 than by ozone alone.
- 313-
The formation of low molecular weight aldehydes which is favoured by the action of ozone as such, as exemplified by the higher aldehyde levels produced at pH2, is much lower after application of 03 + H202 than after application of ozone alone (Fig. 3). This seems to indicate that either the hydroxyl radicals produced through the interaction of ozone and hydrogen peroxide yield less aldehydes, or that they are further oxidized more rapidly to their corresponding acids.
3. Influence on trihalomethane formation potential (THMFP)
Since the UV254 absorption has been shown to correlate well with THMFP, it was tried to find out whether differences in UV254 absorption between ozone and ozone + H202 treated waters were reflected in trihalomethane formation potentials. Both treated waters showed much lower THMFP than the corresponding raw waters, but differences among them were insignificant (Figure 4). H202 alone did not influence trihalomethane formation potentials.
4. Adsorbability on activated carbon
Adsorption on activated carbon (Chemviron F400) was evaluated after treating raw water with 2 mg/l 03 and 2 mg/l 03 + 1 mg/l H202 respectively. Adsorbability on activated carbon was somewhat higher after combined treatment in terms of total organic carbon : at Ce = 2,0 mg C/I, ~ valves of 35 and 22 mg C/g were found. These differences were
not entirely confirmed when adsorption isotherms were set out as a function of UV254 absorbance, in which case a marginally positive effect was observed in favour of ozone alone.
Discussion
Although the interaction between hydrogen peroxide and ozone is theoretically well documented, its applications in water treatment are related to the application of very high ozone doses, which are unlikely to be used in practice.
Although confirming the results of others in terms of the reduction of UV254 absorbance, the differences observed in this study after application of ozone doses which are typically being used in surface water ozonation are limited. As a consequence differences in trihalomethane formation potential are very narrow, in spite of the fact that the patterns observed in the formation of oxidation byproducts (linear aldehydes) point into the direction of different mechanisms of interaction with the organic material.
The presence of bicarbonate ions in the waters investigated may be a handicap for the development of oxidation processes through OHoradicals. It therefore seems worthwhile to study ozone-hydrogen peroxide combinations on surface-waters containing low concentrations of inorganic scavengers. On the typs of water used in this study there seems to be little benefit in combining both oxidants.
- 314-
References
1. STAEHELIN, J. and J. HOIGNE (1982). Decomposition of ozone in water : Rate of initiation by hydroxide ions and hydrogen peroxide. Environ. Sci. Techno!., 16, 676-681.
2. ST AEHELIN, J. and J. HOIGNE (1985). Decomposi tion of ozone in water in the presence of organic solutes acting as promotors and inhibitors of radical chain reactors. Environ. Sci, Technol., 19, 1206-1213. --
3. BRUNET, R., M.M.BOURBIGOT and M. DORE (1984) Oxidation of organic compounds through the combination ozone-hydrogen peroxide. Ozone: Science and Engineering, §., 163-183.
4. DUGUET, J.P., E. BRODARD, B. DUSSERT and J. MALLEVIALLE (1985). Improvement in the Effectiveness of Ozonation in drinking water through the use of hydrogen peroxide. Ozone: Science and Engineering, I, 241-258.
5. MASSCHELEIN, W., DENIS M. and LEDENT R., (1980). Spectrophotometric determination of residual H202' Water and Sewage works, 470, 1980.
6. VAN HOOF, F., WITTOCX, A., V AN BUGGENHOUT, E. and JANSSENS, J.G. (1985). Determination of aliphatic aldehydes in water by HPLC. Analytica Chimica Acta, 169, 419-424.
- 315-
0.10
0.09
0.07
0.06
0 .05
O.oJ
0.Q2
0.01
o
Figure 1
0.05
0 .03
0.02
0.01
o
Figure Z
----. O:! <>->-----<> O:! + 0.25mg/( H2~
"'"--..... 0 3 + 0.50 mg /I ~~
D "-- -C OJ + 1.00 mg /I ~02
6 12 mg II O:! applied
Changes In UV Z54 in function of 03 and HZOZ dose.
3.5 7.0 10.5 1'.0
&---...:..
mg II OJ applied
Ozone
1 mg H2~ r' + Ozone
Ozone (3.5 mg III + 1 mg H2~ • + Ozone
Evolution of UV Z54 in function of time of HZO Z injection.
- 316-
, aldehydes
0, ' 1.5 mgl!
O,lO 1 , pH 8
2 , pH 12
3 , pH 2
,----
0,20
0.15
r--
f---
1 , Unlr. olad
2 , 10mg II 0,
H.z,o.l. mg/lo,
J , lmg/l • 10
--
0,10
1
Figure 3
2 J ~ 2 3
Formation of linear aldehydes through 03 and 03 + HZOZ
T H M F P (11 mole I U
0 ,7 0
0,60 1 : RAW
2 : .2m 9 03 II
0,50 3 : + 2m 9 03 II + 1 mg ~02
0 ,40
0,30
0,20
0,10
1 2 3
Figure 4 Influence of 03 and 03 ~ HZO Z on THMFP.
- 317-
PRESENCE OF POLYCYCLIC AROMATIC HYDROCARBONS IN SURF ACE WATERS USED FOR THE PRODUCTION OF DRINKING WATER
Summary
F. V AN HOOF and S. AERTS Studiecentrum voor Water
c/o Antwerpse Waterwerken Mechelsesteenweg 64
2018 Antwerp - Belgium
Two methods widely applied for isolating PAH from surface water have been compared : solvent extraction using cyclohexane and solid surface extraction. Higher recoveires were obtained through solvent extraction, while the reproducibility for both methods was similar. In spite of the higher costs solvent extraction was selected for further use. Results obtained on surface water samples were strongly influenced by the presence of suspended material. PAH were shown to be mainly associated with the larger suspended particles which are removed easily through sedimentation in impoudment reservoirs. The results obtained stress the need for clearly identifying the sample type (filtered or not) and the sampling point (before or after impoundment) when performing this analysis for evaluating water quality in relation to legal requirements.
Introduction
EEG directive 75/440 (Quality of surface water production) requires surface waters to be monitored for the presence of polycyclic aromatic hydrocarbons (PAH). For this purpose six compounds are taken into account fluoranthene, benzo(b) fluoranthene, benzo(k) fluoranthene, benzo(a) pyrene, benzo(g,h,i) perylene and indeno(l,2,3,c,d) pyrene.
Two methods which are widely used for this parameter have been evaluated solvent extraction with cyclohexane and solid surface extraction, especially with respect to the presence of suspended materials in the samples. In a second stage one of the methods has been used to study the effects of surface water impoundment on the presence of PAH.
Experimental
1. Materials
Acetonitrile, tetrahydrofurane (THF) and water were HPLC Grade. Cyclohexane (Merck 2828) for spectrophotofluorometry was used for solvent extraction. Glass fiber filters (Whatman) and 0,45 pm Teflon membrane filters (Millipore) were used for separation of suspended materials. Sep Pak C18 cartridges (Waters) were selected for solid surface extraction.
HPLC analyses was performed on Zorbax ODS and Vydac 201 TP columns, both 25 cm long and with 5 um particle diameter. Isocratic elution (90 % CH3CN - 10 % H20) was performed on the Zorbax ODS column. On the Vydac 201 TP isocratic elution (70 % CH3CN - 30 % H20 for three minutes) was followed by a linear gradient to 100 %
- 318-
eH3eN in seven minutes. 100 % eH3eN was on for ten further minutes. The PAH were detected with a fluorescence detector ( .A exc. 305 nm, .>.. emm. 4-30 mm). Reference PAH were obtained from Ferak (Berlin).
2. Sampling and sample treatment
Surface water samples were taken from the Nethe- and Albertcanal (derived from the river Meuse) and from impoundment reservoirs. After transportation to the laboratory, the PAH whe isolated the same day. Analysis using solvent and solid surface extraction were performed according to ISO DP 7981 [ISO/Te 14-7 / No 69 (Se2/ WG 19)].
Results and discussion
1. Recovery of PAH
Recoveries of PAH through solvent extraction were determined by adding 100 ng fluoranthere and 20 n g of the other PAH to one liter of Milli Q water. The recoveries encountered corresponded well with those mentioned in the ISO document (80 % for benzo(a) pyrene and 95-98 % for the other compounds), except for benzo(k) fluoranthene, which was recovered for 77,9 % only (S.d. = 4-,5 %, n = 5). Recoveries through solid surface extraction were generally lower (60 - 62 %) than through solvent extraction except for fluoranthene (95 %, s.d. = 12,3 %, n = 5). Although increased recoveries of PAH through solid surface extraction have been reported after addition of isopropanol, the data reported do not indicate higher recoveries than those found by ourselves or reported in the ISO document. Therefore and because of the fact that solvent extraction allowed larger volumes of unfiltered surface water to be proceeded, further work was done with solvent extraction.
2. Influence of suspended solids on PAH levels
In order to find out to what extent suspended solids influenced the levels of PAH encountered several samples of canal water were analysed before and after filtration over a glas fiber filter. Some results are brought together in figure 1. The influence of filtration before analyses has a strong but varying impact on the results obtained. Reductions up to 78 % in the sum of the six PAH are seen after filtration. The influence of filtration on fluoranthene levels is less unambiguous : in one case a higher level was seen in the filtrate, in two other cases a decrease was observed, which was however smaller than the decreases for the five other PAH. Although varying strongly from sample to sample reductions were very similar for different PAH in each sample : 4-5,8 -50,6 % (A), 77;0 - 80,5 % (B) and 91,7 - > 97,0 % (e).
3. Evolution of PAH in impoundment reservoirs
The evolution of the PAH levels was followed at a water intake and in four reservoirs, operated in series and with a residence time of approximately eight days. The evolution of the PAH is represented in figure 2. Strong decreases in PAH levels are observed within the two first reservoirs in which the major part of the suspended material sedimentation occurs. Strong correlations were observed between turbidity
- 319-
and PAH atthe water intakes ( r = 0,893) and between turbidity and PAH within the different reservoirs (20/1 : r = 0,981; 12/5 : r' =0,996; 11/6 : r = 0,952; 1/7 : r = 0,944). These data suggest that PAH are mainly associated with the larger particles, undergoing sedimentation in the first reservoirs. In order to confirm this hypothesis a canal water sample was analysed as such, after filtration over a glas fiber filter and after filtration over a 0,45 fm membrane filter. Results are brought together in Table I.
TABLE I
PAH in canal water before and after filtration
A B C
Fluoranthene 160 60 20 Benzo(b) fluoranthene 60 5 2 Benzo(h) fluoranthene 18 0,1 n.d. Benzo (a) pyrene 44 1,6 n.d. Benzo(g,h,i) perylene 12 n.d. n.d. Indeno{1,2,3,c,d) pyrene 16 n.d. n.d. Total 31O 68,4 22
All figures in ng/l n.d. not detected
A without filtration B : filtration over glass fibre filter C filtration over 0,45 fm membrane.
The data indicate that only a minor part of the PAH are in solution or bound to colloidal material (+ 7 %) while only 22 % passes the glass fibre filter. in both cases fluoranthene represents 90 % of the PAH in the filtrate.
4. Presence of other P AH
Several canal water samples were shown the abovementioned compounds. Fluorene, antracene, pyrene and benzo(a) antracene were occuring substances. The significance of their presence of six compounds quantified until investigation.
to contain other P AH than acenaphtene, phenantrene, among the most frequently presence in relation to the now is at present under
1. VAN NOORT, P.C.M. and E. WONDERGEM, (1985). The Isolation of some PAH from aqueons samples by means of Reversed Phase Concentrator Columns, Analytica Chimica Acta, 172, 335-340.
2. VAN den HOED, N.; M.T.H. HALMANS and J.S. DITS (1982), Determination of PAH at the low ng/l level in the BIESBOSCH water storage reservoirs for the study of the degradation of chemicals un surface waters in : Analysis of Organic Micropollutants in Water, Proceedings of the 2nd European Symposium, Killarney, 17-19/11/1981. A. Bj9Srseth and G. Angeletti eds., pp. 188-192.
- 320-
ng/l
400
300 0 sum os six PAH B.F. IZl sum of six PAH A.F.
e fluoranthene B.F. 200 m fluoranthene A.F.
• six PAH - fluor B.F. III six PAH - fluor A.F.
100
B.F. before filtration
A.F. after filtration
0 A B C
Figure 1 Influence of filtration on PAH levels
ng/l
1000
800
II Intake 600
IB First reservoir
~ Second reservoir 400
D Third reservoir
Iii! Fourth reservoir 200
0 20/1 27/2 12/5 11/6 117
Figure 2 Evolution of PAH levels in impoundment reservoirs.
- 321-
ng/I
400
300
200
100
o
Figure 3
A : surface water
B alter filtration over glass fiber filter
C after filtration over 0,45 IJ.m membrane
m sum of six PAH mI f1uoranthene
A B c
Presence of PAH in unfiltered, glass fIber filtered and 0,45 JLIm membrane filtered samples.
- 322-
RESEARCH AND BEHAVIOUR OF ORGANIC MICROPOLLUTANTS FROM WASTE DISTILERY WINE IN ANAEROBIC TREATMENT
R. SANCHEZ CRESPO and J. PRADA ALVAREZ-BUYLLA Confederaci6n HidrogrAfica del Guadiana.MOPU. SPAIN
SUMMARY The waste from the industry of extraction of ETHANOL from wine and its sub-products, constitute a serius problem of contamination for the continental waters where they are -dumped. In this paper we present the results obtained from an investigation into the organic micropollutants in this waste, their evolution and behaviour, in a process of puri-fying trough accelerated anaerobic fermentation. The results show a high degree of purification which reaches, for some standard parameters, more than 95%, and for the overall total of the micropollutants reaches va-lues of 99,9%.
1. INTRODUCTION Some of the most contaminating organic liquid residue in
continental waters are to be found in the areas of "La Mancha" and "Extremadura", namely the products of the extraction of Ethanol from distillation of wine and its sub-products.
In t:1ese areas all the vines are "vi tis vinifera", of the varieties Airen, Cencibel, Pardillo, Jaen, Malvar, GarnAcha, -and other varieties in a lesser proportion. lhe production of wine in these areas is estimated to be 20.10 Hl., of which so me 6.106 Hl. are used for the production of Alcohol. -
Grape juice is extracted by passing the bunches of grapes through crushing machines, and them through automatic presses, making the juice flOW, and leaving in the presses a mixture of epicarps, pericarps, seed, and all the combination of stems which form the bunch; all this mixture is called "orujos".
The juice thus obtained is an aqueous solution with diver se cellulose-pectic materials in suspension, mucilages, gums,etc. among which are found to a great extent d-glucose, d-fruc tose, potassium tartrate, and to a lesser extent of pentoses.D-tartaric, l-malic and citric organic acids exist. Among the mineral substances we find the elements P, S, K, Na, Ga, Mg, -
- 323-
Si, Fe and Mn, and to a lesser extent F, CI, Br, I, AI, B, Ti, Rb, Mo and Co. As colouring, Anthocyanius, Flavones and Tan- -nins abound.
This juice immediately passes to the fermeJYJ.tation phase, which is possible due to the action of different yeasts of the type Saccharomyces, producing different phosphorilations in which the adenosinetriphosphoric acid plays a fundamental role being the principal produces of phosphoric acid, for the phosphorilation of the glucose.
The hexosediphosphoric acid thus formed, by the action of hexoquinase, is transformed into glyceraldehyde-3-phosphoric and dioxiacetonphosphoric acids
these last two acids are in equilibrium through the action of isomerase, and are displaced in the direction of the former.
The next stage consists of the oxidation of the glyceraldehydephosphoric by a specific ferment whose active group is Cozimase (Co), the schematic reaction being as follows:
E20 3P • OCH2.CHOH. CHCOH2 ~H~:~~_ E203P.OCH2.CHOH.COOH ~H~~~3- H203P.OCH2·CHOH.COOH
The ~ -phosphoglyceric acid thus formed is converted by the action of the enolase ferment into phosphopyruvic acid which, by saponifycation becomes pyruvic acid.
9H20.P03H2 CHOH COOH
9H20H CHOP03H2 COOH
-H 0 SH2 fH3 2 COP03H2 - CO
enolase COOH COOH
This last, through the action of the carboxylase ferment, separates into acetaldehyde and C02
The acetaldehyde is finally reduced to ethanol, the two atoms of H necessary, being provided the dihidrocozymase, beco ming cozymase capable of producing new deshydrogenations.
These reaction are only important stage in the multiple reactions which take place in the fermentation of grape juice.
As a consequence of these reactions, wine is made by the settling and decanting to eliminate the sediment, and by the
- 324-
many cares of the wine-makers. Ethanol is extracted by distillation in columns full of
Raschin rings, effecting in this way a rectification, giving alcohols of the binary azeotrope water-alcohol of nearly 96Q.
The prime material is wine, and to a lesser extent the "lias" and "orujos", previously described, conveniently fermen ted and submitted to diffusion. -
The residual waters of these processes have high temperatures, heavy cloudiness and a strong reddish colours, with con centrations between 5 and 20 gil.
2. PURIFYING TREATHENT The purification of these residues is normally carried
out with one of the following processes. a) - Neutralization with lime, flocculation, sedimentation,
centrifuge and the remaining liquid processed biologically with active sludge and surface electromechanical aereators.
b) Natural evaporation in lagoon, using the resulting solid as a fertilizer.
c) As an alternative an accelerated anaerobic digestion is used, with an adequate pH and temperature.
In a plant of this latter type, we s tudied the organic mi cropollutants which exist in this type of residual waters, ge: nerically named "vinazas", their evolution and behaviour in the different stages of purification of t hi s type of treatment.
The plant selected for this study is in Tomelloso (Ciudad Real, SPAIN), which processes some 370.000 m3/year of "vinazas" with a contaminating potential of 11.106 Kg. C.O.D./year. The principa l parameter of the plant are:
storage and homogenization 600 m3 Anaerobic digester 2 x 4.000 m3 Settling tank
Surface 123 2 m Volume 363 m3
Production of Biogas 15. 900 m3/day 9.200 Kg/day fuel
-FLOW DIAGRAM AND SAMPLING STATIONS
G r--t:]-- '---1 , Ie , --(!)--, r-- J : • r--@---r-- ... .., I I I ~G.>:! I
: ~-----: , rllf"==E==-1ll] L<D-"' L ___ t:j__ r---~ I , C
t ~--. I
L _________ _ ~~-~=~~-_-_j78 0 ~
- 325-
A _ " i x i"~ ond pH control
8 _ Na OH to ••
e _ Ano.rob ic Dil •• t. , o _ Ct ntrifulot ioft
E _ Settlin l 'tonlll;
1 - "Lio,-I"'. t 2 _IIIVlnolo,- . .. I . t
5 - A.a.,obl. outl.t
4 - Fln. ol efflu.nt
3, SAMPLING AND ANALYSIS METODS The samples, ponderly composite, were taking in bottles
of borosilicate glass of 2,5 1, capacity, kept refrigerated at 4QC,
The determinations of organic micropollutants were carried out by GC-MS, according to the following scheme:
ANALlTICAL TECHNIQUE
Adjull to pH 12 wilh No OH 6M Extract with CHt cr. ,30 +30 +30 ml
Adjusl to pH2 with HC r 6 N Extract with CH, cr" 30+30+30 mL
Capillary Column, S.E - 54,0'2 mm., 25 m.- He. Acid - 60' C. Isol.rmal- 2 min, 8' CI min to 250,10 min. N/B-60'C. l,ol.rmal-2min, 8' C/min 10 300,10min
PURGE. H •. - 40 mL / min. 12 min. TRAP SORBENT. TENAX
DE SORP'ION. 4 min. 180· C
Column,Glass 6'x 1/4", 1% SP 1000 on Carbopack B 4S-C. IsotfJ~mQI"'5min, S·C/min to 220·C,5min
{ Quad,upol. Mess Sp'H:lro~atr •. H P 5993 C
EQU!PMENT f/.~PlOYED Pu". af'd f,ot' HP 7675A llbrarv OqC N 2 S
- 326-
4. ANALYTICAL RESULTS The most significant general parameters, are the follo- -
wing:
SAMPLES NQ PARAMETER 1 J 2 I 3 I 4
UNIT
pH 4,9 4,9 7,3 8,0 pH Conductivity (20Q C) 9.200 3.010 5.450 5.920 us/cm Suspended Solid 1l0Q 15.220 866 1.080 355 mg/l Suspended Solid 600Q 2.575 130 292 255 mg/l C.O.D. 2h 16.000 10.520 1.000 900 mg/l O~ B.O.D. 5d 6.200 4.370 300 210 mg/l 0 U.V. Index 9.520 3.120 2.980 2.740 -
The organic micropollutants that we have characterized re present 75% of the separates in the capilar column that we ha: ve used, although we estimate, quantitatively, their overall significance in more than 95%. The results are show in the annex tables.
5. CONCLUSIONS The analytical results show us that in this type of resi
due, alcohols are predominantly found, especially phenyletha-nol, the organic acids and their sters. All these compound are totally eliminated in the purification process under study. The process is efficacious in the elimination of nitrogens and sulphurous compounds. The appearance of phenols as products of the treatment, introduces new elements as a basis for the study of the reactions which take place in this accelerated anaerobic fermentation, which is the basis of the treatment under study.
The efficacy of the purifying system under study is very high, reaching values higher than 99% for the organic micropollutants.
- 327-
;
WASTE DISTILLED WINE ANAER\J!,HC T')IGESTION
TOI1ELLOSO (CIUDAD R8AL) - SPAIN
IDJ;NTIFIED COMPOUND 5
N A H E I FOR M U L A
HALOGENS f ", DICHLORO. 1 f PROPErlE C)H4Cl ?
" ~ DIMETHYL, 4, C"LOROBENZENE Ca H7Cl 'f CHLOROl 1 f 3.5, TRIMETH'fLBENZENE CgHuCl
Total
ALCOHOLS 1·3, BUTAtlEDIOL C4H' 00 2 CYCLOHEXANOL C6 H,,0 PHENYL'IETHANOL C7 11 a O PHENYL ETHANOL CaH, 00
Total
- S:;OLS PHENOL CM'EO , IIETHYLPHENOL C7 HSO 4, ETIlYLPIlENOL Ca"'00 4 , PROPYLPHENOL CgH120
Total
CARBOXYLIC ACIDS PROPIONIC C3"6 0 2 BUTYRIC C4"e 0 2 PEI/TANOIC C5"1002 HEXANOIC C6 H,,0? HEPTANOIC C7 H14O, PHEI/YLACETIC C8H80~ P'lEt/YLPROPIONIC Cg H100 , PHENYLBUTIRIC C'OH'202 DECANOIC C'0H200 2
Total
ESTERS ETHYLOCTA!iEATE C'0H200 2 DIETHYLFl-fl'ALATE C"H'40 4 ET!lYLC/,FP.ATE C, ,1-f~40, ETHYLLAURATE C~4!!2002 ETIIYLFA~I1ITATE C'8113~C~ ETIIYLOLEATE C?O"380~ ETHYLSTEARATE C2oJi4002
Total
NITR(J(lENS " METHYLPYRIDINE I:{;H1" 1, METHYL, " NITROBENZENE ~H1NO llI, INDOLE, 6 ,METHYL CaH9 11 'H, INDOLE. 3,ETHNOL C'0H"NO N, ('.PHEIIYLETHYl.). AC: rA MIDE Cl0H13110
Total
OTHERS II, N' .DIMETHYL'!"·- :JREA C3Ha N,S o. (ETHYLTHIO) -1AljOL C4111COS TIll AZCL. 5. ET!· t CS!171l09 1. "EXENE Cf ih~
1 ~, 4. TRlMETt-iy:. r[;f;ZE~IE CgH,? 4, HETIIYL.DII'ENZO, fURAN C,3 ii ,OO
Total
TOTAL IDE:ITrFn:D CotH'OUNDS
- 328-
SANPLING STATION
,. "Lias" inlet :>. "Vinazas" inlet 3. Anaerobic outlet 4. Final effluent
C ONCENTR A T ION S .uglL-SAl-1PLES HI
1 I 2 I 3 I 4
- 1a '3 23'5 ?3'9
- - . 3'4
- - 13'5 4'2
- 1a' 3 37'0 31, 5
30'4 37'S - . - - - 16
67a ?a 'g - -92,434 24,700 - -g3,14?'1 24,760'4 - 16
- - 47'8 4'8
- - 965 51'~
- - 211 10"
- - 30'7 -- - 1,256'5 66'4
10,143 1.380 - -95. 958 8.9R9 - -
499.687 2'.7?? - -- 3.166 - -7.458 2.675 - -1.582 - - -7.94a 616 - -- 163 - -- 55' - -
622.776 40.263 - -
- 56 - -'90 1?0 - -- 58 - -- 42 - -- 637 - -- 7?8 - -- 175 - -
'90 1.816 - -
- - 4'9 10'5 23'6 - - -- - - 46'S
,. '73 926 - -375 - - -
P. '" '6 9'6 4'9 57
364 - - -313 - - -- 95 '7 - -119 10'2 - -- - - 9'8
- 44-3 - -796 '50' , - 9'8
67.939'9 1.298'4 180'7 ;
MASS SPECTROMETRIC IDENTIFICATION OF HALOGENATED SURFACTANTS IN8ARCELONA'$ WAI ER. lREATMEN1 PLANr
J. RIVERA (1), F. VENTURA (2), J. CAIXACH (1), D. FRAISSE (3), I. ESPADALER (1), A. FIGUERAS (1), M. de TORRES (1). 1- Lab. Espectrometria de Masses. CID - CSIC. Jordi Girona, 18-26. 08034-Barcelona (Spain).- 2- AigUes de Barcelona. P.S. Joan, 39, 08009-Barcelona (Spain). 3- Service Central d'Analyse. C.N.R.S. BP-22, 69390 Vernaison (France).
Summary
Polyethoxylated C12-C15 alcohols, nonylphenols and ((nonylphenoxy) polyethoxy)acetic acid and their brominated derivatives were identified in the River Llobregat and in the tap water of Barcelona by GC/MS and FAB mass spectrometry. Whereas GC/MS is limited to those compounds with a low polyethoxylation degree, FAB mass spectrometry is a fast and simple technique to determine sufactants with a high range of oligomers present in the samples. We believe that both techniques are complementary. The source of these compounds are the dumps of surfactant i ndustri es and dyei ng and textil e processes. Thei r bromi nated deri vat i ves are formed in the ch 1 ori nat i on stages in the water works plant due to a hi gh content of bromi de ions in the upper course of the river.
INTRODUCTION The surfactant end use market in 1982 (1) was household 30%, (laundry
dishwashing), personal care 16% (toilet soap, shampoo) and industrial 54% (industrial and institutional process aids, etc) covering 30 million metric tons worldwide. Among the major synthetic surfactants used in the household product market are the linear alkylbenzene sulfonates, alcohol ether sulfates, whereas alkylphenol ethoxylates are used almost exclusively in industrial applications such as tanning and textile processing.
One of the most important factors that has i nfl uenced the market of surfactants are environmental concerns associated with some components of different detergent formulations, e.g. nonbiodegradable branched alkylbenzene sulfonates causing foaming which have been replaced by linear alkylbenzene sulfonates. It has recently been confirmed that nonionic surfactants of alkylphenol type can be biotransformed into persistent and toxic metabolites (2).
The Llobregat river is an extremely polluted river supplying water to Barcelona (Spain) and its surroundings with a population of 3.2 million. The quality of the water depends on its torrent i a 1 regi me and on the i ndustri es located on its banks. Of these i ndustri es, two are related to surfactant manufacturing and many others to tanning, dyeing and textile processes. Salt mi ne di scharges from the upper course of the ri ver are responsible for high levels of bromide ion in the water, entering the water works plant leading to the formation of high levels of brominated compounds (3) in tap water.
Since Sheldon and Hites (4) first reported the presence of alkylphenol ethoxylates in river water, these compounds and their metabolites have been
- 329-
identified in treated municipal waste waters (5), river raw water (6), di gested sewage sludges (7) and tap water (8) . Rei nhard et a 1 (9) established that the acidic and neutral metabolites of the alkylphenol ethoxylates react during chlorination producing brominated and chlorinated products.
The widespread use of surfactants in the environment has required the employment of both sensitive and specific methods for their determination at trace 1 eve 1 s si nce col ori metry whi ch is the common procedure for the analysis of surfactants is neither specific nor sensitive. For nonionic surfactants, the employment of GC!MS using electron impact (EI) (4,10) or chemical ionization (CI) (11,12) has proved to be useful for the analysis of these compounds with a low po lyethoxyl at i on degree. Ri vera et a 1 (8) identified nonylphenols by GC/MS in river water up to n = 7. HPLC has been used (13,14) to determi ne noni oni c surfactants wi th a hi gher range of 01 i gomers in envi ronmenta 1 water samples. However, the i ncreas i ng use of new ionization techniques such as desorption chemical ionization (DC I) , field desorption (FD) or fast atom bombardment (FAB) and the combination of FD or FAB with co11isionally induced decomposition and mass analyzed ion kinetic energy spectroscopy (CID-MIKE) has been successfully employed for the characterisation of pure industrial surfactants (15,21). Shiraishi (22), Otsuki (23), Crathorne (24) determi ned alkyl pheno 1 sin raw water by FD and Levsen by FD-CID (6). Rivera et al (25) demonstrated the presence of a broad range of surfactants by FAB and FAB-CID-MIKE in raw and drinking water after HPLC diode array fractionation of organic extracts.
The present paper reports the i dentifi cati on of alkyl pheno 1 s, brominated alkyl phenols and acidic derivatives in raw and drinking water by the use of GC/MS and FAB mass spectrometry.
EXPERIMENTAL Organlc compounds of raw and drinking water (ea 2000 1) were adsorbed
with granular activated carbon (GAC), the same type of carbon filters used in the water works plant. The GAC, was removed, drained, transferred into a soxhlet and extracted with dichloromethane (48 h). The organic extract was concentrated and fractionated into acids and bases + neutrals. The acidic fracti on was analyzed as methyl esters after convent i ona 1 deri vat i zati on with BF3/MeOH.
Gas chromatography: A Konik 3000 gas chromatograph equipped with a 25 m x 0.22 mm fused silica column, 0.25 um film BP-5 (SGE, Australia) was used. The temperature was initially 60°C for 3 min, then programmed at 4°C/min to 260°C for 10 mi n. Hydrogen was the carri er gas (12 ps i) and nitrogen used as make-up (30 ml/min). Splitless injections were carried out for 40 sec.
GC!MS: A Koni k 2000 gas chromatograph coupled with a MS-9 VG updated mass spectrometer (VG Ana lyt i ca 1 UK) and VG 11 /250 data system was used. The ana lyses were carri ed out with a BP-5 fused sil i ca column coupled directly to the ion source. Helium was the carrier gas with a back pressure of 12 psi. Same gas chromatographic conditions as described above. For the EI mode, the conditions were as follows: ionization energy 70 eV, mass range 40-500, scan time 2 sec/dec and 1000 of resolving power.
FAB mass spectrometry: 1-2 ul of the extracts were dissolved in a small quantity of thioglycerol saturated with NaCl on the stainless steel of the FAB probe. The samples were inserted into the FAB saddle field source (Ion Tech Ltd) of the MS9 mass spectrometer and bombarded by a neutral beam -8 kV- of xenon.
- 330-
Table I Selected ions in the EI mode of nonylphenols and acidic derivatives used in this study.
Nonylphenols
X=Y=H n=O m/z=121,135 n=l 179,193 n=2 223,237 n=3 267,281
X=Br and Y=H
if R=CH2 see left column
R= CH2CH3 see right column
n=O m/z=213,227 X=Cl n=O m/z=169,183 n=l 257,271 and n=l 213,227 n=2 301,315 Y=H n=2 257,271 n=3 345,359 n=3 301,315
Nonylphenol polyethoxylated
Base peak m/z = 117
X
R "c fr(CH CH 0) CH COOCH if R=CH2 (left column) /' ~ 22m 2 3 R=CH CH (right column)
H3C Y 2 3
X=Y=H m=O m/z=207,221 X=Br m=O m/z=285,299 X=Cl m=O m/z=241,255 m=l 251,265 and m=l 329,343 and m=l 285,299 m=2 295,309 Y=H m=2 373,387 Y=H m=2 329,343 m=3 339,353 m=3 417,431 m=3 373,387
* Dichlorinated, dibrominated and chlorobrominated compounds also were investigated but were not present apparently in the samples.
RESULTS Table I shows the selected ions in the El mode of nonylpheno1s and
their acetic derivatives used in this study in order to enhance both sensitive and reliable identification.
The base peaks for polyethoxylated nony1phenols (NPH) correspond to C4,OI-dimethyl and OI-methyl-ol-ethyl structure in the nonyl chain and addi
tion of oxyethylene groups. The chosen ions are the peaks due to the cleavage of benzylic bond. For nonylphenoxy polyethoxy acetic acid methyle~ters NPHac), the base peak is m/z= 117 corresponding to CH2-CH2-0-CH2-COOH3 . with other relevant peaks showing as before ",DC-dimethyl and «-methy1-~~ethyl cleavage. Mass spectra of NPH's in both electron impact (El) and chemical ionization (CI) modes have been explained by Giger and Stephanou (10,11). The mass spectra of NPHac compounds and their halogenated derivatives have been reported by Stephanou (26).
Figure 1 shows an example, the partial reconstructed ion chromatogram
- 331-
of the aci di c fraction of raw water enteri ng the water works plant. The presence of NPHac with m up 3 becomes apparent, the (nonylphenoxy)ethoxy acetic acid being the most abundant. The same degree of polyethoxylation for NPH in the base + neutral fraction was observed. Samples analyzed along the river course gave the same results as shown in Fig 1. No conclusions about the bi odegradati on coul d be inferred si nce major surfactant manufacturing plants and dumps by dyeing textile processes are located in this area.
Tap water, due the chlorination processes and high levels of bromide ion in raw water, produce large quantities of brominated compounds, such as trihalomethanes (3) and as shown in Fig 2 brominated derivatives of surfactants.
Fig 2 shows the combined selected ions for NPH and brominated nonylphenols (BrNPH) in the base + neutral fractions. With our GC/MS conditions we were able to identify up to n=3 for NPH and n=0-2 for BrNPH.
In the acidic fraction (Fig 3) both NPHac and BrNPHac were present with lower polyethoxylation degree, m=0-2 and m=O-l respectively than raw water samples entering the water works plant. Although chlorinated compounds gave the same fragments as bromi nated ones with one oxyethyl ene group 1 ess, careful examination of mass spectra showed the typical isotopic pattern of bromi nated compounds a 11 owi ng us to di scard the presence of ch 1 ori nated derivatives of surfactants of the alkylphenol type.
The capability of FAB mass spectrometry for the analysis of organic polar compounds was demonstrated. The same base + neutral and
~l ITVz=207 522 J I .. , 'i ,
~I ~=221 557
I.~ , I I I
~I i'T'II.z:=251 6~ I';.J.~
I 0
~1 m~=265
I I ~7 I I I I , I I I •
~l m~=295 Ie I I I I , I I
100- ~=309 732
01 ~~ I I
1~1 m/z =339 790
637 IdIIA~ I • 4
I
10:) ITVz =353
:~; I I I " • I m SCANS 900
Fig 1. Raw water. Acidic fraction analyzed as methyl esters. Selected ion monitoring of NPHac ranging from m = 0-3. See Table I for peak assignation.
- 332-
MlZ r----..,-,~+_' .... ~ ......... --,~ 329 r-~~~~~~~~~
A~~~~~~268288 6!je~~~~~228 248
~~-w~~'~w-____ -~~299 ~--~N~--~------~~I~
168 r~~~~J--~~ __ ~I~
C~~~~::;::::;::;:~I28 ~- 456 509 558 689 658 789 758 sea SCANS
Fig 2: Tap water. Base + Neutral fraction. Presence of NPH (n=0-3) and BrNPH (n=0-2). See Table I for peak assignation.
fig 3: Tap water. Acidic compounds analyzed as methyl esters. NPHac (m=0-2) and Br NPHac (m=O-l) are identified. For peak assignation see Ta-b 1 e 1.
acidic fractions of raw and tap water analyzed by GClMS without further separation were analyzed by FAB using thioglycerol saturated with NaCl as a matrix, which is the most suitable for polyethoxylated compounds (21).
Fi gures 4 and 5 show the pos it i ve FAB mass spectrum of raw and tap water acidic fractions respectively. Raw water shows 44 dalton spaced ion series corresponding to NPHac with m=0-7, m=l being the most abundant. The series (m/z=3l5,359,403,447,49l ... ) is thought to arise from sodium attachment to the molecular ion M+Na~~Other identified peaks are related to the same fragments observed in the EI spectrum as m/z=207,22l and m/z=25l,265 (see Table I) corresponding to the cleavage of the alkyl chain
- 333-
98 88 ~8
68 58 48 39 28 18 8
Fig 4:
98 88 ~8
68 58 48 38 28
o • 0
Positive FAB spe~trum of acidic fraction of raw water NP~ac (*) with m=O-7 (M+Na =359, for m=l; NPH (x) with m=O-6 (M+Na =463 for + n=5); polyethoxylated alcohols ¥12-C15 with n=O-lO,C15-0H (O)(MtNa .427 for n=4); C14-0H (-) (M+N~ =457 for n=5);C13-0H (8) (M+Na = 3.9-9 for n=4); C12-0H (L!.) (M+Na =385. for n=4)
o o
o
.Fig 5: FAB(+). Tap water acidic fraction. BrNPHac 010 with n=O-6~(M+Na+= 437, m=l). Other symbols are explained in Fig 4.
and m=l respectively. For m/z=279 an ~,~-diethyl nonyl chain is suggested. Other peaks (m/z=243,287,331,375,419,463,507) are indicatives of traces of NPH from n=O-6. Polyethoxylated C12 to C15 alcohols with n=O-lO were also observed, showing the characteristic 44 dalton ion series. For example, m/z=427 and 399 correspond to C15 and C13 alcohols respectively for (M+Na~
- 334-
98 88 78 68 58 48 38 28 18 8
x
.Fig 6:
1
98 88 78 ~
58 48 38 28 18
FAB (+). Base + Neutral fraction. NPH (x) with m=O-lO. Polyethoxylated alcohols C12-C13 with n=O-12. For peak assignations, see Fig 4. Commercial silicone based lubricant (~) was also identified by its m/z=73,147,207,221 (not showed) and 355 ion peaks.
8
8 y
x 8
8
8
Fig 7: FAB (+). Tap water base + neutral fraction. Br NPH (0) with n=O-lO. (M+Na+=541 for n=4). See fig 4 for other symbols.
with n=4. Positive FAB spectrum of acidic fraction of tap water (Fig 5) is rather more complex than raw water. A high background is observed, probably due to chlorine action on organic matter, producing low molecular weight compounds. Nevertheless, the presence of NPHac and NPH of low polyethoxylation is shadowed by background but the presence of BrNPHac from
-lli-
m=0-6 is observed. + The series (m/z=393,437,481,525,569,615,657) is assigned to (M+Na) for
different degrees of po lyethoxyl at; on. Thi s seri es ; s the same as for Cl NPHac with one more oxyethyl ene group, but the i sotopi c pattern of monobrominated compounds shown in the spectrum allowed us to discard this possibility. Again po1yethoxylated alcohols with the same degree of po1yethoxylation as raw water are observed, showing no apparent removal of these compounds by granu1 ar activated carbon fil ters of the water works plant. The mai n problems encountered when i nterpreti ng the spectra arose from those low intensity peaks and the possibility of overlapped compounds. For example, pentadecy1 alcohol with a degree of po1yethoxy1ation n, gives the same serie as Br2NPHac with a degree of n-5,C12NPHac (n-3) and C1BrNPHac (n-4). These last three compounds which are byproducts of chlorination processes must show the isotopic profile of halogenated compounds and cannot be found in raw water, and for this reason we assume that they will in both cases be po1yethoxy1ated alcohols.
Figures 6 and 7 depict the positive FAB spectra of raw and tap water, base + neutral fraction. NPH from n=0-10 are identified, n=2 being the most abundant. Also po 1yethoxy1 ated C12 and C13 a 1 coho 1 s are observed wi th a maximum degree of n=12 both in raw and tap water. Brominated non~lpheno1s (BrNPH) are i dent ifi ed in tap water (Fi g II by thei r (M+Na) peaks (m/z=365,409,453,497,541,585 ... ) up to n=10 showing the characteristic isotopic patterns of monobrominated compounds.
CONCLUSIONS The analysis of samples from the river L10bregat and from tap water
showed that po 1yethoxy1 ated compounds of a 1 coho 1, alkyl pheno 1 and a1ky1pheno1 carboxylic acids type, are common and refractory pollutants.
In this work we have reported the formation of halogenated derivatives of alkyl phenols and acidic alky1pheno1s in the chlorination process. The most usual were monobrominated compounds, which is in agreement with majority formation of bromoform and other brominated triha10methanes during normal operating conditions of the water works treatment plant. It has been demonstrated that thi sis due to hi gh 1 eve 1 s of bromi de, proceedi ng from the salt mine discharges in the upper course of the river.
For the analyses, both GC/MS and FAB mass spectrometry were employed. Although both techniques gave similar results when determine these po 11 utants, GClMS was 1 imi ted to the ana 1ysi s of compounds with a sma 11 degree of ethoxylation, whereas FAB mass spectrometry offered good sensitivity and specificity for a higher range of oligomers.
BIBLIOGRAPHY (l)-HAOPT, D.E. Tenside Deterg 20 (1983),6:332-337. (2)-STEPHANOU, E., GIGER,W., Environ Sci & Techno1 (1982), 16:800-805. (3)-VENTURA,F., RIVERA,J., Bull Environ Contam & Toxico1 (1985) 35:73-81. (4)-SHELDON, L.S., HITES, R. Environ Sci & Techno1 (1978), 10:1188-1194. (5)-REINHARD, M., GOODMAN, N., MORTELMANS, KE. Environ Sci & Techno1
(1982), 76:351-362. (6)-SCHENEIDER, E., LEVSEN, K., BOERBOOM, A.J.H., KISTEMAKER, P., Mc
LUCKOY, S., PRZYBYLSKI, M., Anal Chern 56,11 (1984) 1987-1988. (ll-GIGER, W., BRUNNER, P.H., SCHAFFNER, C., Science (Washington DC) 1984,
225,623-625.
- 336-
(8)-RIVERA, J., VENTURA, F., CAIXACH, J., DE TORRES, M., FIGUERAS, A., GUARDIOLA, J., Intern. J Environ Anal Chern (1987) 29, 15-35.
(9)-BALL, H.A., REINHARD, M., In "Water Chlorination", Jolley, R.L. et al eds. Lewis, Chelsea MI 1985; vol 5, 1505-1514.
(10)-GIGER, W., STEPHANOU, E., SCHAFFNER, E., Chemosphere (1981) 10: 1253-1263.
(ll)-STEPHANOU, E. Chemosphere (1984) 13:43-51. (12)-STEPHANOU, E. Organ. Mass Spectrom (1984) 19:510-513. (13)-AHEL, M., GIGER, W. Anal Chern (1985) 57:1577-1583. (14)-AHEL, M., GIGER, W. Anal Chern (1985) 57:2584-2590. (15)-COTTER, R.A., HANSEN, G., JONES, LR. Anal Chimica Acta (1982),
136:135-142. (16)-WEBER, R., LEVSEN, K., LOUTER, G.R., BOERBOOM, J.H., HOVERKAMP, J.
Anal Chern (1982), 54:1458-1466. (171-SCHENEIDER, E., LEVSEN, K., DAHLING, P., ROLLGEN; F.W., Fresenius Z
Anal Chern. (1983), 316:277-285. (18)-SCHENEIDER, E., LEVSEN, K., DAHLING, P., ROLLGEN, F.W., Fresenius Z
Anal Chern. (1983), 316:488-492. (19)-LYON, P.A., STEBBINGS, N.L., CROW, F.W., TOMER, K.B., LIPPSTREAU,
D.L., GROSS; M.L. Anal Chern (1984), 56:8-13. (20)-LYON, P.A., CROW, F.W., TOMER, GROSS, M.L. Anal Chern, (1984), 56:2278-
84. (21l-RIVERA, J., VENTURA, F., CAIXACH, J., FRAISSE, D., DESSALCES, G. In
"Advances in Mass Spectrometry". J.F.J. Todd ed. John Wiley & Sons. (UK). (1986) 1453-1454.
(22)-SHIRAISHI, H., OTSUKI, A., FUWA, K., Bull Chern Soc Jpn (1982), 55: 1410-1415.
(23)-OTSUKI, A., SHIRAISHI, H., Anal Chern (1979), 51 :2329-2332. (24)-WATTS, C.D., CRATHORNE, B., FIELDING, M., STEEL, C.P. In "Analysis of
Organic Micropollutants in Water. Bjorseth, A., & Angeletti, G., eds. D. Reidel Publish. (Holland), (1984),120-131.
(25)-RIVERA, J., VENTURA, F., CAIXACH, J., FIGUERAS, A., FRAISSE, D., BLONDOT, V., In Organi c Mi cropo 11 utants in the Aquatic Envi ronment" • Bjorseth, A. & Angeletti, G., eds. D. Reidel Publish (Holland), (1986), 77-89.
(26)-STEPHANOU, E. In "Air & Water Analysis: New Techniques & Data". Frei, R.W. & A1baiges, J. eds. Gordon Breach Sci Publish. (UK), (1986), 237-250.
- 337-
Summary
EFFECTS OF CHLORINE DIOXIDE PREOXIDATION ON ORGANIC HALIDE FORMATION POTENTIALS
H. BEN AMOR, J. DE LAAT and M. DORE Laboratoire de Chimie de l'Eau et des Nuisances
40 avenue du Recteur Pineau 86022 Poi tiers Cedex - France
The aim of our study was to determine the effects of a preoxidation with chlorine dioxide on the production of organohalogenated compounds (trihalomethanes, dichloroacetic and trichloroacetic acids) formed from the chlorination of different raw surface waters and of isolated aquatic humic substances. Samples of natural waters and of fulvic and humic acid solutions were oxidi zed for a reaction time of 24 hours with increasing doses of chlorine dioxide (from 0 to 2.5 mg of Cl02 per mfh of TOC) , then heavily chlorinated (chlorine doses: 40-60 mg 1 , reaction time : 72 hours). The results obtained with laboratory experiments showed that surface waters consumed about 0.7 mg CIO /mg TOC after a reaction time of 24 hours and ful vic and humic acits about 1.6 mg C102/mg TOC. For doses of chlorine dioxide corresponding to the above values of demands, the experiments showed that there was a significant decrease in the formation potentials of trihalomethanes (45-55 %), of dichloroacetic acid (30-40 %) and of trichloroacetic acid (40-60 %) as well as in the chlorine demand during postchlorination (10-20 %). Compared to chlorine, chlorine di oxide produced very small amounts of organohalogenated compounds. Furthermore, the production of chlorite and the concentration of organic matter (Total Organic Carbon and UV absorbance measurements) were also measured during our experiments.
1. INTRODUCTION II est bien connu actuellement que l'action du chI ore sur la micropol
lution organique dissoute des eaux a potabiliser peut conduire a des productions importantes de composes organohalogenes. Ainsi, des productions de l'ordre de 150 a 300 pg de chI ore organiquement lie (TOX) par mg de Carbone Organique (COT) sont generalement mesurees lors de la chloration en presence d' un large exces de chlore d' eaux de surface ou de solutions de substances humiques extraites d'eaux naturelles (1, 2). Parmi ces composes, les trihalomethanes (THM) et les acides dichloroacetique (DCA) et trichloroacetique (TCA) constituent globalement la majeure partie du TOX (40 a 60 %) et des pourcentages de l'ordre de 15 a 30 %, de 3 a 7 % et de 5 a 30 % du TOX peuvent respectivement etre attribues aux THM, DCA et TCA. Compte tenu du caractere relativement toxique attribue a certains des composes organohalogenes engendres par la chloration, Ie bioxyde de chlore en tant que reactif oxydant et desinfectant peut constituer une des alternatives interessantes au chI ore dans Ie but de reduire la concentration en composes indesirables (trihalomethanes en particulier) dans les eaux de distribution. L' action du bioxyde de chlore sur les composes aromatiques phenoliques (3) et azotes (4), sur les substances humiques (5) et sur la
- 338-
micropollution organique des eaux de surface (6) conduit a la formation de composes quinoniques, d'aldehydes, d'acides aliphatiques et aromatiques et, comparativement au chI ore , a des productions beaucoup moins importantes (5 a 20 fois moins) en composes organohalogenes (6, 7).
Dans Ie cadre de cette etude, nous no us sommes donc proposes d'etudier l' incidence d" une preoxydation chimique au bioxyde de chlore sur Ie potentiel de formation des THM, DCA et TCA a partir de solutions aqueuses diluees de substances humiques extraites d'eaux naturelles ainsi qu'a partir de diverses eaux de surface brutes. Parallelement, les demandes en chlore et en bioxyde de chlore, l'evolution de la matiere organique (parametre COT et absorbance UV) et la production de chlorites ont ete egalement suivies.
2. PROTOCOLE EXPERIMENTAL -1 Les s~tutions meres de chI ore (5 a. 10 g 1 ) et de bioxyde de chlore
(4 a. 6 g 1 ) ont ete preparees au laboratoire, respectivement par action de l' acide chlorhydrique sur Ie permanganate de potassium et de l' acide sulfurique sur du chlorite de sodium, et dosees par iodometrie. Dans ces condi tions de preparation, les solutions meres de bioxyde de chlor:'1 ne contiennent que de faibles teneurs en chlori tes et en chlore ( <50 mg 1 ) •
Les solutions aqueuses de produi ts organiques ont ete preparees dans I' eau ul tra-pure ~~urnie _par un ensemble _~ILLIPO~~) et tamponnees a. pH 7,5 (KH P04 : 10 moll; Na2HP04 : 4 10 moll). Les acides humiques et fufviques ont ete extraits d'une eau de mare situee dans la reserve naturelle du "Pinail" (AHP et AFP) et d'une eau prelevee dans Ie barrage de Moulin Papon (AFMP) en utilisant la procedure d' extraction de THURMAN et MALCOLM (8). L' analyse elementai re a donne des pourcentages massi ques en carbone de 52,5 % pour AHP, de 54,6 % pour AFP et de 48,9 % pour AFMP.
Les echantillons de di verses eaux de surface ont ete prefil tres sur membrane 1,2 pm avant oxydation par Ie chlore ou par Ie bioxyde de chlore.
Toutes les reactions d' oxydation avec Ie chlore et Ie bioxyde de chlore ont ete effectuees a 20°C et a. l' obscuri te. Apres un certain temps de reaction (24 h ou 72 h), Ie bioxyde de ch10r:'1 residuel est dose par spectrophotometrte UV a 360 nm (6 = 1200 1 mol cm ) et par des methodes colorimetriques au rouge de chlorophenol (9) et a. l'acide violet K (10). Le chI ore residuel a ete dose par iodometrie et les chlorites par chromatographie liquide haute performance (HPL~i avec une detection UV a 260 nm (limi te de detection : 0,2 a. 0,3 mg 1 ). Les composes organohalogenes ont ete doses avec un chromatographe en phase gazeuse equipe d'un detecteur a. capture d'electrons (PACKARD modele 438 S ou 439) apres extraction au pentane pour les trihalomethanes et apres extraction a. l' ether en milieu acide et esterification au diazomethane pour les acides dichloacetique et trichloracetique (11). L'evolution de la matiere organique a ete suivie par des mesures de concentration en COT (analyseur de carbone DOHRMANN DC 80) et d'absorbance UV a 254 et 270 nm (spectrophotometre VARIAN DMS 90).
3. RESULTATS EXPERIMENTAUX
3.1. Chloration Les reactions de chI oration ont ete realisees au laboratoire avec des
taux et des temps de chloration tres eleves. Apres 72 heures de reaction, les resultats obtenus (Tableau I) montrent que la chI oration de substances humiques et d' eaux naturelles conduit a des productions comparables en CHC13 ou en THM (entre 30 et 70 pg/mg COT). En ce qui concerne les potentiels de formation de DCA et de TCA, les valeurs obtenues montrent une production plus importante ae TCA (40 a l20pg/mg COT) que de DCA (18 a 30 pg/mg COT) ou de chloroforme. Comme l'a egalement montre RECKHOW (1), les acides humiques presentent une reacti vite plus grande que les acides fulviques vis-a.-vis de la demande en chlore et de la production potentielle
- 339-
en derives organohalogenes. Par ailleurs, les productions de THM, de DCA et de TCA mesurees a partir de diverses substances humiques et d'eaux naturelles semblent etre difficilement correlables avec la demande en chI ore ou la concentration en COT alors que l'absorbance UV a 254 nm (ou a 270 nm) semble constituer un parametre plus representatif de la formation potentielle en composes organohalogenes.
Echantillon Origine COT pH mg 1-1
00 270 nm -1
em
C12 applique C1 2 consomme PF CHC13 PF THM PF TCA PF DCA
mg 1-1 mg 1-1 )l& 1-1 Vg 1-1 )l& 1-1 }Ii 1-1
AF Pinail Mare 2,5 7,5 0,098 60 6,2 150 150 220 65
AH Pinail Mare 2,5 7,4 0,132 40 11,5 170 170 300 75
AF Moul in Papan Barrage 2,5 7,5 0,089 40 7,9 130 130 130 65
Mervent Barrage 7,7 0,150 50 13,1 210 254 200 100
Autize Riviere 5,1 7,9 0,075 40 7,6 125 160 120 90
Saumor Ri viere 2,4 7,6 0,072 40 6,8 85 129 115 75
Clain Riviere 2,7 8,3 0,047 20 8,5 85 120
Moul in Papan Barrage 8,3 7,8 0,160 60 16,5 270 365
Garaon Barrage 6,3 7.3 0.169 30 12,4 205 281
Vienne Rivi~re 6,0 7,6 0,188 40 15,1 260 273 240 130
TABLEAU I Chloration de substances humiques et d'eaux de surface: demande
en chlore et potentiels de formation de THM, de DCA et de TCA (temps de contact: 72 h a 20·C et a l'obscurite)
3.2. Oxydation au bioxyde de chlore Le tableau II rapporte les resultats obtenus pour des taux d'oxydation
appliques proches de 2 mg de Cl02/mg COT et pour des temps de reaction de 24 heures. Dans ces condi tions, les consommations en Cl02 ont ete de l' ordre de 1,7 mg de C102/mg de COT pour les substances humiques et de l'ordre de 0,5 a 1,0 mg de Cl02/mg de COT pour les eaux de surface.
C102 applique CI02 consomme CI02- fOnlle % Abattement TCA DCA Echantillon
mg 1-1 mg/mg COT mg 1-1 mg/mg COT CI02 consomme
270 nm 254 run }lg 1-1 }lg 1-1
AFP 5,0 2,0 4,0 1,6 0,72 29,6 23,3
AHP 4,6 1,9 4,5 1,8 0,60 23,S 19,9
AFMP 5,1 2,0 3,9 1,6 0,67 30,3 25,5 10 10
Mervent 8,1 1,6 5,2 1,0 0,65 18,7 14,3
Autize 10,2 2,0 4,1 0,8 26,7 19,5 25 20
Saumor 4,9 2,1 2,3 1,1 18,1 15,5 15 15
Cialo 5,1 1,9 2,7 1,0 0,65 12,2 13,2
Maul in Papan 14,9 1,8 5,8 0,7 0,70 29,1 21,3
Garaon 11,6 1,8 6,0 0.9 0,62 31.4 27,4
Vienne 12 2,0 9,1 1,5 0,63 26,1 20.0 30 15
TABLEAU II Oxydation de substances humiques et d'eaux de surface par Ie bioxyde de chI ore : Demande en CI02 , productions de Cl02-, de DCA et de TCA,
abattement des absorbances UV a 254 et 270 nm (temps de contact : 24h a 20·C et a l'obscurite)
Les analyses effectuees montrent egalement que 60 a 70 % du bioxyde de chlore consomme se retrouve so us forme de chlori tes et des valeurs similaires ont ete mesurees pour des taux de bioxyde de chlore appliques inferieurs. En ce qui concerne la production de composes organohalogenes, les
- 340-
analyses revHent des productions negligeables de chloro!frme ( <5 ug C l ) et que les productions en DCA et en TCA (5 a 20 pg 1 ) sont egalement tres largement inferieures a celles obtenues par chloration. La formation de composes organohalogenes peut s' expliquer par une liberation d' acide hypochloreux dans Ie milieu reactionnel au cours des reactions d'oxydation de la matiere organique par Ie bioxyde de chI ore (4, 12).
Enfin, l'oxydation par Ie bioxyde de chlore ne conduit qu'a des abattements peu significatifs du parametre COT (environ 5 %) et a des abattements compris entre 13 et 30 % sur Ie parametre absorbance UV a 254 nm ou 270 nm qui permettent d' envisager une reduction de la production potentielle en derives organohalogenes au cours d'une post-chloration.
3.3. Incidence d'une preoxydation chimique au bioxyde de chlore L'etude de I'incidence d'une preoxydation chimique au bioxyde de
chI ore sur Ie potentiel de formation de composes organohalogenes a ete conduite en realisant une preoxydation chimique avec des taux de bioxyde de chI ore appliques allant de 0 a 2,5 mg de ClO /mg COT et addition d'une quanti te fixe de chlore apres 24 heures de reac~ion, les concentrations en THM, DCA et TCA etant determinees apres 72 heures de chloration. Compte tenu de la formation de 0,6 a 0,7 mg de chlorite par mg de Cl02 consomme au cours de l' etape de preoxydation et de la presence eventuelle de ClO resi~~el dans certains echantillons, un taux de chloration de 20 a 66 mg 1 a ete applique afin de satisfaire les demandes en chI ore par la matiere organique.. par les ions chlori tes et par Ie bioxyde de chlore et d'obtenir un large exces de chlore libre residuel pour la determination du potentiel de formation de composes organohalogenes.
Les figures Ia et Ib representent des exemples representatifs de courbes d' evolution du potentiel de formation en composes organohalogenes en fonction du taux de preoxydation obtenu avec des molecules representatives de la micropollution organique des eaux a potabiliser (substances humiques) et avec les eaux de surface. Ces courbes mettent en evidence que l'action du bioxyde de chlore conduit a une diminution significative du potentiel de formation de THM, de DCA et de TCA. Ainsi, pour les taux de bioxyde de chlore mentionnes dans Ie tableau III et correspondant a l'apparition d'un leger residuel~fn bioxyde de chI ore apres 24 heures de reaction ([C102J de 0,1 a 0,5 mg 1 ), des abattements de l'ordre de 40 a 60 %, de 30 a 50 % et de 20 a 30 % ont ete respectivemert mesures pour la production potentielle en THM, TCA et DCA. Il convient dt. noter que l' abattement moindre en DCA qu'en TCA a egalement ete observe apres une preozonation (11).
Apres une post-chI oration , nous avons egalement remarque une elimination complete des chlorites dans les solutions apres 72 heures ~ reactio~a Des c~orations de s~~utions _rqueuses de chlori tes ([C102 ] = 5 10 moll; [Cl] = 10 moll; pH : 7-7,5) ont permis ae mettre en evidence par tes analyses en chromatographie liquide une oxydation totale des chlori tes en chI orates apres 24 heures de reaction. L' oxydation des chlori tes en chlorates s' accompagne d' une consommation de l' ordre de 1 a 1,3 mole de chlore par mole de chlorites et de la liberation de traces de bioxyde de chlore en tant que produit intermediaire de reaction (12).
Cette consommation de chlore par les chlorites formes au cours de la preoxydation chimique au bioxyde de chlore (0,6 a 0,7 mg de Cl02-/mg de Cl02 consomme) et par Ie bioxyde de chlore residuel permet d'expliquer que l'incidence d'une preoxydation chimique au bioxyde de chlore sur la demande en chlore au cours d'une post-chloration dependra en particulier du taux de pretraitement. Pour des faibles taux de traitement en Cl02 , une preoxydation induira une diminution faible mais significative de la demande en chI ore (10 a 20 %). Par contre, pour des taux de traitements plus eleves, la diminution de la demande en chlore par la matiere organique oxydee peut
- 341-
etre annulee par une consommation relative plus importante en chlore par CI02 - et CI02 , Bien que I' elimination des chlorHes par Ie chI ore soH interessante du point de vue sani taire, la formation en quanti te stoechiometrique de chI orates ne res out pas Ie probleme pose par la presence de chlorites et de chI orates dans les eaux de distribution traitees au bioxyde de chlore,
CHC1 J . TCA. DCA
:::;~)·r' CHCI J • TCA . DCA Demand. en CI Z (\.Ig I-I)
50
• ().Ig I-I) (mg I - I)
""-TCA
~.--o~ 15
10 200
10
"" TeA "-.~
A~ 5 l~
0.5 1
1. a . AfM?
1.5 Z CIOZ (mg/mg COT)
A A-A
0 . 4 0.8 1 . 2 1.6 CI02(mg/mg COT)
I.b . Eau de Mervent
Figure:; la et Ib
S
Incidence de la preoxydation chimique au CI02 sur la production de THM, de TCA et de DCA sur la demanae en chI ore
(preoxydation : 24 h, postchloration : 72 h)
C102 applique % Abattement Echan ti lIon -1 mg I mg/mg COT CHC1 3 THM TCA DCA C1Z consomme
AFP 3,8 1,5 52 50 39 0
AHP 3,5 1, 4 35 47 40 6,7
AFM? 3,8 1,5 57 35 20 4,3
Mervent 4 0,8 29 26 38 15 19
Autize 2,6 0,5 52 44 25 17 16,3
Saumor 1,9 0,8 29 34 26 20 0
Clain 2,1 0,8 50 38
Moulin Papon 4,1 D,S 30 46 0
Garaon 4,2 0,7 56 54
Vienne 7,2 1,2 56 55 12
TABLEAU III Incidence de la preoxydation par le bioxyde de chlore sur les potentiels
de formation de THM, de DCA et de TCA et sur la demande en chlore (preoxydation : 24 h, postchloration : 72 h a 20°C et a l'obscurite)
4, CONCLUSION Cette etude, realisee en laboratoire, montre que l'action du bioxyde
de chlore sur la micropollution organique des eaux de surface conduit a des productions potentielles en derives organohalogenes (THM, DCA, TCA) tres nettement inferieures a celles formees par chloration et permet egalement de diminuer d'une maniere tres significative la production de ces derives
- 342-
au cours d' une pos t-chloration. Le bioxyde de chlore consti tue donc une al ternative tres interessante au chlore a I' egard de la formation de composes organohalogenes indesirables et mutagenes. Cependant, Ie developpement de son utilisation dans Ie domaine de la production d'eau potable (en tant que reactif oxydant et desinfectant) sera fortement fonction de la quali te des eaux a trai ter et des reglementations en cours d' elaboration concernant les concentrations maximales admissibles en bioxyde de chlore residuel et en sous-produits d'oxydation (chlorites, chlorates) dans les eaux de distribution.
REFERENCES (1) RECKHOW, D.A. (1984). Organic halide formation and the use of pre-ozo
nation' and alum coagulation to control organic halide precursors. Ph. D. Thesis, Department of Environmental Sciences and Engineering, Chapel Hill (N.C.).
(2) NORWOOD; D.L., THOMPSON, G.P., ST AUBIN, J.J., MILLINGTON, D.S., CHRISTMAN, R.F., JOHNSON, J.D. (1985). By products of chlorination: Specific compounds and their relationship to total organic halogen. Michigan, Lewis Publ., INC/Drinking Water Res. Found., 109-121.
(3) BEN AMOR, H., DE LAAT, J. et DORE, M. (1984). Mode of action of chlorine dioxide on organic compounds in an aqueous medium-chlorine dioxide consumption and reactions on phenolic compounds. Water Res., 18, nO 12, 1545-1560.
(4) BEN AMOR, H., DE LAAT, J. et DORE, M. (1985). Mode of action of chlorine dioxide with certain nitrogenous compounds in an aqueous medium. Environ. Technol. Letters, 6, 489- 504.
(5) COLCLOUGH, C.A., JOHNSON, J.D., CHRISTMAN, R.F., MILLINGTON, D.S. (1983). Organic reaction products of chlorine dioxide and natural aquatic fulvic acids. Water chlorination : Environmental Impact and Health Effects, (Ed. R.L. Jolley), 4, Ann. Arbor Science, 219-229.
(6) STEVENS, A.A. (1982). Reaction products of chlorine dioxide. Environmental Health Perspectives, 46, 101-110.
(7) SAVOIR, R., ROMNEE, L. et MASSCHELEIN, W.J., (1986). Evaluation du bioxyde de chlore comme moyen limitant la formation de derives organohalogenes. Seminaire "Substances Humiques" (GRUTTEE et CGE) , Rennes, France, 15-16 octobre.
(8) THURMAN, E.M. and MALCOLM, R.L. (1981). Preparative isolation of aquatic humic substances. Environ. Sci. Technol., 15, 463-466.
(9) HARP, D.L., KLEIN Jr, R.L. and SCHOONOVER, D.J. (1981). Spectrophotometric determination of chlorine dioxide. Journal American Water Works Ass., 73, nO 7, 387-388.
(10) MASSCHELEIN, W.J. (1966). Spectrophometric determination of chlorine dioxide with acid chrome violet K. Analytical chemistry, 38, nO 13, 1839-1841.
(11) LEGUBE, B., CROUE, J.P., RECKHOW, D.A. and DORE, M. (1985). Ozonation of organic precursors effects of bicarbonate and bromide. Proceedings of the international conference "the role of ozone in water and wastewater treatment", (Ed. Perry R. and Mc Intyre A.E., Selper Ltd, London), November 13-14, 73-86.
(12) RA V-ACHA, Ch., CHOSHEN (GOLDSTEIN), E., SERRI, E. and LIMONI, B. (1985). The role of formation and reduction of THM and chlorite concentration in the disinfection of water with C12 and Cl02 . Environmental Pollution (Series B), 10, 47-60.
(13) TAUBE, H. and DODGEN, H. (1949). Applications or radioactive chlorine to the study of the mechanisms of reaction involving changes in the oxidation state of chlorine. J. Amer. Chern. Soc., 71 (10), 3330-3336.
- 343-
NMR STUDY OF KRAFT PULP MILL WASTE AND NATURAL HUMIC SUBSTANCES --- -----
L. VIRKKI, J. KNUUTINEN, P. MANNILA and J. PAASIVIRTA Department of Chemistry, University of Jyvaskyla Kyllikinkatu 1-3, SF-40100, Jyvaskyla, Finland
Summary
High field proton NMR spectroscopy in DMSO-d6 solution is applied for structural analysis of main high molecular fractions of natural humus and waste lignin from kraft pulp mill. The spectra are similar in great part. Main differences appear in p-disubstituted benzene proton and -CH 2 -CO-proton signals which are absent in spectrum of waste lignin obviously due to chlorination. A new observation of 51 Hz 1:1:1 triplet in both spectra is discussed.
1. INTRODUCTION Chlorobleaching of pulp is a very important source of
water pollution. Analysis methods for small molecular organochlorine components of pulp bleaching wastes are well developed due to applicability of GC/ECD and GC/MS and availability of model compounds. In contrary, high molecular wastes which include most of the organically bound chlorine, have not been analyzed structure-specifically, thus far. Main components of environmental concern in these wastes can be called "chlorolignin" while they obviously contain structures very near to natural lignin/humic substances. In efforts to find out specific analysis methods for chlorolignin, we have used natural humic fractions as reference compounds. Both mill waste and natural water are fractionated with LC and the fractions studied by different chemical treatments and instrumental analysis methods. In the present paper, new experiences in applying high field proton NMR spectroscopy for detecting differences between chlorolignin and natural humic lignin are decribed.
2. MATERIALS AND METHODS Dissolved high molecular organic matter (HMDOM) fractions
of kraft pulp mill waste liquor (MB) and natural humic water
- 344-
(AIT) were isolated applying the method of Thurman and Malcolm (1). The main fractions which were eluted first (60 % of HMDOM) through a XAD-8 column with 0.1 M NaOH were purified with cation exchange resin IR120, freeze-dried and dissolved in DMSO-d6. Fractions were brown in colour. Their proton spectra were recorded with a 270 MHz instrument JEOL GSX-270 and are presented in Figs 1 and 2.
AIT MB 1 DMSO-dS
15 pp m
Fig. 1. Total proton NMR spectra (270 MHz) of the main frac-tions of natural humus (AIT) and kraft pulp mill waste (MB) •
AIT AIT
us
Fig. 2. Baseline corrected partial proton NMR spectra.
- 345-
3. RESULTS AND DISCUSSION
According to the spectra kraft pulp mill waste and natural humic (HMDOM) are very similar. Differences (marked to spectra with stars) indicate that the common p-susbstituted aromatic (AA'BB' spectrum) and CH2 -neighbour of a carbonyl group in humus have been chlorinated away in kraft lignin waste.
A very interesting 1:1:1 triplet centered at 7.2 ppm with peak intervals of 51 Hz is common in spectra of both AIT and MB (marked by arrows) has, according to our knowledge, never been reported and explained earlier. One possible source for this triplet could be protons of the CONH group, which is common in natural substances and inert against chlorination. Chemical shift of phenyl-CONH-R is reported (2) to be 6.03-8.63 ppm. Coupling of nitrogen-14 which has spin 1 explains the triplet appearance. It is, however normally not resolved due to rapid relaxation of nitrogen. Therefore, coupling between nitrogen14 and hydrogen-l is estimated from the measured corresponding coupling of nitrogen-15 and hydrogen-l by multiplying with -0.713 (3). From coupling measured for 15-N ammonium ion the estimated one-bond coupling 14N-1H should be 52.3 Hz. Consequently, in the present case, both chemical shift and coupling appearance support our assumption about CONH appearance as triplets of the spectra. Resolvation of the triplet might rise from the nature of the solvent used and high molecular structure of the substrates which prevent fast quadrupolar relaxation of nitrogen-14 in this special case.
4. ACKNOWLEDGEMENT Financial support from Environmental Science Council of
the Academy of Finland is gratefully acknowledged.
REFERENCES
(1) THURMAN, E.L. and MALCOLM, R.L. (1981). Preparative isolation of aquatic humic substances. Env. Sci. Tech. Vol.15, 463-466.
(2) CHAMBERLAIN, N.F. (1974). The Practice of NMR Spectroscopy. Plenum Press, London, p. 185.
(3) HARRIS, R.K. and MANN, B.E. (1978).NMR and the Periodic Table. Academic Press, London, pp. 98-99.
-~-
IDENTIFICATION OF BIOACCUMULABLE COMPOUNDS IN KRAFT BLEACHING EFFLUENTS
G.E. Carlberg, H. Drangsholt, N. Gj¢s, L.H. Landmark
Center for Industrial Research Box 124 Blindern, 0314 Oslo 3, Norway
Summary
Between 0.25 and 1 percent of the adsorbable organic halogen in spent bleach liquors was shown to be bound to potentially bioaccumulable compounds. Chlorinated thiophenes, phenols and terpene derivates were identified in the lipophilic fraction.
1. INTRODUCTION There is a growing concern about the release of halogenated organic
compounds into the environment as many of them are toxic and may show considerable resistance to biological and chemical degradation. The bleached chemical pulp industry consumes considerable quantities of chlorine and discharge large quantities of chlorinated organic matter into rivers, lakes and oceans (1).
Recently bleach plant discharges were found to affect the diversity, biomass and distribution of invertebrates and plants within the receiving body of water (2). Furthermore, near the effluent outflow the fish biomass was low, the species composition of the fish community was changed, reproduction of perch was reduced and physiological disturbances were noted.
The discharged compounds of greatest concern are those which are toxic, persistent and lipophilic enough to cause a considerable bioconcentration. Fish from waters receiving bleach plant discharges have been found to contain up to 2000 ppm organic chlorine in the fat (3). Chlorinated alkanes, phenols, cymenes and veratroles have been identified in fish (4,5,6). The identified compounds explain, however, only a few percent of the total amount of organic chlorine in the fish fat.
2. EXPERIMENTAL The spent bleach liquor from the chlorination (60 percent chlorine and
40 percent chlorine dioxide) and alkaline extraction stages of a bleach plant were analysed. The mill is a kraft mill using a spruce and pine mixture.
The adsorbable organic halogen of the spent bleach liquors was determined by carbon adsorption, combustion and microcolulometric determination using a Dohrman DX-20 apparatus (7).
The water samples were extracted with cyclohexane at pH2 and the extractable organic chlorine (EOCl) was determined by neutron activation analysis (NAA) (8). Determination of the part of EOCI which is potentially bioaccumulable was performed by reversed phase thin layer cromatography (TLC) according to Renberg et al. (9).
After elution,the TLC coating material containing the compounds with Pow larger than 1000 were extracted with a cyclohexane/isopropanol (1:1) mixture. The cyclohexane was isolated by addition of acidified (pH2) water and the amount of potentially bioaccumulable organochlorine was established by determining the organic chlorine concentration in the extract by NAA.
-~-
The compounds in the cyclohexane extracts were analysed by a combined gas chromatograph/mas spectrometric system, Finnigan 4021 quadrupole mass spectrometer equipped with a Finnigan 9610 gas cromatograph and an Incos 2300 Data System. A 0.25 mm x 30 m OB-5 fused silica capillary column was user. The ionization mode was El a~ 70 eV and the ion source temperature 2700 C. The injector ras.k~ft at 270 C and the oven heated from 50 to 300 C at a rate of 5 m1n
3. RESULTS
Figure 1 shows the relationship between the adsorbable organic halogen (AOX), the extractable organic chlorine (EOCl) and bioaccumulable organic chlorine (Log Pow >3) for the chlorination stage .. The effluent contained 330 mg/L of adsorbable organic halogen. About 2 percent of the AOX was extractable with cyclohexane, and about 0.25 percent was bound to potentially bioaccumulable compounds.
For the alkaline extraction stage nearly 1 percent of the AOX was bound to bioaccumulable compounds and 2.5 percent was extractable with cyclohexane.
100
2
AOX EOCI log Pow>3
Fig. 1. Extractable organic chlorine (EOCl) and organic chlorine bound to potentially bioaccumulable compounds (Log Pow>3) as percent of adsorbable organic halogen (AOX) in the spent bleach liquor from the chlorination (C60 + 040 ) stage.
- 348-
A number of chlorinated compounds were detected in the GC/MS analysis of the two spent bleach liquors (10). The findings are summerized in Table 1. Figure 2 shows the chromatogram of the extract from the chlorination stage effluent. Structures of some of the identified compounds have been included in the figure.
A number of chlorinated thiophenes were found in the extracts. These compounds have been identified previously in spent bleach liquors (11). Through comparison with synthesized standards, which turned out to be isomers to those in the effluents, it was possible to determine the identity of the chlorinated thiophenes. The effluents were found to contain 2 .. 3,4-trichlorothiophene and di- and trichloroderivates of 3-formyl and 3-acetyl chlorothiophenes.
Table 1 Compounds identified in the lipophilic fraction of spent chlorination and alkaline extraction liquors.
Compound
2,3,4 - trichlorothiophene
3 - formyl-dichlorothiophene
3 - formyl - 2,4,5 - trichlorothiophene
3 - acetyl-dichlorothiophene
3 - acetyl - 2,4,5 - trichlorothiophene
2,4,6-trichlorophenol
Dichloroguaiacol
Trichloroguaiacol (2 isomers)
Tetrachloroguaiacol
M 152 probably C10 H1S O
M 184 probably C10 H13 OCl
M 186 probably C10 H1S OCl
M 218 probably Cl0H120Cl2 M 252 probably Cl0Hll0Cl3 M 286 probably Cl0Hl00Cl4
2,4,6 - Trichlorophenol and di-, tri (2 isomers)- and tetrachloroguaiacol were identified in the spent alkaline extraction liquor. Both spent liquors contained a class of compounds with from 1 to 4 chlorine atoms. The identity of these compounds are not known. The molecular weight of the nonchlorinated molecule is 150. These compounds are probably C 0 terpene derivatives containing C H 0. Similar compounds with two mor~ hydrogens, 10 14 C10 H1S O' were also detected.
- 349-
100.0
RI C
C 60 + 0/'0 - EFFLUENT
500 8:20
I CHO
DCI S
1000 16:40
Isee 2S:00
200B 33:20
Figure 2 Gas chromatogram of an extract of the lipophilic fraction of a spent kraft chlorination liquor.
In this particular mill the pulp is washed with stripped condensate before bleaching. The terpene derivates are probably adsorbed on the pulp during this washing and then chlorinated during the bleaching. The thiophenes are formed during pulping. The thiophenes resisting the pulp washing are then carried over and chlorinated during the bleaching.
These results show that the pulp washing procedure before the bleaching has a large influence on the types and amounts of potentially bioaccumulable organochlorine compounds in the spent bleach liquor.
It is interesting to note that the alkaline extraction stage effluent contains about the same amounts of the chlorinated thiophenes and the terpene derivatives as the effluent from the chlorination stage where the compounds are formed. This is probably due to the fact that these compounds are very lipophilic and they will therefore adsorb to the pulp after the chlorination. During the alkaline treatment the adsorbtion characteristics of the pulp is changed and the compounds are removed. The chlorinated thiophenes and the terpene derivatives are probably adsorbed on particles or associated with high molecular weight compounds in the spent bleach liquors.
- 350-
SCAN TIME
The synthesized chlorinated thiophenes, which were used as standards, have also been investigated in biological tests. They were found to bioaccumulate in fish, and some of them showed mutagenic activity (10). The syntesized compounds are isomers of the chlorinated thiophenes in the effluents. However, these results suggest that chlorinated thiophenes might be partly responsible for the biological effects caused by bleachery effluents. The correct isomers will be synthesized to investigate this.
ACKNOWLEDGEMENT
This work was financially supported by the Royal Norwegian Council for Scientific and Industrial Research and performed within the framework of the Cost 641 project Organic Micopollutants in the Aquatic Environment.
REFERENCES
(1) Kringstad, K.P. and Lindstrom, K. (1984). Spent liquors from pulp bleaching. Environ. Sci. Technol., ~, 236A-248A.
(2) Sodergren, A., Bengtsson, B.-E., Jonsson, P., Kringstad, K., Lagergren, S., Larsson, A., Olsson, M. and Renberg, L. (1987) Biological effects of effluents from pulp industries - summary of results from the swedish project Environment/Cellulose. Presented at the second IAWPR symposium on Forest industry wastewaters, Tampere, Finland, June 1987.
(3) Carlberg, G.E., Kringstad, A., Martinsen, K. and Nashaug, O. (1987) Environmental impact of organochlorine compounds discharged from the pulp and paper industry. Paperi ~ Puu, 69(1987)337-341.
(4) Landner, L., Lindstrom, K., Karlsson, M., Nordin, J. and S~rensen, L. (1977) Bioaccumulation in fish of chlorinated phenols from kraft pulp mill bleachery effluents. Bull. Environ. Contam. Toxicol., ~, 663";'673,
(5) Carlberg, G.E., Drangsholt, H., Gj~s, N. and Tveten, G. (1981). Analysis of organochlorine compounds in water, sediment and fish from Iddefjorden. In: Proceedings from the seventeenth Nordic Symposium on Water Research, Porsgrunn, Nordforsk Publication 1, 1981, pp 131-140.
(6) Neilson, A.H., Allard, A.-S., Reiland, S., Remberger, M., Tarholm, A., Viktor, T. and Landner, L. (1984). Tri- and tetrachloroveratrole, metabolites produced by bacterial O-methylation of tri- and tetrachloroguaiacol: an assessment of their bioconcentration potential and their effects on fish reproduction. Can. I. Fish. Aguat. Sci. 11: 1502-1512.
(7) Carlberg, G.E. and Kringstad, A. (1987). Determination of adsorbable organohalogen in discharges from pulp mills. Submitted to Analytical Chemistry.
(8) Gether, J., Lunde, G. and Steinnes, E. (1979). Determination of the total amount of organically bound chlorine, bromine and iodine in environmental samples by instrumental neutron activation analysis. Anal. Chim. Acta, ~, 137-147.
(9) Renberg, L., Sundstrom, G. and Sundh-NygArd, K. (1980). Partition coefficients of organic chemicals derived from reversed phase TLC. Chemosphere, i, 683-691.
- 351-
(10) Carlberg, G.E., Johnsen,S., Landmark, L.H., Bengtsson, B.-E., Bergstrom, B., Skramstad, J., and Storflor, H. Investigations of chlorinated thiophenes. A group of bioaccumulable compounds identified in the effluents from kraft bleaching. Submitted to Water Science and Technology.
(11) Lindstrom, K. and Nordin, J. (1978). Identification of some neutral chlorinated organic compounds in spent bleach liquors. Svensk Papperstidn. 1:55.
-ill-
INFLUENCE OF HUMIC WATER SUBSTANCES ON THE DEGRADATION OF PAH DURING WATER CHLORINATION
1 G. BECHER1 AND S. JOHNSEN2 National Institute of Public Health, Geitmyrsveien 75,
2 0462 Oslo 4, Norway Center for Industrial Research, Oslo, Norway
Summary
14C-labelled 'benzo(a)pyrene (BaP) was added to three types of water samples a) bog water, b) lake water and c) artificial water. After equilibration the samples were chlorinated with NaOCl solution under buffered conditions. After 72 h BaP and degradation products were extracted consecutively with C~C12 at neutral pH then at acidic pH and finally by absorption to XAD-2. The extracts were subjected to fractionation by reverse-phase HPLC.
In contrast to the artificial water about 20% of the reaction products could 'not be extracted from the humic water samples, This might either be due to formation of more water soluble degradation products in the presence of humic matter or strong complexation of these products to humus. On the other hand, no significant difference in the reaction product profile is observed in the neutral CH2C12 extract of neither of the three water samples, i.e. humus has no effect on the distribution of reaction products in this fraction.
1. INTRODUCTION Chlorination is a widely used technique for disinfection of drinking
water. However, the increasing presence and variety of aquatic pollutants, e.g. PAH, raises the question of the chemical fate of these contaminants when subjected to chlorination (1). PAH, present in many natural waters, are degraded by the reaction with active chlorine and the reaction products may in part contribute to the mutagenic activity found in such chlorinated waters (2),
It has been shown that aquatic humus is able to adsorb/complex PAH (3,4) and may thereby alter the reactivity of PAH towards degradation by active chlorine (5).
The aim of this work was to study the effect of humic substances on the distribution of products from chlorination of benzo(a)pyrene (BaP) in water.
2. EXPERIMENTAL 2.1 Water samples
Three different water samples were used in this work. a) Strongly humic water (21.7 mg DOC/L) from the outlet of a bog outside
oslo named Hellerudmyra. b) Lake water (3.6 mg DOe/L) from Maridalsvannet - a drinking water source
for the city of Oslo.
-ill-
c) Artificial water « 0.1 mg DOC/L) produced from doubly-distilled water by addition of NaCI (3.991 mg/L), Na2 So4 (1.204 mg/L), Ca(NOJ)2·4HzO (0.368 mg/L), MgS04 ·7H2 0 (3.001 mg/L), K2S04 (0.701 mg/L), conc. H2 S04 (1.0 ~l/L).
10L of each water sample was spiked with 60 ~g 14C-Iabelled BaP dissolved in acetone and equilibrated by stirring magnetically for 15 h in the dark.
2.2 Chlorination Prior to chloriniaton, equimolar amount of KH2PO and Na HPO were
added to the water samples to yield a 2 mM phosphate soluiion wi~h PH 7. The chlorination was performed by adding concentrated sodium hypochlorite (7.5%, BDH, Poole, England) to give 30 mg free active chlorine/L for the bog water sample and 3 mg free active chlorine/L for the lake water and artificial water, respectively. Free active chlorine was determined colorimetrically with the Neo-komparator (F. Hellige, Freiburg, FRG).
The water samples were stirred magnetically in the dark for 72 h. Residual free active chlorine was less than 0.1 mg/L for the humic water samples and 1.5 mg/L for the artificial water sample.
2.3 Extraction Each sample was extracted twice with 200 ml CH Cl at pH 7 under mag
netic stirring. After addition of 20 ml conc. HCI, s~mpies were extracted once more with 200 ml CH2CI2 . Dissolved CH2 C12 was then evaporated from the water samples using a rotavapor. The samples were filtrated through a XAD-2 column (17 x 2 cm 1.0.) and the column was eluted with 150 ml methanol. Aliquots of all extracts were taken for scintillation counting.
2.4 HPLC fractionation HPLC fractionation of the concentrated extract was performed on a 25 x
4.6 cm 1.0. Lichrosorb RP-18 column using a 30 min linear gradient from 65% MeOH in water to 100% MeOH and holding 15 min at 100% MeOH. The flow was 1 ml/min. Fractions were taken at 1 min intervals and small aliquots of each fraction used for scintillation counting.
3. RESULTS Table 1 shows the extraction efficiency of radioactively labelled BaP
and reaction products using various methods consecutively.
Table 1. Extraction efficiency of BaP reaction products with different methods used consecutively
Bog water Lake water Artificial water
C~C~ pH 7 48% 55% 85%
CH2C12 pH 2 9% 8% 6%
XAD-2 24% 14% 7%
Residue 19% 23% 2%
100% 100% 100%
-~-
From water containing humic substances (bog water and lake water) only about 50% of the reaction products are extractable with CH CI at neutral pH. In contrast, 85% radioactivity is extracted from cohtr~l water under the same conditions. In all cases only a minor percentage (6-9%) of the radioactivity is extracted after acidification. More radioactive products are recovered from the water by adsorption onto XAD-2 resin. The amount of radioactive products adsorbed to XAD-2 seems to increase with increasing amount of humic matter present in the water sample. About 20% of the reaction products are not extractable at all from the samples containing humic matter while only a very small amount remains in the control water free from dissolved organic matter.
Figure 1 shows the HPLC chromatograms of the neutral CH2Cl2 extracts. No significant amount of unreacted BaP is observed in bog and control water samples after chlorination while about 20% of the radioactivity from lake water is due to BaP.
The radioactivity profiles are very similar for all three water sample types. The main radioactive reaction products are observed in fractions 15/16, 27/28, 29/30 and 30/31.
Figure 1 also shows the HPLC profile for the CH2Cl extract from the bog water sample after acidification. Radioactivity is el~ted as broad peak in the beginning of the chromatogram indicating the presence of a variety of very polar products.
GC analyses of fro 28 to 31 from control water revealed the presence of 3 different compounds the identity of which is presently not known. Fraction 16 from control water turned out not to be gas chromatographable without derivatization.
4. DISCUSSION Benzo(a)pyrene is efficiently transformed to other compounds during
water chlorination independent of the presence or absence of humic substances. However, the extraction efficiency of reaction products from both humic water samples is lower than from control water. This might be either due to adsorption/complexation of degradation products to humic substances or the formation of very polar nonextractable compounds catalysed by humic substances. On the other hand, the HPLC profiles for the neutral CH Cl2 extracts showed that humic substances do not significantly affect the ~istribution of degradation products in this fraction. This indicates that humic matter has little catalytic effect on the reaction between SaP and free active chlorine.
Further work is in progress to identify the reaction products in the neutral exctract.
REFERENCES
1. Jolley, R.; Gorchev, R. and Hamilton, D., Chlorination: Environmental Impact and Health Arbor Science, Ann Arbor, MI (1978).
Jr. (Eds.). Water Effects, Vol. 2, Ann
2. Oyler, A.R., Liukkonen, R.J., Lukasewycz, M.K., Cox, D.A., Peake, D.A. and Carlson, R.M. Implication of treatin9 water containing polynuclear aromatic hydrocarbons with chlorIne: A gas chromatographic - mass spectrometric study. Environ. Health Perspect. 46, 1982, 73-86.
3. Carlberg, G.E. and Martinsen, K. Complexation of organic micropollutants to aquatic humus. Sci. Total Environ. 25, 1982, 245-254.
- 355-
I-11 ~
4. Johnsen,S., Martinsen, K., Carlberg, G.E . , Gjessing, E.T., Becher, G. and Lregreid, M. Seasonal variation in composition and properties of aquatic humic substances. Sci. Total Environ. 62, 1987, 13-25.
5. Johnsen,S., Krane, J., Carlberg, G.E., Aamot, E. and Schou, L. PAH in water systems. Effect of aquatic humus and chlorination, in Organic Micropollutants in the Aquatic Environment, Bj~rseth, A. and Angeletti, G. (Eds.), R. Reidel Publishing Company, Dordrecht, 1986, 440-448.
I f
2 ArUfIcIIIIw.er Bog . 8t ... Add • • trect
1 1 ~
lOG
Figure 1. Reverse-phase HP~C of ~eutf~l and acid CH Cl extracts from water samples splked wlth C-BaP prior tb chlorination.
- 356-
i I -
I 1 '000
INFLUENCE OF WASTE WATER DISINFECTION TREATMENTS ON SOME GENOTOXIC CHEMICAL MICROPOLLUTANTS.
A.Savinol , R.Pasquini l , R.Conti l , C. Melchiorri2 ,A. Di Caro2, L.Sebastiani2 , A.Grella2 , S.Bonacci2 .
1 Dipartimento di Igiene, Universita degli Studi di Perugia, Via del Giochetto, 1- 06100 Perugia,Italy ..
2 Istituto di Igiene, Universita degli Studi "La Sapienza", Piazzale Aldo Moro 5, 1- 00185 Roma,Italy.
Summary
This study compares the effects of different waste water treatments using sodium hypochlorite, chlorine dioxide and ozone on some chemical and biological parameters. The samples studied were obtained from an activated sludge treatment plant of the urban effluents of a large city and from a mixed urban and industrial waste sewage (treated only by sedimentation). A bacterial mutagenicity test (Ames test), microbiological examinations (coliforms, fecal streptococci and coliphages), determinations of the principal chemical parameters and of some organic micropollutants (polycyclic aromatic hydrocarbons - PAR -, and volatile haloderivatives - VHO - ) were carried out on the effluents both before and after disinfection. The results showed that all treated and untreated samples were not mutagenic and that only ozone was able to reduce the concentration of all the examined organic micropollutants. Treatment with hypochlorite decreased only PAR concentration, even though their halogenated derivatives a.nd trihalomethanes were formed. Trihalomethanes were found generally in lower concentrations after disinfections with chlorine dioxide. Hypochlorite showed a good bactericidal activity, but a little viricidal activity, whereas chlorine dioxide and ozone had a very good disinfectant activity both on bacterial and viral forms.
1.INTRODUCTION The presence in surface
micropollutants, resistant to reported by many researchers {S,6,8,2l,27).
and waste traditional and is a
waters of toxic chemical treatment processes has been
cause for health concern
with Some of these ,particularly the genotoxic properties ,can also be
halogenated organic compounds formed during disinfection of
- 357-
water with chlorine (7,20,22,25). However, it is not clear whether alternative disinfectants would also lead to the formation of substances with a potential mutagenic/carcinogenic activity (4,7,12,13,15,16,17,23,24).
This study compared the effects of different waste water treatments, sodium hypochlorite, chlorine dioxide and ozone, on some chemical and biological parameters. The samples studied were obtained from an activated sludge treatment plant of the urban effluents of Rome and from a mixed urban and industrial waste sewage (treated only by sedimentation) of Perugia.
2. EXPERIMENTAL 2.1.Samplings and disinfection treatments
The research was carried out on waste water from Perugia city sewage, which had only undergone sedimentation, and was composed of domestic sewage and industrial sewage (small industrial plants and artisan workshops) and on the effluent of an activated sludge treatment plant in the city of Rome The 50 litre samples of water were rapidly taken to the laboratory and disinfected with one of the following chemicals: sodium hypoclorite, obtained by diluting a commercial product, chlorine dioxide, prepared before use from sodium chlorite, and ozone, produced from oxygen by a small-scale apparatus constructed by the Perugia research group ( Fig. 1 ).
The concentration of the different products was checked by iodometry immediately before use.
Chlorine treatments were done in glass containers with watertight closures. The effluent was additioned with the disinfectant and duly shaken. Instead, ozone was bubbled into the effluent. After 30 min the active chlorine and ozone residues were neutralised with stoicheiometric amounts of sodium thiosulfate. The concentrations of disinfectants used were based on the different characteristics of the sewages. 2.2.Analytical determination
To evidence the effect of the disinfectants under study, before and after treatment, the effluents underwent a series of chemical,microbiological and mutagenicity tests.
Unless otherwise stated the techniques described in Standard Methods (1985) and the Annual Book of ASTM Standard (1978) were followed for the chemical tests.
The following traditional chemical parameters were considered; BODS, COD, N-ammonia, N-nitrous, N-nitric, total phosphorous, total oils and fats. These last parameters were determined by the gravimetric method after extraction of 10 litre samples with trichlorotrifluoroethane by Ultra Turrax homogenizer.
Attempts were made to evidence the presence of organic micropollutants with known mutagenic/carcinogenic potential, such as polycyclic aromatic hydrocarbons (PAR) and volatile halogenated organic
-~-
compounds (VHO). The analytical techniques followed are described later on in the text. 2.2.l.Polycyclic Aromatic Hydrocarbons (PAH)
The dry residue of oils and fats, solubilised in 10 ml cyclohexane, was divided into two aliquotes. One, prior to extraction and purification (9) (fig.2) was used to test for PAH content by HPLC, following the analytical method described by Monarca et al. (18). The other aliquot was used for the mutagenicity test described below.
The hydrocarbons identified and dosed were: fluoranthene, pyrene, benzo(a)anthracene, chrysen, benzo(b)fluoranthene, benzo(k)fluoranthene, benzo(a)pyrene, dibenzo(a,h)anthracene and benzo(g,h,i)perylene. 2.2.2.Volatile Halogenated Organic Compounds (VHO)
Determination of these compounds was carried out following a static head space~as chromatographic method Ziglio and Beltranelli (26), under the analytical conditions described by Melchiorri et al. (14). The haloderivatives identified and determined were: chloroform, chlodibromomethane, bromodichloro-methane, bromoform, methylchloroform, carbontetrachloride, tetrachloroethylene, trichloroethylene. 2.3.Microbiological analyses
The microbiologic analyses were performed mainly to reveal and determine the indices of fecal contamination (fecal coliform bacteria, fecal streptococci and coliphages). 2.4.Mutagenicity assays
The Salmonella/microsome assay (Ames test) was performed on cyclohexane extracts of oils and fats. The tests were carred out with Salmonella typhimurium TA 98 and TA 100 strains. Aroclor-induced male Sprague-Dawley rats were used for preparation of liver post-mitocondrial supernatant (S9), according to Ames et al.(197S).
The oils and fats extracts, dissolved in dimethylsulfoxide (DMSO),were tested in duplicate at increasing concentrations, with and without metabolic activation (± S9 mix), following the standard plate test.
The criteria for positive results were the observation of a dose related response and a 2-fold increase in the number of induced revertants/plate over spontaneous revertants.
Appropriate positive and negative controls were included in the assay.
3.RESULTS AND CONCLUSIONS The mean values of the
and the bar graphs in figures. From the results of the
the following can be said :
assays are given in Tables 1, 2, 3, and 4 3, 4, 5, and 6. traditional chemical parameters (Tab. 1)
a) the analytical values show, even though sometimes in a contradictory fashion, the effects of an oxidating activity on the organic matrix of sewage. These activity was greater with the ozone treatment;. b) the contradictory results, especially those regarding the amounts of BODS and COD determined, cannot be explained by variability in the
-359-
sewage method. evaluate chlorine products
characteristics, the analytical methods employed or the sampling In our opinion this shows that these indices are not able to the characteristics of the treated sewages, especially when dioxide is used. The active residues of chlorine or the of the reaction with sodium thiosulfate, used to neutralise,
could in fact cause considerable analytical interference. Instead, the haloderivatives determined (Tab. 2 and figs. 3 and 4)
were constantly formed when a chlorine based disinfectant was used. Hypochlorite, in particular, caused a notable increase in chloroform (~ia and Roma) and tetrachloroethylene content (Perugia). The latter was probably due to the presence in Perugia sewage of precursors correlated to the industrial activities of the zone.
In previous experiments carried out on drinking water the use of chlorine dioxide resulted in a smaller increase of these compounds (Melchiorri et al. 1987), and though in the present study this characteristic was confirmed it did, in at least one trial (Perugia), lead to the formation of higher concentrations of tetrachloroethylene than those produced when treating with hypochlorite.
Ozone was able to drastically reduce the concentrations of all the haloderivatives identified by use.
It should also be pointed out that in the Perugia sewage all the treatments led to the formation of other VHO which need to be investigated further.
The data relevant to polycyclic aromatic hydrocarbons (PAR) show that all the treatments applied to both the waters under study led to a considerable decrease in single and total PAH (Tab. 3, figs.5 and 6).
The ozone treatment registered the greatest oxidating power followed by dioxide, whereas hypochlorite, although its disinfectant activity was considerable, gave rise to the formation of unidentified hydrocarbons.
With reference to the microbiologic examination (Tab. 4) it is seen that there is a different initial load of the microrganisms which are fecal contamination indicators, due to the different characteristics of the sewages under study. Hypochlorite was effective, especially against bacteric forms, but had little or no effect on the viral forms (coliphages). Th8 degree of reduction of the loads was, strangely enough, greater in the untreated waste waters. Instead, the activity of chlorine dioxide and ozone was valid in both types of effluent and against both the bacteric and viral forms. The most resistant proved to be the fecal streptococci in empirical sewage.
The residue amount of the microbic loads after treatmen~ was, however, proportional to the initial load.
From theese the results (given in this study) it is clear that chlorine dioxide gives the most effective disinfectant action against viral forms followed by ozone and then hypochlorite.
The most effective treatments as regards the bacteric forms proved to chlorine dioxide, followed by hypochlorite and ozone.
-~-
The studied results limited employed
mutagenicity test did not reveal mutagenic activity waters either before or after treatment. We believe are due to the small amount of sewage examined, the doses. of disinfectants used and/or the short contact
in the these
fairly times
which caused the appearance of non-measurable concentrations of mutagenic compound in the 5 litres of sewage tested.
The overall examination of the data obtained, despite the uncertainities due to the occasional appearance of volatile haloderivatives still unidentified by us, and the non-optimal bactericide activity (which can however be improved by bettering the treatment conditions) reveal the net superiority of ozone over the other oxidants. This superiority is particulary important when you consider its capacity for oxidating those organic micropollutants known to be dangerous to health and the environment, and the absence of persistant residual activity.
Comparing hypochloride with chlorine dioxide treatment it can be pointed out that chlorine dioxide reduced PAR con ted and gave rise to lower VHO concentrations, without the formation of other compounds, and had a greater bactericide action. For these reasons, chlorine dioxide proved to be more suitable treatment than hypochlorite which, though maintaining its good disinfectant power, increased rather than decreased the toxic and/or mutagenic hazards associated with waste waters.
REFERENCES
1) Ames, detecting microsome 2) APHA, water and 3) ASTM
B.N., McCann, J., and Yamasaki, E. (1975). Methods for carcinogens and mutagens with the Salmonella/mammalian
mutagenicity test. Mutat.Res.~: 347-364. AWWA, WPCE (l985}."Standard Methods: for the examination of wastewater" 16th Edition .
(1978)." 1978 Annual Book of ASTM Standards. Part 31:Water" American Society for testing and Materials, Philadelphia. 4) Butkovic, W., Klansic, L., Orhanovic, M. and Turk, L. (1983).Reaction of polynuclear aromatic hydrocarbons with ozone in water. Environ. Sci. and Technol, 17 : 546. 5) Cotruvo, J. A. (1985). Organic micropollutans in drinking water.Sci. Tot. Env., 47 : 7. 6) Craun, G. F.(1985). Epidemiologic studies of organic micropollutans in drinking water. Sci. Tot. Env., 47 : 461. 7) Dolara, P., Ricci, V., Burrini, D., Griffini, O. (1981). Effects of ozonation and chlorination on the mutagenic potential of drinking water. Bull. Env. Contam. Toxic.,27, 1. 8) EPA-USA ( 1980). Water quality criteria documents; availability. Federal Register, 45 (231) : 19318.
- 361-
9) Grzybowski, J., Radecki, A. and Renkowska, G. (1983). Isolation, identification and determination of polycyclic aromatic hydrocarbon in sewage. Environ. Sci.and Technol, 17 (1) : 44. 10) IRSA.(1973). "Metodi analitici per Ie acque" Quaderni dell'Istituto di Ricerca sulle Acque (CNR) n. 11 vol. 111°. 11) IRSA (1984). "Metodi di analisi per Ie acque di mare" Quaderni dell'Istituto di Ricerca sulle Acque (CNR) n. 59: 540. 12) Janssens, J. G., Meheus, J. and Dirick (1985). Ozone enhanced biological activated carbon filtration and its effect on organic matter removal, and in particular on AOC reduction. Wat. Sci. Tech., 17 : 1055. 13) Legube, B., Langlais, B., and Dore, M. (1980). Reactions of ozone with aromatics in dilute aqueous solution Reactivity and biodegradability of oxidation products. Prog. Wat. Tech, 12 : 553. 14) Melchiorri,C., Grella, A., Di Carlo, A., and Bonacci, S. (1985).Indagine 5ulla presenza di composti organo alogenati volatili nelle acque potabili della citta' di Roma. Nuovi Annali di Igiene e Microbiologia, 36 : 1. 15) Monarca, S., Meier, J. R.,and Bull, R. J. (1982). Evaluation of drinking water treatment and distribution system by bacterial short-term mutagenicity assay. Presentato al 3 Congresso Internazionale della NIA, 15-18 settembre 1982, Spoleto. 16) Monarca,S., Pasquini,R, mutagens in unconcentrated and
and Arcaleni,P. (1985). Detection of concentrated drinking water supplies
before and after treatment using a microscale fluctuation test. Chemosphere, 8 : 1069. 17) Monarca,S., Pasquini,R., and Scassellati Sforzolini,G. (1985). Mutagenicity assessment of different drinking water supplies before and after treatments. Bull.Environ.Contam.Toxicol., 34 : 815. 18) Monarca, S., Pasquini, R., Scassellati Sforzolini, G., Savino, A., Bauleo, F. A., and Angeli, G. (1987). "Environmental monitoring of mutagenic/carcinogenic hazards during road paving operations with bitumens". Int. Arch. Occup. Environ. Health, 59 : 393. 19) Morozzi, G., Savino, A., Conti, R.,and Manenti, R. (1986). The fate of some organic pollutants in wastewater during ozone and ozone-GAC adsorption treatments. J. Environ. Sci. Health, A21 (6), 523. 20) Pana', A., Patti,A. M., Zaratti, L., Grella, A., and Paroli, E.(1984). Chlorine action on the hepatitiS A virus infectivity. L'Igiene Moderna, 81 : 454 . 21) Petrasek, A. C., Kugelman, I. S., Austern, B. M., Preslley, T. A., Winslow, L.A., and Wise, R. M. (1983). Fate of toxic organic compounds in waste water treatment. J.W.P.C.F., 55 : 1286-1296. 22) Rannung, U. (1981). Mutagenic effects of effluents from chlorine bleaching off pulp. J. Tox. Envir. Health, 7 : 33. 23) Rice, G. (1985). Ozone for "drinking water treatment.Evolution and present status. In: "Safe Drinking Water: The Impact of Chemicals on a Limited Resource". G. Rice (ed.). Drinking Water Research Foundation. 24) Scassellati Sforzolini, G., Pasquini, R., Savino, S. and Conti, R.(1985). Influenza del trattamento con cloro e con ozono sulla
- 362-
attivita' mutagena di Congresso Nazionale di Bari 2-4 ottobre 1985.
acque contaminate con 3,4-Benzopirene. XXXII ° Igiene e Medicina Prevent iva e Sanit~' Pubblica.
25) Szejnwald, B. H., Bishop, D.R.,and Rowan, C.A.(1984). The role of skin absorption as a route of exposure for volatile organic compounds (VOCs) in drinking water. Am. J. Publ. Health, 74 : -479. 26) Ziglio, G. ,and Beltramelli, G. (1980). "Determinazione di idrorocarburi alifatici alogenati (C1-CS) nelle acque potabili con la tecnica dello spazio di testa statico". Quaderni Istituto di Igiene -Universita' di Milano, n.2,luglio (1980). 27) WHO (1984). Guidelines for drinking water quality. Vol. 11°. Health criteria and other supporting informations. Geneva.
« H
" ::> gj
""
~ IX:
Table 1 . PHISICO - CHEMICAL PARAMETERS DETERMINED IN THE WATERS UNDER STUDY , BEFORE AND AFTER TREATMENT WITH SODIUM HYPOCHLORITE , CHLORINE DIOXIDE AND OZONE (CONTACT TIME 30 MIN)
Parameters Untreated NaC10 C102 03 assayed
(ppm) (lOmg/!) (lOmg/l) (4.8mg/l)
pH 8.40 8.60 8.60 8.00 N-ammonia 30.00 25.00 25.00 15.00 N-nitrous -- n.d(*) - 0.07 N-nitric -- 1.03 1. 03 1. 03 B.O.D. 5 98.30 40.30 152.00 51.50 C.O.D. 240.00 250.00 350.00 148.00 Total phosphorous 3.80 3.50 4.20 3.50 Total oils
and fats 23.81 25.43 23.62 13.47
Parameters Untreated NaC10 C102 03 assayed
(ppm) (5mg/l) (5mg/l) (2.5mg/l)
pH 7.80 7.90 7.30 8.30 N-ammonia 16.00 15.00 15.00 13.50 N-nitrous 0.17 0.12 0.06 0.03 N-nitric 0.26 0.39 0.20 0.91 B.O.D·S 13.30 10.00 51. 60 8.20 C.O.D. 29.60 41.60 120.40 40.00 Total phosphorous 2.40 2.40 2.30 2.30 Total oils and fats 1.06 1.10 0.90 0.60
(*)n.d.= non determinable
- 363-
I I
< H l!J ;:J ~ f;<:I p..
~ ~
Table 2 . ANALYSES OF TOTAL HALODERIVATIVES IN THE STUDIED WATERS BEFORE AND AFTER TREATMENT WITH SODIUM HYPOCHLORITE , CHLORINE DIOXIDE AND OZONE ( CONTACT TIME 30 MIN).
Compounds (ppb) Untreated NaClO Cl02 03 (lOmg/l) (lOmg/l) (4.8mg/l)
Chloroform 1.130 2.980 1. 750 1.330 Chlorodibromomethane - - 0.340 -Bromodichloromethane - 0.030 0.550 0.100 Bromoform - - - -Methylchloroform 0.090 0.040 0.040 0.010 Tetrachloroethylene 13.910 16.040 27.660 0.320 Trichloroethylene 2.540 1.650 1.600 0.240 Carbontetrachloride - 0.013 - 0.009 Other (*) + ++ ++ ++
Total VHO 17.770 20.753 31.540 2.009 Variation % (+17) (+78) (-89)
Compounds (ppb) Untreated NaClO Cl02 03 (5mg/l) (5mg/l) (2.5mg/l)
Chloroform 0.370 0.700 0.330 0.040 Chlorodibromomethane - - 0.090 -Bromodichloromethane - - - -Bromoform - - - -Methylchloroform 0.050 0.050 0.040 -
Tetrachloroethylene 0.270 0.130 0.140 0.070 Trichloroethylene - - - -Carbon tetrachloride 0.007 0.021 0.008 0.008 Other (*) - - - ++
Total VHO 0.697 0.901 0.608 0.118 Variation % (+29) ( -13) (-83)
(*)Presence of unidentified VHO with relative retention time to chloroform, 0,702 (+) and 0,786 (++).
- 364-
..: H 0
~ r>:< p..,
~ 0 ~
_. ---
Table 3. ANALYSES OF POLYCYCLIC AROMATIC HYDROCARBONS (PAR) IN THE STUDIED WATERS BEFORE AND AFTER TREATMENT WITH SODIUM HYPOCHLORITE,CHLORINE DIOXIDE AND OZONE (CONTACT TIME 30 MIN.).
P A H (ppb) Untreated NaClO(*) Cl02 03 (lOmg/l) (lOmg/l) (4.8mg/l)
Fluoranthene 104.0 58.4 31.4 13.8 PJ:rene 104.3 33.6 25.0 15.0 Benzo(a)anthracene 35.0 15.2 5.9 3.4 Chrysen 15.3 11.7 2.6 2.1 Berrzo(b)fluoranthene 51.0 31.0 24.0 16.0 Benzo(k)fluoranthene 16.8 10.9 2.2 1.3 Benzo(a)pyrene 18.5 10.8 2.9 2.4 Dibenzo(a,h)anthracene 21.6 10.3 - -Benzo(g,h,i)perylene 8.0 4.9 3.2 -
Total P A H 374.5 186.8 97.2 54.0 Variation % (-50) (-74) (-85)
P A H (ppb) Untreated NaClO(*) Cl02 03 (5mg/l) (5mg/l) (2.5mg/l)
Pyrene 50.3 37.1 30.3 10.0 Benzo(a)anthracene 17.3 11.8 11.1 2.0 Chrysen 23.6 16.7 15.1 6.8 Benzo(b)fluoranthene 17.4 9.8 6.2 3.3 Benzo(k)fluoranthen 43.5 24.4 34.0 3.7 Benzo(a)pyrene 14.9 8.8 7.4 1.0
Total P A H 167.0 108.6 104.1 26.8 Variation % (-35) (-38) (-84)
(*) Presence of unidentief PAR with retention time included to: Pyrene and Benzo(a)anthracene Benzo(b)fluoranthene and Benzo(k)fluoranthene; Benzo(a)pyrene and Dibenzo (a,h)anth~
- 365-
<: H
'"' ii;l fJ:l ~
~ ~
Table 4. MICROBIOLOGIC ANALYSIS AND MUTAGENIC TEST OF WATERS UNDER STUDY,BEFORE AND AFTER TREATMENT WITH SODIUM HYPOCHLORITE CHLORINE DIOXIDE AND OZONE (CONTACT TIME 30 MIN)
Parameters Untreated NaC10 C102 03 (lOmg/l) (lOmg/l) (4.8mg/l
Total Coliforms 1.1x198 <3.0 <3.0 7.Sx102 /lOOml
Fecal Coliforms 9.Sx107 <3.0 <3.0 2.3x102 /lOOml
Fecal Streptococci 6.4x107 9.1x10 7.Sx102 2.1x104 /lOOml
Coliphages 2.4x10S 2.4103 7.5 1.lx10 /lOOml
Ames Test neg. neg. neg. neg.
Parameters Untreated NaC10 C102 03 (Smg/l) (Smg/l) (2.Smg/l
Total Coliforms 7.SxlOS 7.3 <3.0 l.SxlO /lOOml
Fecal Coliforms 9.1x103 <3.0 <3.0 <3.0 /lOOml
Fecal Streptococci 2.3x104 9.1 <3.0 7.3 /100ml
Coliphages 9.3xl03 2.4x103 0.3 0.3 /100ml
Ames test neg. neg. neg. neg.
- 366-
o X I G E
N OZONIZER CONTIICTOR
FLOW CONTROL
FIGURE 1 - OZONE TREATMENT MODEL
5 ml CYCLOHEXANE SOLUTION
1 ADD 100 ml OF CYCLOHEXANE IN THE SEPARATORY FUNNEL
1 ADD DIMETHYLFORHAMMIDE (DHF) - WATER ( 90 ml + 10 ml )
~ MIX for 5 min
1 ADD 200 ml of CYCLOHEXANE AND 100 ml of WATER to DMF : H20
{-
MIX for 5 min
'>I
------------------~~
WASH the CICLOHEXANE (200 ml) with 50 ml of WATER (add NaCl to avoid emulsion)
t Mix for 5 min
EVAPORATE CYCLOHEXANE AT SMALL VOLUME (3-4 ml)
1 FILTER on an Si02 COLUMN 107. ACTIVATED
I v
ELUTE with 120 ml CYCLOHEXANE
COLLECT 110 ml of CYCLOHEXANE
1 EVAPORATE at SMALlJ VOLUME (1-2 ml)
'"
UPPER LAYER DISCARD (ciclohexane)
DISCARD DMF H20 LAYER
DISCARD WATER LAYER
DISCARD FIRST 10 ml
FIGURE 2 -HPLC ANALYSIS EXTRACTION AND DETERMINATION OF PAH
- 367-
o ::c :>
o
27.66
\6.04
13.9\
CICH=CC12
FIGURE 3 - VOLATILE HALODERIVATIVES (VHO) IN THE WATERS UNDER STUDY, BEFORE AND AFTER TREATMENTS
32
30 28 26 24
22 20
18 ~
16 .. 14 12
10 8
6 ROMA 4
2 0
"0 0 0 0 '" ~
u -'
'" '" U
<I> Z L
c: ::l FIGURE 4 - TOTAL VHO
- 368-
110
100
90
..: 80 H
~ 70
~ 6
70
60 s SO
40 .a 8: 30
o UNTREATED
13 NaClO
~ ClOZ
III 03
FIGURE 5 - POLYCYCLIC AROMATIC HYDROCARBONS (PAH) IN THE WATERS UNDER STUDY BEFORE AND AFTER TREATMENTS
80 60 40 20
300 80 60 40 20
200 .c 80 '" '" 60
40 20
100 80
60 ROMA 40 20 0
"C 0 '" 6' ~ -' 0
'" u U
'" '" '- z FIGURE 6 - TOTAL PAH
C ::l
- 369-
SESSION V
MATHEMATICAL MODELLING
Chairman R. SCHWARZENBACH
EV8.luation of some chemical fate and transport models - A case study on the pollution of the Norrsundet Bay (Sweden)
Modelling of groundwater transport of microorganic pollutants : State-of-the-art
EVALUATION OF SOME CHEMICAL FATE AND TRANSPORT MODELS. A CASE STUDY ON THE POLLUTION OF THE NORRSUNDET BAY (SWEDEN).
K.Kolset, B.F.Aschjem, N.Christophersen, A.Heiberg, B.Vigerust. Center for Industrial Research, P.O.Box 124 Blindern, N-0314-0SLO-3,Norway.
SUMMARY
Mathematical models in the marine environment can be grouped into several classes. These models may be splitted into physical, chemical and toxicological models. Further they may be classified by their geographical extension, their accuracy and/or amount of input data needed. This paper gives a brief discusson of the main classes of models.
We have used three different models to investigate transport and fate of chemicals in the aquatic environment. In this paper the models FEQUM, EXAMS and QWASI are presented, their characteristics explained and a comparison of the models is made.
The three models have been applied to the Norrsundet area. Norrsundet is a heavily polluted bay on the eastcoast of Sweden. The pollution is mainly due to a kraft mill located in the area. The models were calibrated using data on chloroform in wastewater, and tested on four chlorophenolics. All models give satisfactory results for the compounds investigated exept for Tetrachlorocatechol. Correlation coefficients between calculated and measured concentrations vary from 0.86 to 0.97. The results obtained for tetrachlorocatechol are probably due oxidition of TeCC before reaching the first compartment.
INTRODUCTION
During the past two decades methods in analytical chemistry have been through a revolution. The amount of accurate data which can be obtained on a routine basis has increased dramatically. This revolution has created a need for developing methods that can handle analytical data in an efficient way and extract from them useful information. One way to utilize the data is to use them as the basis for a mathematical modelling of the processes and reactions occuring in nature. Development of 50-called fate and transport models is one example of such an effort. Others are developing of pure dilution models, models describing transport of a compound across a boarder zone between two phases and models describing uptake of a compound to an organism.
Models can be grouped into several classes. One may distinguish between compartment models and continuous models, between models modelling the micro cosmos and models modelling the macro cosmos, between the high precession and the more coarse models.
Use of models is generally aimed at i) improving the understanding of the
-m-
system studied, e.g. by testing various hypotheses about the system, and ii) predicting future developments under various scenarios. Before predictions can be made confidence in the model must be established. A common procedure is to calibrate the model, i.e. determine the model parameters in terms of one particular set of data, and then test its predictive capability by attempting to reproduce another data set while keeping the model parameters unchanged. In spite of the advances made in analytical techniques, collecting the data needed to perform a satisfactory model validation remains a. costly and time-consuming task.
MODELS FOR AQUATIC DOMAIN
The models used in the aquatic domain can be grouped into several classes. Firstly, there are two principal different ways of modelling the fate and transport of chemicals in the aquatic domain; soft and hard modelling. The soft models have been developed during the last years in the field of chemometrics. They describe the different relationships in nature using Principal Component Analysis (PCA) or Partial Least Squares (PLS) techniques. These methods establish prediction models using measured data, without knowing the exact processes in nature. The hard models are based on known or estimated differential equations representing important processes in nature and represent the traditional way of modelling the environment. In this paper we will concentrate on the hard models.
The hard models can further be splitted into physical, chemical and toxic effect models. There are, however, no clear lines between these classes of models. In example, WASP (Water quality Analysis Simulation Programme) (1) contains a physical, a chemical and a toxicological part. The models known as fate and transport models are classified somewhere between a physical and a chemical model. Figure 1 shows a classification of the different models used in the aquatic domain.
The physical models can be divided into compartment and continuous models. The compartment type of models assumes that each segment is homogenous
WASP
Physical properties
PHOENICS
Chemical ==) properties ==)
Figure 1. Classification of models for the aquatic domain.
- 373-
throghout its volume. The water and sediment volume is horisontally andvertically divided into segments. The compartment type is usually used to describe the transport and fate of a chemical in the marine environment. Alternatively, in the continuous models the system can be modelled using differential equations as in hydrodynamic modelling. These models are usually pure dilution models, modelling only the hydrodynamic part of the system. One example is the Phoenics (2) model.
The physical parameters included in the model may be topographical data, wind direction and speed, tidal flows, water and sediment area, volume, and advective or circular flows.
The chemical models describe phenomena as diffusion, chemical reactions and degradation and fate and transport of a compound. Common for these models are their use of compound specific parameters, e.g. physical, chemical and reaction parameters and the distribution of the chemical compound between phases. A large group of chemical models describe the fate and transport of chemicals in the marine environment. This information combined with load data and hydrodynamic data from a physical model make it possible to calculate the concentration gradient in a continuous system or the concentration in each compartment.
The fate and transport models, FEQUM and EXAMS, calculate the concentrations of a compound in different compartments. The QWASI model can handle only one compartment in each phase.
Models concerning the toxicological effects on the environment are another type of models. The toxicological models are used to describe the effect of a chemical compound on the biological and ecological environment.
Output from a physical model can be input to a chemical model and output from a chemical model can be used as input for a toxicological model. This is indicated with the arrows in figure 1. However, the opposite direction of the data flow is not possible. The WASP modelling system contains these three parts and the data flow is icorporated in the system.
Modelling the aquatic domain has usually been concentrated on site specific models or models bound to specific systems and compounds, resulting in models containing small pieces of the above mentioned types of models. Generalizing such models results in ad hoc and inaccurate solutions. The reason might be:
* A model generated for a specific area often gives better accuracy than a general model calibrated to an area. The accuracy will also depend on
- the exactness of the equations involved. A good description of the study site and its configuration will give better accuracy.
* The uncertainty of the model results depend on the model's accuracy and the uncertainty of the input data. Many uncertain parameters make the uncertainty of the results even greater.
The more elaborated a model is, the more computer resources are necessary. Besides, the more complex a model is the more painful is it to keep the full control over it. This is a limiting point in the modelling work. The modeller must usually reduce the model to fit a reasonable limit in expences in computer time and human frustration.
-374-
The fate and transport models used for our purpose might be a best approach. All models are the general type, adapted to a specific site by calibration. The models can handle up to 50 compartments and they use compound specific information. They all run on a usual IBM PC compatible.
DESCRIPTION OF EXAMS. FEOUM AND OWASI.
In the present study three different models of the compartment type, EXAMS, FEQUM and QWASI, have been applied in studying the dispersion of industrial pollutants in a coastal area.
EXAMS (EXposure Analysis Modelling System) was developed by the US Environmental Protection Agency (EPA) in the late 1970s and early 1980s (3). EXAMS is an aquatic model, i.e. it deals with the water column and benthic sediments. It is very well documented and is probably the most widely known model of its kind.
FEQUM (Fugacity EQUilibrium Model) and the QWASI (Quantitative Water, Air, Sediment Interaction) model are different modifications of the fugacity model developed by Mackay and coworkers (4-8) since 1979. The fugacity model is a multimedia model comprising water, air, soil, sediments and biota. FEQUM, which has been designed at our own institute, comprises the same media. Its major new feature is that it can handle an arbitrary number of compartments within each physical medium (the original fugacity model could treat only one compartment of each type). Furthermore several key model parameters are made time-dependent and the integration routine used is changed from a simple Euler to a Runge-Kutta algorithm of fourth order.
QWASI was developed by the authors of the original fUgacity model (9,10) Like EXAMS, the model is limited to the aquatic domain (water and sediments). Therefore, in a certain sense it may be regarded as a simplification of the original model, which also included air, soil and biota. However, a number of refinements relative to the parent model have been introduced. In particular, sedimentation processes are described in a more elaborate way than in the fugacity model. The QWASI model has been implemented in two versions: a single compartment (9) and a linear continuous (10) variant.
While the above models may appear quite different at a first glance, they rely on the same basic principles, the difference lying only in the detailed formulation of physical-chemical processes and in the limitations pertaining to system definition. In the following paragraphs the common basis of the models will be outlined and the the most important differences between the models will be discussed.
Basic principles. In the models, the compound investigated is thought of as being carried among the system compartments by flowing water, either in dissolved form or bound to particles present in the water body. In EXAMS mobile planktonic matter is also considered as a possible carrier of the chemical. Once the flow rates of water and suspended sediments across the compartment boundaries have been specified, the rate of change in chemical mass, m(i), in each compartment can be determined simply as the difference between the mass entering the compartment (Q. (i)) and the mass being removed from the compartment either through t2ansport to other compartments or to the external world (Qt (i)), or through loss processes (QI (i)): rans oss
-lli-
dm(i)/dt = Qin(i) - Qtrans(i) - Qloss(i). (1)
In addition to the transformation reactions accounted for by eq 1 the chemical will also participate in other processes within each compartment. In particular, there will be a partitioning of the chemical between the water and solids . All three models investigated assume that partitioning between the water and suspended particulates and between the pore water and the solid matter in the bed sediments is in equilibrium. The justification of these assumptions is that sediment-water exchange is fast relative to the processes incorporated in eq 1. The same is true for various other chemical processes of concern, including ionization reactions of acids and bases and complexation with dissolved organic carbon: (Such rea~tions can only be handled by EXAMS).
Masses in eq. 1 may be expressed in terms of chemical concentrations (c) and the volumes of the system compartments. The characteristic feature.of FEQUM and QWASI is the use of fugacity instead of concentration. The fugacity of the chemical is related to its concentration by the equation
c = Z·f , (2)
where f is the fugacity and Z is the so-called fugacity capacity. The fugacity is a measure of a chemical's tendency to escape from the phase in which it exists to adjacent phases. It has the dimension of pressure. In general, the fugacity capacity depends on both the phase and on the chemical itself. Within limited ranges of the concentration it is roughly constant. Express~ons for.z for various.types o!3me~fa ~ave.been given by Mackay (4). For alr Z, WhlCh has the unlt mol m Pa , lS slmply 1/RT.
The principal advantage of introducing fugacity is that it enables all concentrations to be expressed in a single unit. This facilitates the interpretation of calculated results to an appreciable extent. In particular, assessing the relative importance of the relevant transport and transformation processes can be easily done.
In one particular context, fugacities do have a more fundamental significance than do concentrations. As is well known from chemical thermodynamics, stability constants for chemical equilibria are true constants only if expressed in terms of activities, or fugacities, rather than in concentrations. This means that FEQUM and QWASI have the potential of treating water-sediment partitioning more adequately than EXAMS.
Water and sediment flows. EXAMS considers two distinct modes of water transport, advection and turbulent dispersion. The rates of water flow between system compartments due to advective flows are determined from a water mass balance set up on the basis of information on flow paths and input flow rates. Turbulent dispersion denotes the exchange of water between adjacent system compartments arising from various small-scale advective processes.
FEQUM also takes into account two types of water movement. In addition to advective flows, which are treated in essentially same way as in EXAMS, the model can describe circular currents. These currents may extend over two or more water compartments. Circular currents involving ~ (adjacent) compartments, correspond to the dispersive water exchange defined in EXAMS.
-m-
The QWASI model accounts for advective flows only.
Obviously, FEQUM is somewhat more flexible than EXAMS and QWASI with regard to the treatment of water transport. In fact, FEQUM allows one to define flow patterns in a completely free manner within the framework of a compartment model.
In EXAMS suspended particulate matter is thought of as a non-conservative substance that is carried passively with the water masses. The rate of flow of particulates between two interconnected compartments is calculated from the corresponding water flow and the particulate concentration in the source compartment. Exchange of solids between the water column and benthic sediments is treated in a similar manner (see below).
The QWASI and FEQUM models differ from EXAMS in that sediment transport through the water column is treated independent of the water flow. Thus, in these models sediment transport is not entirely driven by the hydrodynamics of the system. Input to QWASI and FEQUM pertaining to sediment transport includes inflow and outflow of particles, sedimentation and resuspension rates plus the total volume of suspended sediments.
Sediment-water transfer. QWASI, FEQUM and EXAMS regard the transfer of compounds between water and sediment as a process involving two mechanisms: a) an exchange of water between the pore-water in the bed sediment and the overlying water masses, and b) a resuspension and deposition of solid sedimentary material.
In the three models the first process accounts for diffusive transfer between the water phase and the pore-water caused by chemical concentration gradients. with regard to the interpretation of the second mechanism, however, there seems to be a certain difference between the models. In EXAMS sediment-water transfer is thought of as involving a dispersive exchange of volumes between the water column and the sediment layer. The process includes the following physical events: a saltation of a unit volume of bed sediment, a subsequent equilibration with the water column and a resettlement on the bed. QWASI and FEQUM also considers water and solids exchange between the sediment and water phases. These models treat the two types of transport processes independently of each other. Furthermore, they describe a simple mass transfer of solids between the two phases and do not require the exchange of solid matter to be symmetrical.
EXAMS cannot account for situations in which a net sediment deposition is taking place, which is a common case. Another implication of the restrictions inherent in this model is that with equal partition coefficients for the suspended particulates and the bed sediments, steady-state concentrations of the solute are bound to be the same in the water column and the pore-water, provided no degradation reaction takes place in the sediment layer. The same holds true for the sorbed fractions in the two phases.
Loss and transformation processes. All three models assume chemical transformation and loss processes to be of first order in the solute. Processes of concern include volatilization, direct photolysis, hydrolysis, radical oxidation and microbial degradation. Although the processes are first order, the overall reaction order may be higher. Oxidation, for example, also depends on the amount of oxidants present in the water. In such cases a ·pseudo· first-order rate constant must be estimated to obtain
-m-
the correct reaction rate. EXAMS is more elaborate than the other two models in that it calculates (pseudo) first-order constants for each individual reaction from basic chemical data provided by the user. The model also adjusts rate constants according to system temperature and to other relevant environmental parameters.
In FEQUM the air column is considered as part of the total system. volitilazation in this model, therefore, is a transport process rather than a loss process. The treatment of volatilization is based on the same theory in all three models, the two-layer film model (11).
Both EXAMS and the fugacity models of Mackay and coworkers were developed primarily for application to rivers, lakes and ponds. The hydrodynamics of these systems is therefore simple. In estuaries and other coastal areas, water movement is generally much more complex, it is driven by a number of different mechanisms, including tidal waves, wind, the occurrence of water density gradients and stable large scale currents. It is not obvious that models like those studied here are capable of giving a realistic picture of the flow pattern. In particular, it is questionable whether the combined effect of such mechanisms can be lumped into a set of dispersion coefficients, as is implicitly assumed when EXAMS is employed. Therefore, as applied in the present study, EXAMS, FEQUM and the QWASI model are to be considered as essentially empirical models.
CHEMICAL DATA
The fate and behaviour of a compound in the environment are to a large extent dependent on its physico-chemical properties. Table I shows some relevant data for chloroform and the other four compounds used in validating the investigated models. The degradation processes taken into account are oxidation and base-promoted hydrolysis. Hydrolysis was assumed to ~e ifPortant only for chloroform. The rate constant was set to 0.23 . 10~ h~ , assuming pH = 7. In the present calc~ld-at~pns the following -3 o~fdatlon rate const~~ts were us~?: Cf: 0.7,10 _2h -1 2,4,6-TrCP: 1.0,10 h ; 3,4,5-TrCG: 0 h ; TeCG:O h ; TeCC:0.5·10 h . The rate constants for the first two comp~unds were obtained from (12) assuming an oxidant concentration of 1'10- M. For TeCC the constant was estimated from (13).
TABLE I Chemi~alLfb~si~~l f~'~m~t~rs Used
Compound Abbrev- Mol. Vapour Water Henry's law log, iation weight Solubility consttnt Pow pressure
gLmole Pa eem ~tm, m LmQl 2,4,6-Tri~hloro-
1.594 800.0· '10-6 A 3.61· J2.henol 2, 4, 6-TrCP 197.5 4.0 3,4,5-Tri~hloro-
0.648 9.10c * guaiacol 3,4,5-TrCG 227.47 4.13 Tetra~hloro-
0.1388 4.20C * guaiacol TeCG 261.92 4.42 Tetra~hloro-
5.3'10- 4 c 2. 7~C * ~atechol TeCC 247.89 - 4.19. Chloroform Cf 119.38 2.1104 • 8200.0 2.88'10- 3 • 1. 97
1 Pow: octanol-water partition coefficifnt Data are collected or estimated from: =(12) 8 =(14) c =(15) = (16)
- 378-
Norrsundet
Lindon
Gulf of Bothnia
Iggon
Figure 2. The Norrsundet area about 230 km. north of stocholm.
THE NORRSUNDET STUDY
The Norrsundet area is a heavily polluted bay on the east coast of Sweden, situated about 230 km north of Stockholm (Figure 2). The main source of pollution is a kraft mill (Norrsundet Bruk) which dircharges wastewater into the nearby coastal water at a rate of about 1 m Is. The effluent is mostly uncharacterized. A few individual compounds have been identified, however, including chloroform, some chlorophenols, chloroguaiacols and chlorocatechols. Generally the Gulf of Bothnia has a high background level of these compounds (17) probably as a result of discharge from a s.eries of pulp and paper mills located around the gulf.
The effluent from Norrsundet Bruk is discharged into a shallow pond. Into this pond seawater flows from the south and is pumped to the north at a
TABLE II Pollutant Loads
Compound Conc. in effluent Discharge via effluent Background level Ilg/l g/h ng/l
Chloroform 440 1247 - 1584 30 2,4,6-TrCP 22 62 - 79 30? 3,4,5-TrCG 117 328 - 421 10 TeCG 92 258 - 331 2 TeCC 217 609 - 781 40?
- 379-
IABLE HI ~~Qgraehi~al Qata, Norrsundet greg
Compartment Area Water Water Water Air Sediment Name Number volu9le m.depth residence t. volure volure
106 m2 10 m m days 10 m 10 m
Milgrund 1 4.4 11.5 2.6 0.5 4.4 0.44 Svartgrund 2 3.7 20.5 5.5 3.7 0.37 Klunken 3 10.1 60.7 6.0 10.1 1.01 Klackarna 4 36.4 581.0 16.0 36.4 3.64 R~dh<ellan 5 17.8 147.0 8.2 0.9 17.8 1. 78 Utposten 6 7.9 200.0 25 0.8 7.9 0.79
rate of about 20 m3 Is. The main current in the Gulf of Bothnia moves from the north to the south along the Swedish coast with an average speed of 0.5 knot (18,19). Besides being influenced by this current, water motion and pollutant transport in Norrsundet are to a large extent determined by the wind, which blows predominantly from the southwest.
We did not have the opportunity to carry out a measurement program specifically designed for model validation. Instead we have used existing data from previous studies.
Measurements of chloroform and various chlorophenolics in water and sediments in Norrsundet have been carried out by Xie et al. (17) on four different occations between September 1982 and November 1983. The sampling points were well distributed over the investigated area, but the vertical resolution was rather poor. Pollutant concentrations in the effluent were measured on two occations, chloroform on only one. No information on the number of samples taken at each station or on the precision of the measured concentrations.is given. Such information would have been of great value in the calibration and validation of the models investigated.
Mean observed chloroform concentrations were determined for each water compartment by averaging the concentrations measured on November 3, 1983, (17) horizontally over the whole compartment area. The values thus obtained are given in Table IV. All measurements were made 5 meter below the sea level.
Combining the data on effluent concentrations with information on the effluent flow rate enables one to estimate the discharge rate of the relevant compounds. T~e flow rate is somewhat uncertain but was probably in the range 68-86 000 m Iday on November 3, 1983. Effluent concentrations measured on that date (17) and corresponding discharge rates are given in Table II. Since additional chloroform may be formed after the effluent has left the kraft mill, the calculated values for chloroform may underestimate the amount of chloroform flowing into the receiving area.
Partitioning of receiving area. The division of the receiving area into compartments closely follows that employed by Jonsson (20). The partitioning used is shown in Figure 3. The compartment names and their geometrical dimensions are given in Table III. Except for "utposten" and "Klunken", in which cases "data were obtained from a sea chart, volumes, areas, mean depths and residence times were taken or calculated from the work of Jonsson (20).
- 380-
ROdhdllan
Figure 3. The right figure shows the partitioning of Norrsundet, while the left one showssthj flows defined. The estimated flow rates are (all figures in 10 m Ih): W-1 11.0, W-2 27.1, W-3 6.8, W-4 .072, W-5 1.9, W-6 0.9, W-7 15.0.
Each subarea was split horizontally into two layers, a water and a bed sediment layer. The depth of the sediment was set to 0.1 m in all cases. In FEQUM an air layer is also needed. The height of the air column was chosen as 1 m. (An average residence time of 1 hour used for all air segments.)
using the calculated residence times and compartment volumes, a total flux of water across the boundary of each water column compartment can be estimated. Assumed flow directions and associated estimated flow rates are shown in Figure 3.
Results from EXAMS and FEOUM calculations. In a first step, EXAMS and FEQUM were calibrated to give a best possible fit to the average observed water concentrations of chloroform, given in Table IV. Chloroform was used as a tracer by Xie et al. (17) to determine the dilution of the effluent from Norrsundet Bruk. The compound is relatively inert and has little tendency to adsorb to sediments. These properties make it well suited for calibration purposes. The flows W-2, W-5, W-6 and W-7 (see Figure 3) were chosen as the basic adjustable parameters in the calibration.
With the estimated discharge rate of chloroform (Table II), it proved impossible to reproduce the observed ambient concentrations. Both models invariably predicted too low concentrations in the two innermost compartments. Therefore, emphasis was shifted from reproducing absolute values to predicting correctly the relative magnitude of the chloroform concentrations in the various compartments. In order to achieve this the
- 381-
TABLE IV Mean water concentrations in Norrsundet on 3rd of November 1983
Compound Milgrund Svartgrund Klunken Klackarna R~dh~llan Utposten
Chloroform 4500 2800 1000 360 50 2,4,6-TrCP 235 52 60 31 3,4,5-TrCG 1050 360 161 22 15 15 TeCG 850 290 119 13 9 8 TeCC 119 110 97 79
- = unknown concentrations. All figures in ngtl.
flow rates had to be modified, some rather drastically. Thus, the flows W-2 and W-7 were both reduced to 20 \ of their initial values and W-5 to 50 \. These changes might appear quite large. One should keep in mind, however, that the initial values are based on a regression relation developed for the whole east coast of Sweden. Therefore, there is a considerable uncertainity associated with these values. Furthermore, the initial'flow rates do represent at best a long-term average of the flow conditions in Norrsundet, whereas the calibration will describe the situation prevailing at the time when the concentration measurements were carried out.
Keeping the flow rates at the levels specified above, one could reproduce the observed chloroform concentrations with a reasonable accuracy by increasing the discharge rate relative to the value given in Table II. With EXAMS a two-fold increase was required, giving a rate of 2490 gth. With FEQUM the corresponding value was 2050 gth. Even though the validity of the models must be considered uncertain, these results do indicate that the reported discharge rate of chloroform underestimates the real load. As mentioned previously, a possible explanation is that additional chloroform is formed after the wastewater has left the plant.
The final results from the calibration are shown in Table V. The poor fit found in the compartment "R~dh~llan' can be ascribed to the fact that the concentration in "R~dh~llan· is bound to be almost the same as in the neighbouring compartment, "Klackarna". consequently, as long as the models are calibrated to give good agreement with the average concentration in "Klackarna", which is much higher than that of ·R~dh~llan", an overestimation of the concentration in ·R~dh~llan· is unavoidable.
To test the calibrated models, calculations were performed for 2,4,6-TrCp, 3,4,5-TrCG, TeCG and TeCC while leaving the flow pattern established in the calibration phase unchanged. The only changes made in the model input were the chemical-specific data and the pollutant load (see Tables I and II). The results of these calculations are summarized in Table VI.
Measured EXAMS FEQUM
TABLE V Calibration results for EXAMS and FEOUM
Milgrund Svartgrund Klunken Klackarna R~dh~llan Utposten
4.5 4.7 4.5
2.8 2.7 2.7
1.0 0.84 1.0
0.36 0.30 0.37
0.05 0.27 0.32
0.095 0.15
The flows used are: 20\: W-2,W-7. 50\: W-5. 100\: W-1,W-3,W-4,W-6. All figures in ~gtl (chloroform).
- 382-
TABLE VI Measured and calculated concentrations of four chlorinated phenols
Co 2, 4, 6-TrCP 3, 4, 5-TrCG TeCG TeCC mp min max EXAM FEQU min max EXAM FEQU min max EXAM FEQU min max EXAM FEQU
1 150 320 161 158 500 1600 749 821 400 1300 665 650 38 200 1570 1550 2 - 71 101 98 320 420 458 510 250 340 419 405 63 160 988 967 3 43 74 37 38 120 237 156 194 45 214 152 155 -4 - 46 17 17 - 36 64 5 - - 17 17 - 15 62 6 - - 6 6 15 15 21
B 44 63 0 0 8 21 0
min = minimum measured concentration max = maximum measured concentration EXAM = conc. calculated by EXAMS . .e. = Background.
unknown.
87 4 25 68 69 -86 4 16 68 69 -28 8 8 22 23 -
0 2 12 0 0 -
in the compartment (17) in the compartment (17) ~ = conc. calculated Comp = Compartment
97 356 371 97 157 168 - 156 168 - 52 55
48 0 0
by FEQUM.
All concentrations are in ng/l. No background input to "Utposten" was used.
As can be seen from Tables IV and VI, calculated and measured concentrations agree quite well for most of the compounds. The only notable exception is TeCC, for which calculated values in the innermost compartment are almost ten times higher than those measured. The explaination could be that some of the TeCC have been oxidized before reaching the first compartment ("Milgrund"). Another explaination could be TeCC's high lipophilic character. Although not reflected in its octanol-water partition coefficient, TeCC has been shown to have particularly high affinity for suspended particulates (21). The wastewater from Norrsundet Bruk contains a large amount of fibres and other kinds of undissolved organic matter, a large fraction of which will settle on the bottom. This solid material will therefore tend to act as a sink to TeCC.
Results from the OWASI calculations. In its present form the QWASI model cannot describe flow patterns characterized by streams flowinq eithe.r way across compartment boundaries or by more general circular currents. Hence, with the flow pattern assumed for Norrsundet (Figure 3), the model could not be applied to this area as a whole. Therefore, in using QWASI the environmental system was redefined to include the innermost segment, "Milgrund", alone. A major purpose of making these calculations was to find out whether sedimentation processes as those mentioned above could be responsible for the large error in predicted TeCC concentrations obtained with EXAMS and FEQUM. In the calculations the input data for flow rates, diffusion and pollutant loads were the same as that used for FEQUM.
TABLE VII TeCC and CF concentrations calculated by OWASI.
Compound
Chloroform TeCC
Water (Milgrund) Measured Calculated
4500 119
3130 871
Sediment (Milgrund) Measured Calculated
7200 333000
- = Data not available All fiqures in nq/l
- 383-
concentrations of CF and TeCC calculated using the QWASI model are given in Table VII. The results are satisfactory as far as CF is concerned. The predicted concentrations of TeCC, however, is still too high compared to measurements. This indicates that there are other more important reasons for the overestimation of the TeCC concentrations than a high degree of sedimentation. As mentioned above, extensive oxidation of TeCC is one possible explanation.
DISCUSSION AND CONCLUSIONS
The model output also includes information on how the studied chemicals are removed from' the system. The main sinks for the different compounds and their relative importance are given in Table VIII. According to the calculations, the input of chlorophenolics to the Norrsundet area is to a large extent transported out of the receiving area. For CF evaporation is a very important process.
Disregarding TeCC all models give reasonably good predictions for the concentrations in water of the compounds included. In most cases the predicted values are within the range of the observations. To obtain a numerical measure of the goodness of fit, linear regression on observed and calculated concentrations was performed. (Table IX).
As can be seen from Table IX, the correlation coefficients are quite close to 1, the smallest one being 0.86. This, however, merely demonstrates a linear relationship between measured and calculated concentrations. If the regression equations are considered, the picture is changed. TeCC differs from the other compounds by a large slope and y-axis intercept value. This shows that the models are unable to give correct predictions for TeCC. For the other compounds the slopes are close to 1 and the y-axis intercepts close to O. The main cause of the deviation is the models inability to separate the concentrations in "Klackarna· and ·R~dh~llan·.
Considering the coarse nature ot th~ models themselves and the uncertainty in many of the model parameters and parts of the chemical data one can hardly expect better agreement than that obtained for the test compounds (exept TECC). One should also recall that the measurements themselves are somewhat uncertain.
The question if the model performance is satisfactory has no simple answer. Since many of the model parameters are quite uncertain, some deviations between observed and simulated values must be expected. In the present work
TABLE VIII Fate Qf chemi~al gS eer ~~nt of lOgd
Compound Evaporized Exported to Chemically Gulf of Bothnia degraded
EXAMS FEQUM EXAMS FEQUM EXAMS FEQUM
Chloroform 41.7 36.4 58.3 63.6 « 0.01 « 0.01 2,4,6-TrCP 0.4 0.2 97.3 99.8 2.3 3,4,5-TrCG 30.3 3.3 69.7 96.7 TeCG 5.6 1.5 93.4 97.5 TeCC 0.03 0.01 92.5 96.0 7.5 4.0
-~-
TABLE IX Rearession And Correlation Coefficients
Coefficient Cf 2,4,6-TrCP 3,4,5-TrCG TeCG TeCC EXAM FEQU EXAM FEQU EXAM FEQU EXAM FEQU EXAM FEQU
Y-axis * Interceptt(A) -0.01 0.1 22.6 22.5 63.4 85.2 71.6 72.3 -2758 -2667 Slope (B) 1.02 .96 0.60 0.58 0.70 0.74 0.75 0.73 35 34 Corr. Coeff. (R) 1.0 1.0 0.86 0.87 0.97 0.97 0.97 0.97 0.94 0.94 t
Regression equation: (Predicted cone.) = A + B . (measured cone.)
we were somewhat hampered by the fact that much of the data needed to adapt the models to the studied environment in a satisfactory way were unavailable. Information on water currents was particularly scarce. Furthermore, data on concentrations in sediments and biota are essential for a satisfactory evaluation.
There are a number of parameters in the models which we had to estimate or simply leave out. Several of these, including currents and amounts of suspended particulate and biotic matter in the receiving water, could easily have been determined through field measurements. This suggests that one should start the modelling work at an early stage in the investigation, when the measurement programme is being planned. Taking due account of the data required by the models will give a security for high quality results.
In conclusion: The agreement obtained between observed and calculated concentrations in water is in general promising. FEQUM and EXAMS give very similar results. This is perhaps to be expected since key parameters are in general the same even though the actual formulas used may differ.
The results for chloroform obtained by QWASI for Milgrund were lower than that calculated for the two other models. For TeCC QWASI's results were very similar to EXAMS and FEQUM. QWASI is a much simpler model than the two others and it has a smaller user area. In spite of this, and because of its simplicity, the model may be a supplement to the other models.
Our experience shows that by closely coordinating modelling work and field measurements the investigations of water pollution by industrial wastes can be greatly improved. The modelling helps to identify key parameters to' be measured and the model may be used to estimate concentrations in areas and under conditions where measurements have not been made. Most important, the model may be used to predict concentrations for various emission scenarios.
ACKNOWLEDGEMENT
We are grateful to Lars Renberg of the Swedish Environmental Protection Board for drawing our attention the Norrsundet case and for his support during the work.
REFERENCES
(1) Ambrose,R.B., Vandergrift,p.E.S.Bh Wool,T.A, (1986) WASP3, A Hydrodynamic And Water Quality Model. u.S.Environmental Research Laboratory Report EPA/600/3-86/034.
- 385-
(2) Rosting, H.I., Spalding D.B., (1986). PHOENICS 84 and Beyond.
g~~i~tg;~ i*~~k~ft;nc~~s;I~R~ iK§~o;~ds~;t~4=~S&eff:g~f!§BRr3:~ig:lG3~1~8. (3) Burns, L.A' I Cline D.M. and Lassiter R.R. (1982). Exposure Analysis Modeling System ,EXAMS): User Manual and System Documentation. ~ Environmental Research Laboratory. Report: EPA-600/3-82-023.
(4) Mackay~ D. (1979). Finding fugacity feasible. Environ. Sci. Technol., .11, 1218-12.l3.
(5) MackaYi D. and Paterson S. (1981). Calculating fugacity. Environ. Sc~. Techno ., li, 1007-1014.
(6) Mackay~ D. and Paterson S. (1982). Fugacity Revisited. Environ. Sci. Technol., lA, 654A-660A
(7) Mackay, D., Paterson S. and Cheung B. (1985a). Evaluating the alf environmental fate of chemic The fugac~ty level III approach as applied to 2,3,7,8-TCDD. Chemosphere, ii, 859-863.
(8) Mackay, D., Paterson S. Cheung B. and Neely W.B. (1985b). Evaluating the environmental behavior o~ chemicals with a level III fugac~ty model. Chemosphere, ii, 335-374.
(9) Mackay, D., Joy, M. and Paterson, S. (1983a) A quantitative water, air, sediment interaction (QWASI) fagacity model for describing the fate of chemicals in lakes. Chemospfiere ~ No 7/8 981-997.
(10) Mackay, D., Paterson, S., and Joy, M. (1983b) A quantitative water, air, sediment interaction (QWASI) fagacity model for describing the fate of chemicals in rivers. Chemosphere ~ No 9/10 1193-1208.
{11) Liss P.S., (1974). Processes of Gas Exchange across an air-water ~nterface. Deap-sea res. 20, 221-238.
(12) Mabey, W.R., Smith J.H./ Podoll R.T., Johnson H.L., Mill T., Chov T.W., Gates J., Waight Partr~dge I., Jaber H. and Vandenberg D. (1982). Aquatic fate process data for organ~c priority pollutants. EPA report No. 440/4-81-014.
(13) Abrahamsson, K. and Xie T.M. (1983). Direct determination of trace amounts of chlorophenols in fresh water, waste water and sea water. J. Chromatoqr., 279, 1.99-208. -
(14) Bidleman, T.F. and Renberg L. (1985). Determination of vapor pressures for chloroguaiacols, Chloroveratroles, and Nonylphenol by Gas Chromatography. Chemosphere., ii, 1475-1481.
(15) Lyman, W.J., Reehl W.F. and Rosenblatt D.H. (1982). Handbook of Chemical Property estimation methods. McGraw-Hill. ISBN 0-07-039175-0.
(16) Xie, T.M. and Dyrssen D. (1984). Simultaneous Determination of Partition Coefficients and Acid~ty Constants of Chlorinated Phenols and Guaiacols by Gas Chromatography. ~nal. Chim. Acta., 160, 21-30.
(17) Xie T.M. Abrahamsson K., Fogelqvist E. and Josefsson B. (1986). Distribution o~ Chlorophenolics in a Marine Environment. Environ. Sci. Technol., 20, 457-463.
(18) Falkenmark, M. (1986). Hydrology of the Baltic Sea Area: Temporal Fluctuations in Water Balance. AMBIO, li, 97-102.
(19) Jonsson, P., Jonsson B., HAkanson L. and Martinsen K. (1986). Spr~dning av klorerat organiskt material frAn skogsindustrier. Report 3228 from Statens NaturvArdsverk, Sweden. ISBN 0282-7298.
(21) Martinsen, K., Carlberg G.E. and Kringstad A. (1987). Methods for determination and character~zation of chlorinated organic substances from paper and pulp mills.in water,' sediment and fish. Paper presented at the Second IAWPRC Sympos~um on Forest Industry Wastewaters, ~n Tampere, Finland, 1987.
- 386-
MODELLING OF GROUNDWATER TRANSPORT OF MICROORGANIC POLLUTANTS : STATE-OF-THE-ART
Summary
F. DE SMEDT Laboratory of Hydrology
Vrije Universiteit Brussel
A review is presented of some basic concepts and recent achievement concerning the modelling of transport of organic substances in groundwater. Organic pollutants are subjected to complexe physical, chemical and' biological transf orma tions while moving through the groundwater layers, such that no individual possesses a complete understanding of all the processes involved, and consequently the study of this subject becomes an interdisciplinary field of research. Recently, a large number of new developments have emerged giving new insight into processes as hydrodynamic dispersion, sorption, biodegradation and immiscible flow. It is concluded that considerable work has been done with respect to modelling of organic pollutant transport in groundwater, but a need to further developments remains necessary in order to put an effective stop to the very serious deterioration of the groundwater reserves.
1. INTRODUCTION The pollution of our natural water resources has increased
tremendously, especially since World War II. At first, it was believed that the groundwater was fairly protected and would remain a water source of good quality for always. Nowadays, it is recognised that already many groundwater layers are polluted and that the continued widespread use of chemical products and disposal of large volumes of waste materials, form a serious threat to all existing groundwater reserves. New instances of groundwater contamination are continually being recognized. Every day hazardous organic compounds, e.g. pesticides, herbicides and solvents are intentionally disposed, accidentally spilled, or applied for agricultural reasons, resulting in a general increasing low-level organic contamination of the groundwater. More than 100 organic contaminants where found until 1981 in groundwaters in The Netherlands (1) : a list of these together with the highest detected concentrations is presented in Table I. This list is probably indicative for all industrialised countries (2), but when updated will likely grow as many thousands of organic chemicals are presently manufactured in industry.
The reliable assessment of groundwater contamination is possible if the behavior and transport of the contaminants in the flowing groundwater can be predicted. Quantitative predictions of the distribution of micro-organic pollutants in the groundwater layers can be made if the processes controlling transport, hydrodynamic dispersion,
- 387-
Table I. Groundwater contaminants and max. concentrations detected in the Netherlands until 1981, taken from (1).
HYDROCARBONS 62. Dimedon 63. 1,1-Dibutoxyisobutane
I. Cyclohexane 30 64. Dibutylether
2. C I-Cyclohexane I 65. Dipen ty lether
3. C2-Cyclohexane 66. Diphenylether
4. C3-Cyclohexane 67. Butylacetate
5. C4-Cyclohexane 0.1 6B. Butylbenzoate 6. C6-Alkanes 10 69. Diethy lphthala te 7. C7-Alkanes 40 70. Tetrahydrofuran B. CB-Alkanes 10 71. Oibenzofuran
9. Cg-Alkanes 10 72. Trimethyl-penta-diol-10. CIO-Alkanes 10 di-isobutyrate II. Benzene .100 12. Toluene 300 COMPOUNDS CONTAINING HALOGEN 13. Xylenes 1000 14. Ethylbenzene 300 73. Dichloromethane 15. C 3-benzenes 300 74. Bromochloromethane 16. C4-benzenes 30 75. Trichloromethane 17. CS-benzenes 10 76. Bromodichloromethane lB. C6-benzenes 0.3 77. Dibromochloromethane 19. Styrene 10 78. Tr ibromomethane 20. Limonene 10 79. Tetrachloromethane
21. Indane 10 80. I,I-Dichloroethylene 22. MethYl1ndane 10 81. I , I-Diehl oroethane 23. Dimethylindane I 82. 1,2-Dichloroethane 24. Naphthalene 30 83. 1, 1, 1-Trichloroethane
25. Methylnaphthalene 10 84. Tr ichloroethy lene
26. Acenaphthene 85. Tr ibromoethy lene 27. Biphenyl 3 86. Tetrachloroethylene
28. Methylbiphenyl 0.3 87. Bromocyclohexane
29. Fluorene 0.03 88. Chlorobenzene
30. Fluoranthene 10 90. Dichlorobenzenes
31. 3,4-Benzopyrene 91. Trichlorobenzenes
32. 3,4-Benzofluoranthene 92. Chlorocresols
33. 11, 12-Benzofl uoranthene 93. 2.3.6.-Trichlorephenol
34. 1,12-Benzoperylene 94. 2.4.5.-Trichlorophenol
35. Indenopyrene 95. 2.3.4. -Tetrachlorophenol 96. 2.3.5.6. -Tetrachlorophenol
COMPOUNDS CONTAINING OXYGEN 97. Pen tac hloropheno 1
36. Diethyleneglycol COMPOUNDS CONl'AINING NITROGEN
37. cyclohexanol I 38. Phenolpropanol 100 98. Phenylisocyanate
39. Methylphenol 10 99. Diphenylamine
40. C2-phenol 30 100. 2.2.4-Trimethyl-6-etoxy-
41. C3-phenol 30 quinoline
42. C4-phenol 100 43. C5-phenol 3 COIIPOUND5 CONTAINING SULPHUR
44. Phenylphenol 3 45. 2,4-Di-sec.butylphenol 101. 2-Hydroxybenzothiazol
46. Methylnaphtol 102. 2.3-Benzothiophene
47. cis-Menthol I 103. 5-Me thy 1- ,2 • 3 -benzothio-
48. Dimethylbenzaldehyde I phene
49. Cinnamaldehyde 0.3 104. 4-Thiaheptane
50. Cyclohexanon 30 105. Methylpropyldisulphide
51. Di-isobutylketon 0.3 lOG. Methylisobutyldisulphide
52. Octanon-3 0.3 107. Ethy li sobuty ldi sulphide
53. Acetophenon 10 108. Dipropyldlsulphlde
54. I-Indanon 3 109. Butylpropyldisulphlde
55. Conmarin 3 56. Trimethylcyclohexanon 30
MISCELLANEOUS COIIPOUNIlS
57. Isophoron 10 58. camphor 1000 110. Tri-ethylphosphate
59. Fenchone 100 Ill. Tri-butylphosphate
60. Benzophenon 0.03 112. Tri-isobutylphosphate
61- 2,6-Di -tert.Butylbenzo- In. N-n-butylsulfonaaide
quinon 0.3
- 388-
10
0.3 3 O. J
1000
3000 8
10 0.3 0.3 4
30 10 10
3 3000 1000
100 )0 10
3 3 I
10 I 2
10 3
0.3
30 100
30 I 0.1 0.1 0.1 I 0.1
0.1 0.3
10 3
and chemical, physical, and biological reactions are well understood. During the last years, significant progress has been made in the modelling of organic pollutant transport (3,4,5). This paper gives a brief review of some basic concepts and recent achievements.
2. TRANSPORT OF ORGANIC COMPONENTS
2.1. General Two types of flow are possible when organic contaminants enter the
groundwater. Miscible displacement occurs when the contaminant is completely soluble in the groundwater. Figure 1A shows a hypothetical example. The pollutant dissolves gradually in the infiltrating water from rainfall and migrates downwards to the watertable where it enters the saturated groundwater layers. Here, the pollutant flows together with the groundwater predominantly in horizontal direction. The physical transport processes involved are advection and hydrodynamic dispersion. In case the organic pollutant not immediately dis sol ves in the water immiscible displacement results, as shown in fig. 1B and 1C. The pollutant moves downwards under the influence of gravity as a seperate phase, occupying part of the pore space. When reaching the zone of saturation the behavior becomes determined by its density. When the density is less than 1, the pollutant will float on the water table while gradually spreading outwards (Fig. 1B). A contaminant that is more dense than water continues to move downwards through the saturated zone until it encounters an impervious stratum, whereupon it will migrate downdip along the slope of the contact plane (Fig. 1B). Although immiscible, eventually small amounts might mixe with the groundwater and be transported in miscible form. At the same time of the transport, either in miscible or in immiscible form, the pollutant is subjected to other chemical, physical and biological processes. In case of organic pollutants the most important of these are sorption and biodegradation. We will now focus on the different processes separately.
2.2. Advection and hydrodynamic dispersion In miscible displacement the pollutent is transported by advection
and dispersion. Advection refers to the transport together with the macroscopic groundwater flow. Hence, the first step in the prediction of pollutant transfer is to solve the groundwater flow equation for the flow velocity distribution in space and time. In simple cases, analytical solutions are available but generally numerical solution techniques have to be used. The theory of groundwater modelling is well established. Good references are available (6,7,8).
Generally, groundwater movement is described by means of a macroscopic approach; i.e. physical properties are averaged over space such that values of the hydraulic potential, the flow velocity, etc. can be attributed to any point of the porous system. Hydrodynamic dispersion is a mixing process resulting from the microscopic flow variations, as shown schematically in Fig. 2. Dispersion results in a spreading of the contaminant as it flows through the subsurface, which cannot be predicted solely from the macroscopic groundwater flow velocities. The equation describing advective-dispersive flow is written as (6) :
d(eC)/at = V(DeVC) - V(veC) [ 1)
- 389-
,I!:. ... oundw ... tcr ~77?"'77>""'7?",""
~roundw .. t c r fJo w
7?//,T/,T,Q?777777771777/
8. I •• i~cible flaw -li~htcr th~n w~tcr
77777777777777777777.
c. I •• i sc i ble rlo~ -hC~T i c~ th~G wa ~er
stratu .
Fig. 1. Different types of groundwater pollution
_croscopic nov Teloci ty
Fig. 2. Advection and hydrodynamic dispersion
- 390-
llicroscopic flov ~ths
where C contaminant concentration e water content of the porous formation t time V del operator v macroscopic groundwater velocity D dispersion tensor
The classical theory of dispersion is well established (6,9). The basic assumption is that the dispersion is a second-rank tensor, with elements depending upon the coefficient of molecular diffusion and the components of the macroscopic flow velocity, in the following way
where D o
£L' £T °ij
D .. = (£L-£T) v.v./lvl + 6·. (£T Ivl + D ) 1J 1 J 1J 0
diffusion coefficient longitudinal and. lateral dispersivities Kronecker symbol
[2]
The dispersivities are defined as constant properties of the porous medium characterizing the degree of microscopic flow deviations. Recent studies however, indicate that these parameters depend upon the distance scale of the pollutant transport problem. Fig. 3 shows values of the longitudinal dispersivities versus the field scale, taken from (8). This phenomenon might be explained by the heterogeneity of the natural ground water layers, i. e. the farther a pollutant travels the more it will encounter heterogeneities and the more the movement will deviate from the average flow velocities. A whole new branch of research has subsequently come into existence, where the transport of water and solutes is analyses by means of stochastic processes. Recent reviews of this interesting field of research are available (4,10,11). Nevertheless, when appropriate values are chosen for the dispersivities (as given in Fig. 3), the classical theory represented by equations [1] and [2] can provide for resonable estimates of the advective and dispersive movement of pollutants. This approach also applies to the unsaturated zone of the groundwater system (12), although complications might arise from the presence of immobile water phases (13,14).
As an illustration of the advective-dispersive transport process, let I s consider a simple example, where a contaminant is injected .as a instantaneous point source in an uniform flow system. The concentration distribution, neglecting molecular diffusion is given by (9) :
where M t x y,z
222 __ -::,M'-T::---o-...,.".-__ ex [_ (x-vt ) y + z ] 8(TIvt)3!2£L1!2 £T p 4£Lvt 4£Tvt
mass injected at x=y=z=t=O time coordinate in the flow direction coordinates perpendicular to the flow
The concentration distribution is illustrated in Fig. 4. the position of the centre of gravity of the pollution cloud at time twill lie along the flow path at a distance corresponding to advective movement, i.e. vt. The concentration distributions along the x- and y-axes illustrate the dispersive movement, i.e. the spreading of the pollution cloud around its centre of gravity.
- 391-
YD.,ror. rlow
•
•
~'
.~........... .,
.000000ft SucIcIry.1I'1!I "~82 .'~: 1IIl
• a.;",... arz -~:-
• NIIICN. 9IZ
d~~~~~~~~~~~~~~~~~ d d d d ~ d
Area scale length [m)
c
Fig. 3. The scale effect in dispersion, after (8)
1
• t
Coaccacratcd coato.r. .at ti8c t
to ..... .. . .. .. -;
u
Fig. 4. Solution of an instantaneous point source injection in a uniform flow
- 392-
For real groundwater contamination problems, the actual flow velocity distribution will be more complicated, such that the pollutant transport equation has to be solved with numerical techniques, as will be discussed.
2.3. Sorption Contaminants migrating through the groundwater system are, due to
the dispersive mixing, continuously brought into contact with the surfaces of the solid material, such that adsorption and desorption reactions can occur. The transport equation for a sorbing pollutant can be written as :
a(Sc)/at ~ V(DCVC) - V(V 8C)- a(pS)/at [ 4]
bulk density of the solid matrix where p S amount of pollutant adsorbed per unit weight of solid matrix
Previously, it was believed that sorption could be considered as an instantaneous equilibrium reaction, described by an equilibrium isotherm, of for instance the liniar, Freundlich or Langmuir type. However, laboratory and field experiments have shown that non-equilibrium effects should be taken into account (3,15).
Recent approaches (16,17) take into account chemical and physical non-equilibrium effects, as shown schematically in Fig. 5. The sorption process is assumed to be governed by diffusion of the pollutant particles in an immobile water phase surrounding the sorption surfaces, local adsorption equilibrium between the solute adsorbed onto the sorption surfaces and the solute in the immobile water, and surface and pore diffusion as intra stationary phase mass transport. With such complex models, the amount of adsorption becomes very difficult to predict. However, sensitivity analyses of the different components of such models indicate that not all processes are relevant and that simplifications are allowed (3). Hence, for practical field predictions it is possible to describe the sorption of organic solutes by the most simple approach, i.e. a liniar reversible isotherm (5,18,19) :
[ 5]
where Kd : distribution coefficient.
For nonpolar organic compounds, this distribution coefficient is related to the amount of organic carbon present in the solid material and the partition coefficient for a mixture of water and octanol. A relation of this type is for instance given as (19) :
0.62 OC.K ow
where OC : weight percent of solid-phase organic carbon K : octanol/water partition coefficient.
Other s£~ilar equations have been published (1,5,20).
[6]
In case of linear isotherm adsorption, the transport equation can be written as
R a( ec) / at V(DSVC) - V(v SC) [7]
- 393-
80bile ~ater phase
advection and dispersion
diffusion in ~bile water
sorption and surface diffusion
pore diffusion
Fig. 5. Sorption with non-equilibrium effects, after (16)
ad't'ection and dispersion
diffusion in ~bile water
.obile water phase
Ilicl"'D(;Oloay
Fig. 6. Biodegradation, after (25)
- 394-
where R retardation factor, given by R = 1 + pK/e [8]
Equation [7] reduces to the transport equation [1] of a non-sorbing chemical, when apparent velocity and dispersion coefficients are defined by dividing the real values by factor R. For instance, the solution for the instantaneous point source injection of a sorbing pollutant in a uniform flow, is identical to equation [3] , where v is replaced by viR:
C1 (v + viR) [9]
such that the distance travelled by the centre of gravity of the pollution cloud becomes vt/R.
2.4. Biodegradation Contrary to previous belief, recent investigations have shown that
high levels of microorganism are present in the groundwater layers, even at great depth (21). The amount of biomass, almost exclusively in form of bacteria, is much larger than what normally occurs in surface waters. Biodegradation of a broad range of organic compounds has been demonstrated in laboratory and field studies (2,3,5); although mostly aerobic microbial mineralization has been shown to occur, also anaerobic mineralization is possible (21,22,23). Table II presents prospects for biotransformation of several important classes of organic pollutants in groundwa'ter, according to (21).
To date, the accurate prediction of actual biodegradation of organic contaminants in groundwater remains difficult due to the complexity of the processes involved. Recent modelling efforts are given by (24,25,26). The most advanced model (25) is depicted schematically in Fig. 6. It is assumed that microorganism grow in microcolonies attached to the solid materials; the growth and decay of the organisms is controlled by the energy source and free oxygen, modelled by modified Monod kinetics. The transport of organic solutes and oxygen is described by means of advection-dispersion in mobile water, diffusion in immobile water and adsorption in the microcolonies. Results of the mo~el calculations show that biodegradation has a major effect on pollutant transport, when proper conditions for microbial growth exist, but that anaerobic conditions develop rapidly when large amounts of contaminants are present.
Considerable additional research is required before such models can be used for accurate predictions, because the actual biodegradation processes and controlling parameters are not exactly known for real field situations. Simplified approaches are possible in two cases. At low concentrations the biodegradation can be described by a first order rate expression in terms of the pollutant concentration. The transport equation becomes :
a(eC)/at = v(DevC) - v(veC)- AeC [10 ]
where A : biodegradation rate coefficient Solutions of [10] can easily be related to solutions of a non-degrading chemical. For instance, in case of the instantaneous point source injection in an uniform flow, the solution becomes :
[ 11]
- 395-
Table II. Prospects for biotransformation , taken from (21)
CIu. al Compound.
Halogenated aliphatic hydrocarbons Trichloroethylene Tetrachloroethylene I. I. I·Trichloroethane Carbon tetrachloride Chloralorm Methykne chloride I, 2-Dichloroethane
Bramm.ted met~ OdorobeDzenes
a.Iorobenzene I. 2-Dichlorobenzene I, 4-DiChIorobenzene I, 3-Dichlorobenzene
AlkyIbenzenes Benzene Toluene Dimethylbenzenes Styrene
Phenol ...d a1yk1 phenols CbIoropbenoJs Aliphatic hydrocubons Polynuclear aromatic hydrocarbon.
Two end three rings Four or more rings
·From Wilson ...d McNabb (1983) ·Posslble but probebIy incomplete, <ProbobIe but at hip concentration,
~scible orf:anic liquid
phase
solid uteri.al
Aerobic WoI~r. Concmtration al Pollutant (..wL)
100 10
NOM Non~
Non~ Non~
None None None None NOM None Possible Improbeble Pouible Improbeble Improbeble ImpoobebJe
Probable Pouibk Probable Possible Probable Possible Improbeble Impmbeble
Probable Pouible Probable Possible Probable Pouible Probable Possible Probable Probable Probable Possible Probable Possible
Pouible Pouible ImpoobebJe Impmbeble
Fig. 7. Immiscible flow
- 396-
Anwrobic Wat ..
Possibk' Possibk' PosslbW' Possible· P .... ibk· P .... ible Poalble ProbobIe
None None None None
None None None None Prob.bIe' Possible None
None None
The transport resembles the situation depicted in Fig. 4, but with the concentration values decreased by the exponential decay factor. A second simplification is possible as described by (26) in case of large injection of organic pollutants resulting in anaerobic conditions. The consumption of oxygen and substrate by the microorganisms may be approximated as an instantaneous reaction, resulting in a contamination plume with no free oxygen present. The solution of such transport can be approximated as
C4 = max[C I - COif, 0] [12]
where CO : original oxygen concentration present in the groundwater f ratio of oxygen to substrate consumption
The transport resembles the situation shown in Fig. 4, with all concentration values below the level COif removed from the system.
2.5. Immiscible transport Immiscible transport is far the most complicated process to model.
Water and organic liquid (and possibly air in the unsaturated zone) coexist as separate phases in the porous medium, as shown in Fig. 7. The resulting multi phase transport depends upon the pressure discontinuities at the fluid-fluid interfaces and the relative phase permeabilities, both strongly varying functions of the saturation degree of the different phases in the porous medium. At the same time, other processes intervene, e.g. volatilization, dissolution, chemical reactions, adsorption and biodegradation. Earlier research on immiscible organic groundwater contamination has been compiled in (27). Quantitative modelling has only been undertaken very recently. An overview of the requirements and difficulties is presented in (28). Recent modelling results are given in (28,29,30). A model for pure immiscible flow, excluding all other processes has been presented in (29), and applied to an actual contamination case of a chemical waste disposal site, involving contaminants as mirex, lindane, trichlorophenol and dioxin. A more extensive model is presented in (28), where a hypothetical case of groundwater pollution by trichloroethylene, both in immiscible and miscible form, is analyzed. The most complicated model is developed in (30), for describing the fate of hydrocarbon constituents of petroleum products in unsaturated soil. Transport of each constituent is assumed to occur as solute in the water phase, vapor in the air phase, and as an unaltered constituent in the oil phase. The model allows for adsorption and biodegradation, incorporating equations for the free oxygen consumption and transport. In a subsequent paper (31) simulation results are presented for a hypothetical pollution case of gasoline, consisting of eight groups of hydrocarbon constituents. Rates are calculated at which the contaminants are removed from the soil, by either entering the atmosphere or groundwater, or by being biodegraded. It is clear that for such models a number of problems remain unresolved, such as the exact understanding and description of the occuring processes and the exact assessment of the parameters involved. Nevertheless the progress'made in this field is substantial.
3. NUMERICAL MODELLING It is clear from the discussion above, that accurate descriptions
of the processes controling organic pollutant transport in groundwater, mostly result in complicated mathematical expressions, which when
- 397-
applied to real field conditions are very difficult to solve. Therefore, a large amount of research has been devoted to numerical solution techniques on either mainframe or .microcomputers. Accurate numerical solutions for miscible or immiscible pollutant transport in combination with groundwater flow, can be obtained by using techniques as finite difference, f~nite element, and related methods. The body of literature devoted to these topics is too extensive, such that no general overview can be presented here. Basic introductions to the subject can be found in (6,7,9). Recent states of the art have been given by (32,33). The new trends towards using microcomputers in modelling applications are discussed in (8,34).
4. CONCWSION Organic contaminant transport modelling in groundwater has rapidly
progressed in recent years. A large number of new developllents have emerged to such an extend that mathematical lIodelling can provide for confident predictions, although not all pheno~na are co.pletely understood and not all para.eters are fully known. Further research is necessary to clear up the remaining barriers.
Due to the co.plexity and variety of the different processes, modelling of organic contamination of the subsurface has beco.e an interdisciplinary field, such that no individual processes a co.plete understanding of all the phenomena involved. Researchers from a wide variety of disciplines will have to cooperate in the future. It is anticipated that in the coming year such progrss can b.:: ude that an effective protection and restoration of the groundwater becomes possible.
REFERENCES
(1) ZOETEHAN, B.C.J., DE GREEF, E. and BRINKMAN, F.J.J. (1981). Persistence of organic contaminants, lessons from soil pollution incidents in The Netherlands. Studies in Environmental Science, Volume 17. (Elsevier) : 465-480.
(2) N.N. (1986). Behaviour and transformation of organic pollutants in groundwater treatment. Water Pollution Research Report 3, C.E.C., EUR 11094 : 172 pp.
(3) ABRIOLA, L.M. (1987). Modeling contaminant transport in the subsurface : An interdisciplinary challenge. U.S. National Report to International Union of Geodesy and Geophysics 1983-1986 125-134.
(4) ANDERSON, W.P. (1984). Movement of contaminants in groundwater : Groundwater transport - advection and dispersion. Studies in Geophysics, Groundwater Contamination, National Ac. Pres's, Washington D.C. : 37-45.
(5) CHERRY, J.A., GIUfAM, R.W., and BARKER, J.F. (1984). Contaminants in groundwater Chemical processes. Studies in Geophysics, Groundwater Contamination, National Ac. Press, Washington D.C. : 46-64.
(6) BEAR, J. (1979). Hydraulics of groundwater. McGraw-Hill Int. Book Co. : 576 pp.
(7) WANG, H.F., and ANDERSON, M.P. (1982). Introduction to groundwater modeling - Finite difference and finite element methods, V.H. Freeman and Co. : 237 pp.
(8) KINZELBACH, V. (1986). Groundwater modelling - An introduction with
- 398-
sample programs in BASIC. Developments in Water Science 25. Elsevier : 333 pp.
(9) FREEZE, R.A., and CHERRY, J.A. (1979). Groundwater. Prentice-Hall Inc. : 604 pp.
(10) DAGAN, G. (1986). Statistical theory of groundwater flow and transport Pore to laboratory, laboratory to formation, and formation to regional scale. Water Resources Research, 22-9 120S-134S.
(11) GELHAR, L.W. (1986). Stochastic subsurface hydrology from theory to applications. Water Resources Research, 22-9 : 135S-145S.
(12) NIELSEN, D.R., VAN GENUCHTEN, M.TH., and BIGGAR, J.W. (1986). Water flow and solute transport processes in the unsaturated zone. Water Resources Research, 22-9 : 89S-108S.
(13) VAN GENUCHTEN, M.TH., and WIERENGA, P.J. (1976). Mass transfer studies in sorbing porous media, I : Analytical solutions - Soil Science Society of American Journal, 40-4 : 473-480.
(14) DE SMEDT, F. and WIERENGA, P.J. (1984). Solute transfer through columns of glass leads. Water Resources Research, 20-2 : 225-232.
(15) VAN GENUCHTEN, M. TH., WIERENGA, P.J., and O'CONNOR, G.A. (1977). Mass transfer studies in sorbing porous media. III : Experimental evaluation with 2,4,5-T. Soil Sci. Soc. Am. J., 41 : 278-285.
(16) CRITTENDEN, J.C., HUTZLER, N.J., GEYER, D.G., ORAVITZ, J.L., and FRIEDMAN, G. (1986). Transport of organic compounds with saturated groundwater flow : Model development and parameter sensi ti vi ty. Water Resources Research, 22-3 : 271-284.
(17) MILLER, C.T., and WEBER, W.J. (1986). Sorption of hydrophobic organic pollutants in saturated soil systems. Journal of Contaminant Hydrology, 1 : 243-261.
(18) CHIOU, C.T., PETER, L.J., and FREED, V.H. (1979). A physical concept of soil-water equilibria for nonionic organic compounds. Science 206 : 831-832.
(19) KARICKHOFF, S. W., BROWN, D. S., and SCOTT, T. A. (1979). Sorption of hydrophobic pollutants on natural sediments. Water Resources, 13 : 241-248.
(20) SCHWARZENBACH, R.P. (1985). Sorption behavior of neutral and ionizable hydrophobic organic compounds. Proceedings of the Fourth European Symposium on "Organic micropollutants in the aquatic environment", Vienna, Austria, October 22-24, 1985, D. Reidel Publ. Co. : 168-177.
(21) WILSON, J.T., and McNABB, J.F. (1983). Biological transformation of organic pollutants in groundwater. EOS : Trans. Am. Geophys. Union, 64 : 505.
(22) ZEYER, J., KUHN; E.P., EICHER, P., and SCHWARZENBACH, R.P. (1986). Behaviour and transformation of organic pollutants in groundwater treatment. Water Pollution Research Report 3, CEC, EUR 11094 : 117-119.
(23) SCHINK, B. (1985). New reactions involved in non-aerobic degradation of aromatic compounds. Proceedings of the Fourth European Symposium "Organic micropollutants in the aquatic environment", Vienna, Austria, October 22-24, 1985, D. Reidel Publ. Co. : 321-330.
(24) CORAPCIOGLU, M.Y., and HARIDAS, A. (1984). Transport and fate of microorganisms in porous media A theoretical investigation. Journal of Hydrology, 72 : 149-169.
(25) MOLZ, F.J., WIDDOWSON, M.A., and BENEFIELD, L.D. (1986). Simulation
- 399-
of microbial growth dynamics coupled to nutrient and oxygen transport in porous media. Water Resources Research, 22-8 1207-1216.
(26) BORDEN, R.C., and BEDIENT, P.B. (1986). Transport of dissolved hydrocarbons influenced by oxygen-limited biodegradation. 1 Theoretical development. Water Resources Research, 22-13 1973-1982.
(27) SCHWILLE, F. (1984). Migration of organic fluids immiscible with water in the unsaturated zone. Pollutants in Porous Media, The Unsaturated Zone Between Soil Surface and Groundwater, Springer-Verlag : 27-48.
(28) PINDER, G.P., and ABRIOLA, L.M. (1986). On the simulation of nonaqueous phase organic compounds in the subsurface. Water Resources Research, 22-9 : 1095-1195.
(29) OSBORNE, M., and SYKES, J. (1986). Numerical modeling of immiscible organic transport at the Hyde Park landfill. Water Resources Research, 22-1 : 25-33.
(30) CORAPCIOGLU, M.Y., and BAEHR, A.L. (1987). A compositional multiphase model for groundwater contamination by petroleum products. 1 : Theoretical considerations. Water Resources Research, 23-1 : 191-200.
(31) BAEHR~ A.L., and CORAPCIOGLU, M.Y. (1987). A compositional model for groundwater contamination by petroleum products. 2 : Numerical solutions. Water Resources Research, 23-1 : 201-213.
(32) NARASIMHAM, T . N. ( 1984) . Formula tiot:l of numerical equations. Fundamentals of Transport Phenomena in Porous Media. Martinus Nyhoff Publ. : 773-804.
(33) LEWIS, R.W., and ROBERTS, P.M. (1984). The finite element method in porous media flow. Fundamentals of Transport Phenomena in Porous Media, Martinus Nyhoff Publ. : 805-898.
(34) BEAR, J., and VERRUIJT, A. (1987). Modeling groundwater flow and pollution. D. Reidel Publ. Co. : 414 pp.
- 400-
POSTER SESSION V - MODELLING
Modelling of surfactants in the Comunidad de Madrid as subbasin of Tagus River
Modelling of anthropogenic substances in aquatic systems MASAS - A personel computer approach
MODELING OF SURFACTANTS IN THE COMUNIDAD DE MADRID AS SUBBASIN OF TAGUS RIVER
F. CUBILLO Models and Technology Division
Comunidad de Madrid
Summary
The need to have an accuracy knowledge about concentrations exposure and fate of surfactants in the Comunidad de Madrid rivers determined the construction of a water quality model that was able to simulate the behaviour of some of them.
This model has been used in two main ways,to get a better knowledge and to select the investment in waste water treatment plant that would fulfill the water quality goals planned in the surface waters
In adition to the normal constituents in water quality studies has been model iced an anionic surfactant, linear alkylbenzene sulfonate (LAS). this paper show a general overview to this model.
The model construction has been based on U.S. Environmental Protection Agency computer program QUAL2E,being very helpful the cooperation, advising and cospo.(Detergents Research Association Barcelona), and the Comite Conjunto Hispano Norteamericano para la Cooperacion Cientifica y Tecnologica.
h INTRODUCTION
One of the most important problems is the appearance of foam in the little river go across Toledo city mainly beautiful historical place
of Tagus river Agency dams existing wen the because Toledo is a
It is normal to look for the origin of this phenomena where the biggest wastes are produced .This is the case of Madrid region where are released about 80 %'of basin total liquid wastes. (figure 1)
- 402-
..... - COMUN I DAO 0 [" IIIAOIt ID TUilUTOlty
........ -. TAGUS BASI N
- LIQUID "ASTn
~ "'AtN W" STI1S
SCOPE OF THE MODEL
FIGURE 1
In order to stablish the wastewater treatment plants optimal investment,to get the fixed objectives in surface water quality,to avoid this undesirable phenomena, has been developed a mathematic model based on U.S. Environmental Protection Agency computer program QUAL2E.
Tho model can simulate the surfactant concentrations along the rivers network under diferent conditions being able to make predictions and comparations between the results and the desirables levels.
A previous activity to the construction of a model is the compilation of all the basic information that ·has influence on the phenomena to modelice .
In this sense was collected information about next issues: .
-sources of surfactants wastes. -point loads from urban zones and industries. -basins and rivers bed geomorphology. -hydrology,climatology and hydraulic characteristics.
- 403-
-withdrawals. -reservoirs and canals operation rules. -canals and irrigation zones. -existing historical data of L.A.S. concentrations.
~ COMPUTER PROGRAM
The program steady-state and QUAL2E program.
used is deterministic, onedimensional,based in the
simulation, well known
It can simulate till thirteen river water quality constituents: Disolved Oxigen, Biochemical Oxigen Demand, Temperature, Algae as Chlorofil A, Nitrogen as Ammonia, Nitrites and Nitrates,Disolved Phosporus,organic Nitrogen and Phosporus, Coliforms, one nonconservative and three conservative arbitrary constituents.
One modification has been included in the program to allow simulate the alquilbenzene sulfonate surfactant, considering two kinetics: one first order biodegradation rate and one adsorption to sediment rate. (figure 2).(1).
The program has been developped in a 2655 PRIME computer property of Comunidad de Madrid.
The model is suitable for detritic rivers with perfect blend and assumes that changes on transport and dispersion mechanisms are significant only in the flow direction. It allows for multiple point loads and can simulate natural incremental inflows through point or nonpoint discharges.
~ MODEL CONSTRUCTION
The model construction was in two groups of calibration monitoring surveys. (figure 3).
based on historical data and data obtained in intensive
This calibration surveys were carry-out under diferenf conditions: one group of data were collected during winter (high flows, no irrigation,low temperatures,all the people in their residence places), the other were collected in august (lower flows, an important percentage of people on hollidays out of Madrid and the irrigation zones like a kind of big filters).
The laboratory quantitative determinations of LAS in aquous samples were realized by A.I.D. (Barcelona) using the standard methylene blue method (MBAS) and a new procedure for solid phase extraction followed by high performance liquid chromatography (HPLC).
-~-
DRG-N
(j3 0 I S
Gi, IF) S 0 I L V E
t.l , 0
0 NO z
X ORG- P Y G E 134 N
cl 6 13 z
NO, 015- P d , (I - F)
0/3,u ci 4 p
d,)1 d.2)J.
cl,P CHL~
d. 2 P ALGAE
CONSTITUENTS INTERATIONS CONSIDERED IN THE MODEL
FIGURE 2
- 405-
E3 WINTER
SUMMER
CONDITIONS
CONDITIONS
SURFACTANTS CONCENTRATIONS DURING CALIBRATI ON SURVEYS
FIGURE 3
With these data was calibrated the excellent agreement between predicted
model and
obtaining mesured
concentrations. (figure 4).
Once the model was developped, it was used to make diferent predictions, comparing the influence of diferent type of wastewater treatment proccess in diferent places with the final concentrations of LAS in relation with the necesary levels to avoid the appearance of foam in undesirable places (2) •
Next months viII continua the study surfactants in Tagus basin analyzing the influence methods to diminishing surfactant concentrations rivers network.
- 406-
of anionic of another
along the
RIO TAJO RIO JARAMA
A, .. 'I'1ivl .... ...
• ~ c .. rr .... • IOT~"
• WINTER DATA
• SUI'IH£R DATA
L.A.S. CONCENTRATIONS ON JARAMA AND TAGUS RIVERS FIGURE 4
REFERENCES
(1) BROWN, L. and BARNWELL, T.,The enhanced stream water quality models QUAL2E and QUAL2E-UNCAS: Documentation and user manual. Environmental Research Laboratory. U.S. Environmental Protection Agency.Athens GA.
(2) CUBILLO, F. situacion actual en los rios de la Comunidad General de Recursos Hidraulicos.
- 407-
de la calidad de las aguas de Madrid (1986).Direccion
Comunidad de Madrid.
MODELLING OF ANTHROPOGENIC SUBSTANCES IN AQUATIC SYSTEMS:
MASAS - A PERSONAL COMPUTER APPROACH
R.P. SCHWARZENBACH, J. WETZEL", J. HELDSTAB", and D.M. IMBODEN Swiss Federal Institute for Water Resources and
Summary
Water Pollution Control (EAWAG), 8600 DObendorf, Switzerland
" Present Address: INFRAS, 8002 ZOrich, Switzerland
Mathematical models for evaluating the dynamic behaviour of anthropogenic substances in aquatic systems (MASAS) are being developed and implemented on a personal computer. The MAS AS system will allow the user to construct models of increasing complexity for lakes, rivers, and groundwater aquifers, and to build up compound and system libraries. In this paper, the general concept of the MASAS system is described and a summary of the present status of implementation is given.
1. INTRODUCTION: THE ROLE OF MODELS IN SCIENCE
Anthropogenic substances, by definition, are compounds whose presence in nature is due uniquely, or at least largely, to human activities. As a consequence of their use, disposal (or destruction), and accidents, anthropogenic substances and their transformation products sooner or later find their way into the environment, where their concentrations are controlled by transport and transformation processes. Depending on the chemical properties of the various substances, they may accumulate in the ecosystem or in some parts of it and eventually lead to ecotoxicological effects.
Today, our capability to understand and interpret the distribution of cnemical species in the natural environment is still a step behind our capability to measure them. In view of the increasing set of field information on a growing number of synthetic chemicals, it has become a major task of environmental scientists to structure field data, to look for common phenomena, and to develop concepts, e.g. mathematical models, for the description of the fate of anthropogenic substances in the environment.
Though in the literature, the use of mathematical models for predictive purposes has gained special attention, it is by no means the only, and perhaps not even the most important goal of modelling. Scientists have since ever used models (or conceptual ideas which in the present terminology would be called "models") to arrange information derived from measurements in order to formulate general hypotheses. This was (and is) done because measurements alone yield information on .QD..e. specific situation only. It is the understanding which provides the base for generalizing the knowledge acquired for one situation to other systems and other times. Models are tools for surmounting the barrier between measuring and understanding.
As pointed out by the philosopher Karl Popper (1), there is no logic way to prove a general law based on a limited number of observations. Popper says: There is no principle of induction. This also means that there is no way to prove the correctness of a model in a strict mathematical sense. The best we can dE) is to derive hypotheses (models) from our experience and to check them over and over again against reality (observations). With every confirmation the theory (or model) becomes more established, more probable, more correct. Yet, one single falsification of the theory can destroy the
- 408-
hypothesis. This method of deduction relies on the formulation of models. In Popper's philosophic framework, the establishment of theories and their subsequent test against reality are the essence of science. He concludes that only those theories can be classified as empirical-scientific which, in principle, could be falsified by new observations. A theory which cannot be checked against measurements is not part of science.
We should realize that in environmental sciences we are confronted with a set of difficulties which, on one hand, limit the use of Poppers principle and, on the other hand, open a new role for mathematical modelling. Two points shall be mentioned:
(1) Environmental systems are often not controllable. There exist phenomena which only occur in natural systems and cannot be studied in an artificial, small-scale "microcosmos". For instance, the inpact of the rising C02-concentration on climate belongs to this category of problems. In natural systems where experiments cannot be conducted, different processes are simultaneously active making it difficult to establish simple cause/effect relationships. Models can serve as substitutes for experiments. They yield information on possible reactions of the system to changes, on the sensitivity of the system to various processes and parameters, and on possible competing processes.
(2) Environmental sciences often deal with problems which are at the borderline of our present knowlwedge and analytical possibilities. It is for instance not possible to identify by direct measurement, the various chemical forms in which a chemical pollutant (such as a heavy metal) may be present in a given system. In these cases, model calculations based on knowledge of thermodynamic properties can replace the direct analytical measurement. Thus, the distribution and transformation of a chemical pollutant can be assessed, the important parameters identified and possible ranges of concentrations estimated. Obviously, such models do not reach Popper's deduction criterion as long as the models cannot be directly checked against measurements. At this point, the urgent need for decisions in environmental pOlicies requires that science has to step beyond the safe ground of certainty. We could call this the merging point of natural and social sciences.
2. THE GOALS OF THE "MASAS" pROJECT
The goal of the MASAS project (Modelling of Anthropogenic Substances in Aquatic Systems) is to develop an "hierarchically" ordered class of mathematical models and their implemention on a personal computer. The models are aimed to describing the dynamics of anthropogenic substances in different aquatic environments including rivers, lakes, and groundwater. Here we report on the first phase of the implementation of MASAS which deals with lakes, only. The same general ideas will also be applied to the other aquatic compartments.
There exist already different models whose goals coincide with the above formulated goals. Probably the most complete model of this kind is the program EXAMS developed by the US Environmental Protection Agency (EPA) (2). The main disadvantage of this very sophisticated model lies in its complexity, making it impossible to adapt it to the often insufficient level of knowledge about organic compound characteristics and, particularly, about environmental systems.
The essentially new idea of MASAS is the possibility of the user to construct models of successively increasing complexity with respect to both, system and compound description. Starting with a one-box model, the user is didactically led to more complex models, whereby the program must inform him or her at each step to what extent a more complex model is justified or necessary and what information about compound and environment is needed to make this step. This step-by-step procedure allows the user to
- 409-
weigh the results with respect to the relevance of the different processes and differentiations successively introduced into the calculation. In addition, the sensitivity of the results can be estimated with respect to assumptions that are based on insufficient data and knowledge.
Let us illustrate the "hierarchical" structure of MASAS with the simple example of a radioisotope in a lake. The first step, the one-box model (Fig.1 a),
a c
b d Exchange with Sediments
Figure I: The hierarchical model structure of MAS AS exemplified by the fate of a radioactiv isotope in a lake: (a) One-box approach; (b) Two-box model in case that the half-live of the isotope is not much larger than the duration of the stratification; (c) Additional complexity that has to be introduced if compound exhibits both a dissolved and particulate form: If a more refined vertical structure is required, a multi-box may be used; (d) Three-dimensional mixing for fast reacting (decaying) isotope.
requires knowledge of input rate and half-life on the compound level, lake volume and water renewal rate on the system level. If the half-life of the isotope is of the same order (or smaller) than the duration of the annual vertical stratification of the lake, after performance of the one-box calculation the program gives a message to the user telling him that the one-box approach may not be adequate for his case. The program will also list the additional information needed for the next hierarchical level, the two-box model (Fig.1b): Duration of stratification, depth of thermocline and corresponding volumes of subsystems, vertical mixing between boxes during stratification, fraction of isotope input into either box. In case that these data are not available the user can still move on to the two-box model by making estimates for the missing information and check out the sensitivity of the result to parameter variations. A comparison of the two-box and one-box results serves to evaluate whether the extra model complexity can be justified.
As a next step, the program may point out additional complexity on the compound side (e.g. division of the total isotope concentration into solid and dissolved fraction, see Fig.1 c). This may also require additional data on the system level (e.g. concentration and settling velocity of particles). For highly reactive species, mixing along all three space axis, horizontal and vertical, may be necessary (Fig.ld) which obviously opens a whole new set of information needed on the system side. In summary, the MASAS concept shall lead the applier from simple to complex models by telling him the gain (better performance of
- 410-
model?) and the prize (more data required), and by giving him information on possible discrepancies between model assumptions (e.g. completely mixed system) and probable compound behaviour (e.g. not completely mixed due to high reactivity).
3. GENERAL QESCRIPTION OF THE "MASAS" SYSTEM
As already mentioned, the program MASAS is developed as a userfriendly, practical tool for evaluating the dynamics of anthropogenic substances in aquatic systems. In order to provide maximum flexibility, the whole system is divided into three mayor program modules, namely "COMPOUNDS", "SYSTEMS", and "MODELS". Each of these modules plays a different role in the modelling process, the modules "COMPOUNDS" and "SYSTEMS" may also be used as independent units. The module "COMPOUNPS" fulfills three major tasks:
(I) It serves to assemble the record of compound properties and compound specific reaction data that are needed for describing the processes the compound is subjected to in the selected environmental system.
(2) It serves as a (standalone) database containing the properties and compound specific reaction parameters of a large number of compounds. It assists the user to build up and/or to extend his own compound library.
(3) It provides the necessary programs to estimate compound properties from either data of structurally related compounds (that are present in the library) or from propertyproperty, structure-property, and structure-reactivity relationships.
The module "SYSTEMS" has the same general structure and tasks as the module "COMPOUNDS", that is:
(I) It assists the user in the description of the relevant properties of the phYSical system (e.g., a lake, river, groundwater aquifer) in which the behavior of the compound is to be evaluated. It also helps the user to choose the "right" set of system data that is appropriate with respect to the complexity of the model used.
(2) It serves as a data base for the description and comparison of different aquatic systems (e.g., different lakes). It enables the user to build up and/or to extend his systems library.
Finally, the module "MOQELS" combines the data derived from the two other modules and provides the mathematical framework for the desired model calculations. It provides the user with models of increasing complexity (e.g., one-box, two-box, multi-box models for lakes). In addition, it contains programs that enable the user to modify directly compound and system parameters for the compound and system at hand, without having to change the compound and system records in the two other blocks. This gives a maximum flexibility when trying, for example, to evaluate the sensitivity of the results of a model calculation towards changes in a given parameter. Hence the module "MODELS" is on a higher hierarchical level as compared to "COMPOUNDS· and ·SYSTEMS·, in that it not only provides the mathematical tools for the model calculations, but also manages all the data required to perform these calculations by using the two other modules as input
In addition to the above mentioned program modules, the MASAS system provides a series of additional features that may be very useful for certain applications. For example, whole case studies including compound records and the results of previous model calculations may be stored for later use. Furthermore, the program offers a variety of tabular and graphical output options. All output data may also be transferred to other software including word processing programs.
- 411-
4. IMPLEMENTATION
4.1 Choice of Hardware and Supporting Software
One of the major criteria for selection of the hardware was userfriendliness. Hence, a computer had to be selected that provides a user interface that is easy to learn, simple to understand and easy to operate. Furthermore, the computer had to provide extended graphical output features. Finally, in order to make MASAS widely applicable, a personal computer system had to be chosen. In our opinion, all these requirements are to date only optimally fulfilled by the Apple Macintosh computer series. Therefore, it was decided to develop the MASAS system on such a computer. This does, however, not mean that in the future, the program could not be implemented on other personal computers (e.g., IBM).
The software used for development of the MASAS system is the MacMETH-system (MODULA-2 developed bY.IFI at the Swiss Federal Institute of Technology (ETH) in Zurich). To facilitate the programming of the user interface, the program package "DIALOG MACHINE" (also developed at ETH) is employed.
4.2 State of Implementation
During the nine months that this project has been running so far (status Fall 1987), a significant part of the general software of the modules "COMPOUNDS" and "SYSTEMS" has been implemented. This includes editing programs for creating and updating compound and system libraries (so far only for lakes), programs that allow to combine compound and system parameters; as well as output programs for displaying data in tabular and graphic forms. Presently, models for evaluating the dynamics of pollutants in ~ are being implemented. The one-box and the two-box model are already available. They may be used alone or in combination to simulate the annual stratification/circulation cycle in a lake. A multibox-model which includes the lake topography and sediment/water interactions (Le., resuspension) will be implemented in the near future. The basic concepts of the lake models are described in detail by Schwarzenbach and Imboden (3) and Imboden and Schwarzenbach (4). A simple example to illustrate the application of MASAS is given in the appendix.
4.3 Outlook
In the near future (Le., within the next six to twelve months), the program package for modelling organic pollutants in lakes should be completed. It is also planned to augment the compound as well as systems (lakes) data bases. Finally, the development of programs aSSisting the user to apply a model of appropriate complexity, and a series of plausibility tests to minimize erroneous input data are intended to be implemented within this time frame. The next step will then be the implementation of river models.
REFERENCES
(1) POPPER, K. (1973). pie Logik der Forschung. 5th ed., J.C.B. Mohr, TObingen. (2) BURNS, L.A., CLINE, D.M. and LASSITER, R.R. (1982). Exposure Analvsis
Modeling System (EXAMS): User Manual and System Documentation. EPA-Report No. 600/3-82-03, US EPA, Athens, Georgia 30613.
(3) SCHWARZENBACH, R.P. and IMBODEN, D.M. (1984). Modelling concepts for hydrophobic chemicals in lakes, Ecological Modelling, 22,171.
(4) IMBODEN, D.M. and SCHWARZENBACH, R.P. (1985). Spacial and temporal distribution of chemical substances in lakes: Modelling Concepts. In: Chemical Processes in Lakes, W. Stumm, ed., Wiley Interscience, New York, pp 1-30.
- 412-
APPENPIX:
Simple Example illustrating the use of MASAS
A very simple example shall help to illustrate the use of MASAS. Using a one-box model we shall consider the steady-state concentration of the volatile pollutant perchloroethylene (PER) in a shallow, well-mixed test lake as a function of the average windspeed above the lake. Figure 2 shows the information record that is created by the module "MODELS" using the modules "COMPOUNDS" and "SYSTEMS" as input sources.
~o OnebOH-Model Run (One box): Onebox compound: PER
'~'tem : Leke Epi3010 .... eter[oKI: 293 .0 pAirletml : 0 .00 'w'; nd[ m/3 I: 1.00 ~iff.1iq.l-l : 0.510 diff.Qe,.[ - I: 0.300
It m.tren'f.[ m/hl : 7 .67E-03 extinc DOC[ m3/g/ml 3.00E- 0 1 extincS[ m3/g/m l : 5.50E-Ol
React . Rete,: ga, ex.[ l /hl : 1.28E-03 hlldrolll';'[ l/hl: O.OOE+OO photolll' ,[ l/hl : O.OOE+OO biologicel[ I/hl : O.OOE+OO ,ediment. [I Ihl: O.OOE+OO to tel[ 1 Ihl : 1.28E- 03
~ed .vel.!m/hl: 1.00E-Ol prod[mol/m3/hl : O.OOE+OO di33 .frec.[ -I: 1.00E+ 00 rene .... . rete [ l/hl : 1.20E-03
SteedvStete : oncln! mol/m31: 2.40 E- 0 7 n uxln[ mol/hl : 6.91 E-05 onc. ,bt.[ mol/m31 : 1.16E- 0 7 lime 10 33[dl: 1.68E+ 0 1
Figure 2: One-box Model data record for PER in the lake "Episolo".
The window consists of 9 boxes containing the most important parameters and results of the model. In the top three boxes we find a selection of parameter values from the modules "MODELS· and "COMPOUNDS" underlying the calculation. The next three boxes show the various first order (reaction and transport) rates. The bottom boxes give the relation between input concentration and system concentration at steady state together with the time needed to reach steady state within 37%.
From the data shown in Figure 2 we can, for example, see that only two processes, gas exchange ("gas ex.") and dilution by water renewal ("renew. rate"), determine the fate of PER in this lake. Note that at the chosen windspeed of 1 mls the two processes are about equally important since the corresponding (pseudo) first-order rate constants.have roughly equal size.
For evalutation of the effect of the windspeed on the gas exchange rate and on the steady-state concentration, the wind parameter (upper right box) can be directly varied in
- 413-
the "MODELS" module without going back to the "SYSTEMS" module. As can be seen in Figure 3, variation of the winds peed over an order of magnitude leads to a variation in the steady-state concentration of about a factor of four.
PER 0,1 2 0,010
-G- conc. slst.
0,1 0 gas ex.
0,008
!!! M
'" E x ""- 0,08 ~ ..; 0 0,006 Q)
c E '" :::::. 0 '" <> ~ 0>
0,06 0,004
0,04 0.002
0,02 ·0,000 0 2 4 6 8 10 12
windspeed
mls
Figure 3: First-order gas exchange rate constant (gas ex.) and steady state concentration (conc. stst.) of perchloroethylene (PER) in lake "Episolo" as a function of windspeed (measured at 10m above water surface).
Acknowledgements - This work is supported by the Commission of the European Community, and by the Swiss Environmental Protection Agency (BUS).
- 414-
SESSION VI
LABORATORY DATA TREATMENT
Chairman R. WEGMAN
Chemometrics in environmental analytical chemistry
CHEMOMETRICS IN ENVIRONMENTAL ANALYTICAL CHEMISTRY
H.A. VAN 'T KLOOSTER
National Institute of Public Health and Environmental Protection (RIVM)
Laboratory of Organic-Analytical Chemistry
P.O. Box 1 - 3720 BA Bilthoven - The Netherlands
SUMMARY
In environmental analytical chemistry chemometric tools are used in quantitative
rather than in qualitative analysis. In this paper chemometric concepts and
methods for the identification and structure elucidation of organic compounds are
discussed. Attention is focussed on the application of computer-aided library
search and artificial intelligence to the interpretation of spectrometric and
chromatographic data, whether or not combined. Library search systems for
mass spectra as. well as for combined ultraviolet spectra and HPLC retention data
were developed based on mathematical statistical models of the reproducibility of
the data involved. A pilot version of an expert system for structure analysis of
organic molecules was developed based on combined infrared and mass spectral
data, using artificial intelligence and information theory. Here too, the
significance of results is indicated by numerical (relative) probabilities. General
concepts and results are presented.
1. INTRODUCTION
Since the early seventies chemometrics is the name of the chemical discipline that
applies mathematical and statistical (formal) methods and computer techniques
a) to design or select optimal measurement procedures and experiments,
b) to provide a maximum of chemical information by analysis of chemical data [1].
Examples of chemometrical tools are: factorial design, experimental optimization, factor
analysis, pattern recognition, cluster analysis, principal components analysis, signal
- 416-
flitering, curve resolution, library search and (more recently) also artificial intelligence. Due
to the revolutionary development of computer technology chemometrical tools become
more and more powerful.
For computer-aided extraction of information about the identity or structure of organic
molecules from spectral and/or chromatographic data two main approaches can be
distinguished:
1) by comparison of analytical results with known reference data, using library search;
2) by identification of substructures from (combined) measured features (e.g. molecular
spectra) using empirical rules, by application of artificial intelligence.
Library search is primarily used for the identification of more or less frequently
occurring compounds of which reference data are likely available. If no reference data are
available, i.e. when dealing with unknown or completely new compounds, the second
approach applies. This situation might occur in the aquatic environment, for example when
new chemical products (e.g. pesticides) are released.
2. COMPOUND IDENTIFICATION BASED ON "MOLECULAR FINGERPRINTS"
The fact that mass, infrared, NMR and ultravioletspectra and to a certain extent also
chromatographic retention data can be considered as "molecular fingerprints" forms the
basis of most computerized library search systems. Retrieval methods for characteristic
chemical data and techniques for the comparison of human fingerprints have similar
elements: the first step is to clean-up the raw data, then in many cases a data reduction is
carried out by selection of prominent features. Finally, there is the comparison of unknown
and reference data patterns, which, for a useful result, requires a statistical correlation to be
established.
In library search three main items can be distinguished:
1) the method of feature selection;
2) the design of the similarity measure;
3) the quality (i.e. the reproducibility) of unknown and reference data.
In this paper no attention will be paid to feature selection. As for items 2 and 3: the design
of the similarity measure and the quality of the reference data have much to do with each
other. As a matter of fact, the n;producibility of the data involved in a search system is a
crucial element in the design of the similarity index. Especially the interlaboratory
r~roducibility plays an important role, whenever multi source databases are being used.
This reproducibility is determined by differences in samples, instruments, experimental
conditions, performance of analysts and operators, introduction of coding errors, etcetera.
etcetera. As a consequence a key factor determining the usefulness of computer-aided
- 417-
library search systems is the extent to which the reproducibility of the relevant data is
accounted for in the design of the similarity measure.
The following questions are imperative:
> what is considered to be "similar", what is "different"?
> which difference is "acceptable"?
> which difference is "significant"?
> which rational (formal) criterion is to be used?
3. A REPRODUCIBILITY-BASED SIMILARITY INDEX
The concept of a reproducibility-based similarity index will be introduced for a simple
one-dimensional (fictive) example. Suppose we want to identify a compound of which we
already know it's a methylpentanol, by library search based on gaschromato- graphic data.
We have measured a retention value of 740 units with a standard deviation of 5 units. The
following small library of reference data is available:
Compound Retention value (arbitr. units)
2-methyl-l- pentanol 804
3- -1- 816
4- -1- 803
2- -2- 717
3- -2- 778
4- -2- 736
2- -3- 762
3- -3- 747
Evidently, the 4-2 isomer and the 3-3 isomer have reference values being closest to the
measured value of 740. But why should we reject the 2-2 isomer or the 2-3 isomer? In
other words: which reference values are significantly different from the measured value -
say on the 1 % level of significance - and which are not? Our library search system should
provide a clear answer to these questions.
In comparing measured and reference data the statistical theoy of hypothesis testing
applies. Here, the null hypothesis Ho for every comparison is that the unknown compound
is identitical to the reference compound. The test parameter is the "difference QJlantity" .1q,
representing the difference between two retention values:
.1q = Retmeasured - Retreference. (1) In this case the reproducibility model of.1q is simply a normal probability distribution
- 418-
function p(ilq) under the null hypothesis, with mean = 0 and standard deviation = 7. The
parameter for testing Ho is the similarity index (SI), for our simple library search system
defmed as the integral of the reproducibility function, in this case a symmetrical Gaussian
curve:
00
81 = 2· J Po (Llq) dLlq ilq = ilQ
(2)
with ilQ being the actually measured value of ilq.
The integration is done from the point of the actually measured value of the difference
(ilQ) to infmity (2-sided):
Po (6q)
T Pa(6Q)
-6Q o --+)6q
For example: the similarity index value for an observed difference of 7 units is equal
to 2 x 16% = 32%. If the observed difference is zero, the SI equals the whole area under
the curve: 100%.
Our library search program, based On the above mentioned similarity index, allows the
specification of a minimum value for the similarity index. If, for example, a threshold
value of 1 % is specified (which corresponds to a significance level (X of 1 %) references
with an SI less than 1 % are rejected as possible candidates. The output of the retrieval
- 419-
system is a compound identity or a list of identities (if any) with SI values above the preset
threshold. For the pentanol problem our library search system gives the following output:
IN SEARCH FOR: UNKNOWN PENTANOL
SI (THRESHOLD) = 1.0 %
SEARCH RESULTS:
tID lli.%l. COMPOUND NAME
1 57.0 4-METHYL-2-PENTANOL
2 32.4 3-METHYL-3-PENTANOL
END OF HIT LIST
Thus, applying a significance level of (l = 1 %, the null hypothesis is accepted for the
4-2 and the 3-3 isomer and is rejected for all other isomers. The conclusion is that the
unknown methylpentanol is probably the 4-2 or the 3-3 isomer.
4. LIBRARY SEARCH FOR MULTIDIMENSIONAL DATA
For the comparison of analytical data, e.g. molecular spectra and/or chromatographic
retention data, a matching criterion in the form of a multidimensional similarity index has
been developed [2].
This similarity index requires that unknown and reference data are characterized by a
set of continuous 'feature quantities' ql ... qi ... qm' For mass spectra one might think: of the
peak intensities at a certain number of selected masses, for 13C-NMR spectra the feature
quantities could consist of the chemical shift values.
The actual comparison is made on the basis of the values of a set of 'difference
quantities' .1ql ... .1qi ... .1qm' representing the differences in value of the feature quantities
for the unknown and reference data, by calculating the value of the similarity index SI,
given by:
SI J ... J ... J Po (.1ql···.1qi .. ·.1qm) .1ql .1qi .1qm
R [.1Ql·" .1Qi ···.1QmJ
(3)
where Po is a probability density function called the 'reproducibility function', representing
the probability that difference quantity i has a value between .1qi and (.1Qi+MQi), in case
the reference compound considered is identical to the unknown compound (the 'null
hypothesis' Ho)' Further, .1Qi represents the actually observed value of the ith difference
quantity, while R is in the region of the multidimensional space of difference quantities
- 420-
defmed by the condition:
Po(~ql···~qi···~qm) < Po(~Ql···~Qi···~Qm) (4)
The model of reproducibility of the search data, as expressed by the reproducibility
function, forms the basis of the similarity index.
Application of the similarity index provides a classification of the references by
separating these in two classes: compounds that could be and compounds that cannot be the
unknown. In terms of hypothesis testing this is equivalent to acceptance or rejection of the
null hypothesis that the unknown and the reference compound are identical. A library
search system based on this principle should retrieve all references of the 'could be' class,
i.e. all references with a similarity index exceeding a predefined threshold value, rather than
the 5 or 10 'best matches' (which may also be very bad matches).
Based on this general concept, library search systems for mass spectra [2,3],
13C-NMR spectra [4], high-resolution-lH-NMR spectra [5] and ultraviolet spectra
combined with HPLC-retention data [6] were developed. Reproducibility models for
molecular spectra were elaborated from some hundreds (sometimes more than thousand) of
pairs of re.plicate spectra: different reference spectra of a same compound, recorded under
different experimental conditions [2-6].
5. THE CASSAM CENTER
The developed systems mentioned above are implemented in the national CAS SAM
Center in the Netherlands, CAS SAM being an acronym for: Computer-Assisted
Spectroscopic Structure Analysis of Molecules. The CAS SAM Center is (provisionally on
an experimental basis) accessible through national networks to Dutch universities and other
science institutes. Participants in the CASSAM project are the University of Utrecht, the
Netherlands Organisation for Applied Scientific Research (TNO) and the Netherlands
National Institute of Public Health and Environmental Protection (RIVM).
CASSAM data bases include:
• mass spectra (Wiley/McLafferty collection)
• 13C-NMR spectra (CIS-CID/fNO collection)
• infrared spectra (ASTM flle)
• 500 MHz-HNMR spectra of carbohydrates (Univ. Utrecht collection)
• UV spectra and LC retention data of organophosphorus pesticides (RIVM collection).
Extensions and updates are planned with new releases of commercial data bases and
reference flles made available by Dutch science institutes (CAS SAM users).
The CASSAM software consists of various programs for the comparison or interpretation
- 421-
of molecular spectra.
6 .LmRARY SEARCH OF MASS SPEcrRA
The Mass Spectral Reproducibility-based Retrieval (MSRR) system of the CASSAM
Center uses the Wiley-McLafferty data base of electron impact mass spectra. The 1987
release of this data base contains some 120.000 mass spectra of some 100.000 organic
compounds. A reproducibility model of the mass spectra was developed from some 1400
pairs of replicate spectra, originating from an earlier version of the data base (containing
some 39.000 spectra). This model is a mathematical statistical description of the observed
systematic and random differences in the selected features, being peak intensities at
maximum 24 selected masses [2,3]. The criterion for the selection of the peaks is based on
the (empirical) fact that peaks with relatively high intensities and peaks at relatively high
masses contain more information than small and low-mass peaks. For the similarity index
of the MSRR system a threshold value can be specified, analoguously to the
one-dimensional chromatographic example discussed above. An example of an output of
the MSRR system is given below.
100 rei. % intensity
t
------7 mass/
/charge
100 reI. % intensity
i
unknown
2-hydroxyxanthone
mol. W.: 212
emp.rormula: C13Ha03
212
200
212
(xO.S)
(xO.S)
B~~ ______ ~~ __ ~ ____ ~ ____ ~ ____ ~ ------7 mass/
/charge
200
- 422-
For an 'unknown' mass spectrum (A), in this case of a pre-identified component
(2-hydroxy-xanthone) extracted from a plant, the system retrieves from a library of 39000
reference spectra five references having a similarity index (SI) value of at least 5% as
compared with spectrum A. This similarity index can be considered as the (relative)
probability that the retrieved reference is identical to the unknown. On top of the 'hit list':
2-hydroxy-xanthone, of which the reference spectrum (B) has an SI-value of 69.2%.
References 2, 3 and 4 are isomers, of which the mass spectra are expected (on mass
spectrometric grounds) to show much similarity. Reference 5 is a different compound,
which however has some common structural features with hydroxyxanthones. As to the
SI-values: only an exact copy of a reference spectrum yields an SI value of 100%. Two
different spectra of a same compound, recorded under different experimental conditions,
however, always show differences which in some cases may be quite substantial. The
silimarity index used in this library search system takes account of such differences [2,3].
********************************************************************* CAS SAM CENTER - UTRECHf
MASS SPECTRAL RETRIEVAL SYSTEM
DATABASE: WILEY/MCLAFFERTY COLLECTION
*********************************************************************
IN SEARCH FOR:
TEST SAMPLE, PRE-IDENFIED AS: 2-HYDROXY-XANTHONE
RETRIEVAL RESULTS:
N.Q SIe&l SERNR MOLW FORMULA CQMPOUND NAME
1 69.2 14606 212 C13HS03 2-HYDROXY-XANTHONE
2 47.1 14604 212 C13HS03 3-HYDROXY-XANTHONE
3 46.9 14605 212 C13HS03 1-HYDROXY-XANTHONE
4 40.1 14603 212 C13HS03 4-HYDROXY-XANTHONE
5 S.l 14601 212 C13HS03 1,4-DIHYDROXYFLUORENONE
END OF HIT LIST
*********************************************************************
- 423-
7. LIBRARY SEARCH OF COMBINED HPLC AND UV DATA
The library search system for identification of pesticides based on diode-array UV
spectra combined with HPLC retention data uses an (experimental) reference data base
consisting of some 200 LC-UV data sets of organophosphorus pesticides, measured (at the
RIVM) from standard solutions under various experimental conditions [6]. For the
comparison of the UV data (normalized) absorbance values at 107 wavelenghts were used
as the feature quantities. The HPLC data are compared using a feature quantity in the form
of a capacity factor. A combined similarity index was developed based on the
reproducibility models of both the UV data and the HPLC data. The sample output shows
the identification of two organophosphorus pesticides as known test cases. For testcase 1
the hit list consists of one identity: the correct one. The second test case yields apart from
the correct compound also a structurally very similar compound, which shows hardly
different LC and UV data. For 95 % of some 100 "unknowns" the target reference (correct
positive) was on top of the "hit list", with a similarity index value being significantly
higher than values found (if any) for false positives .When the same eluent was used both
for the unknown and the reference (which fact usually is recorded and thus can be checked)
optimal results are obtained with combined UV and LC data [6].
************************************************ CASSAM CENTER - UTRECHT
HPLC/UV RETRIEVAL SYSTEM
DATABASE: RIVM COLLECTION/PESTICIDES
************************************************ IN SEARCH FOR: TESTCASE 1: CARBOPHENOTHION
RETRIEVAL RESULTS:
tID SIr&l CAS REG NR
1 98.0 786-19-6 CARBOPHENOTHION
END OF HIT LIST
IN SEARCH FOR: TESTCASE 2: DEMETON-S
RETRIEVAL RESULTS:
!:ill Sle&l CAS REG NR
1 45.8 126-75-0
2 41.6 298-03-3
END OF HIT LIST
NAME DEMETON-O
DEMETON-S
************************************************
- 424-
8. EXSPEC: AN EXPERT SYSTEM FOR STRUCTURE ANALYSIS OF MOLECULES
The purpose of a chemical expert system is to provide fast, easy, efficient and
effective access to chemical information and knowledge in a specific domain of expertise,
via computer-representation of Integrated reference data, theoretical and empirical
knowledge (e.g. in the form of 'rules'), system models and reasoning mechanisms. An
expert system should not only contain (integrated) relevant "hard facts" (e.g. numerical
correlations and statistics) but also the expertise ("soft knowledge") of experienced
specialists in the field.
For structure elucidation of organic molecules various spectrometric methods are
available. In many cases different methods such as mass, infrared, NMR and uiltraviolet
spectrometry may provide complementary structural information. This is one of the basic
elements of EXSPEC, an expert system for computer-aided interpretation of combined
spectra data [7-11].
EXSPEC is writen in PROLOG and runs on an Apple Macintosh II computer.
PROLOG is a flfth generation computer language, especially designed for the application of
artificial intelligence. Interpretation rules describe the relationship between certain molecular
substructures and spectral features. On the basis of a set of reference compounds,
containing these substructures of which infrared and mass spectra are available, an
automated 'rule generator' was developed [10]. By application of information theory the
conditional probability (p) of the presence of a structural unit (S), given a certain spectral
feature (for MS a peak representing a molecular or fragment ion at mass m/e, for IR an
adsorption in wavelength interval co), can be calculated. Using Bayes' theorem the
probability P(Sklcoi) of a functionality Sk' given an absorption coi for n functionalities
equals:
(5)
n with P(coi) = L P(Sk).p(~ISk) (6)
k=l
The available spectra can be read from disk or typed in from keyboard. Apart from
options input and interpret there is also an option explain, which provides the possibility
of requesting the reasoning process used to reach the conclusion. In the example shown a
secondary alcohol has been identified as a functional group with a probability of 99%.
Examples of dialogues with EXSPEC options input, interpret and explain are given
below. Under explain the rules specifying the relevant correlations between spectral and
structural features and the conditional probabilities are listed.
- 425-
:start
*** EXSPEC IRIMS INTERPRETATION *** Option: Input
Do you want to enter new spectra (n) or
update old ones (u) ? n
Do you want any old spectra to be deleted ? n
What spectra are available? Infrared Mass
Do you want to read an Infrared spectrum from disk? y
Infrared spectrum number? 178
Do you want to read a Mass spectrum from disk? y
Mass spectrum number? 178
Please. enter the molweight of the compound: 116
Option: Interpret
Possible molecular formulas are:
C5HS03
C6H1202
C7H1601
Alcohol has probability 0.9
Primary_alcohol has probability 0.01
Secondary_alcohol has probability 0.99
Tertiary_alcohol has probability 0.01
Phenol has probability 0.01
Option: Explain
What functionality do you want to have explained ?
Secondary_alcohol
Secondary_alcohol was found to have probability 0.99
because:
fragment (45 any) --> peS 1m/e) = 0.64
fragment (44 any) --> peS 1m/e) = 0.66
absorption (1003 950 moderate) --> peS I w) = 0.65
fragment (42 any) --> peS 1m/e) = 0.61
absorption (2977 2923 very_strong) --> peS I w) = 0.62
fragment (19 any) --> peS 1m/e) = O.SO
- 426-
9. THE EXSPEC STRUCTURE GENERATOR
In EXSPEC the generation of molecular structures is carried out in three steps [11].
As a ftrst step (not shown here) possible elemental compositions being compatible with a
specifted molecular weight and possible elements are calculated . .In the second step the
user selects a formula (here C4~CIO ) and the program determines which molecular
fragments (small or larger groups such as methyl, hydroxyl and fenyl) are compatible with
the selected formula. The user then makes a choice. In the example the fragment "-OH" has
been, the fragment "-0-" has not been chosen, by which the user indicates being interested
in alcohols and not in ethers having the formula C4~CIO. The program then determines
which combinations of fragments are plausible. In the third step the user selects a set of
fragments, for which the program then generates all possible unique (acyclic) structures in
the form of a "structure list". As an illustration of the interpretation of this list the
corresponding structures are drawn. Eventually, by processing the other selected fragment
sets the isomeric structures can be found (without duplications). As shown above the
EXSPEC module interpret "translates" spectral information into structural information, e.g.
in the form of statements concerning the presence or absence of a secondary alcohol group.
Integration of these modules provides the possibility to consider larger fragments, which
strongly reduces the number of possible structures. If, for example, it has been deduced
that the unknown compound contains a tertiary alcohol group, the total number of possible
structures with formula C4~CIO is reduced to only two.
: start
*** EXSPEC STRUCTURE GENERATOR Ver 4 ***
Molecular formula: (C 4 0 I CII H 9)
List of possible fragments :
I = "-CH3"
4 = ">C<"
2 = "-CH2-"
5 = "-OH"
Please enter numbers: (1 2 3 4 5 7)
Selected fragments:
3 "-CH<"
6 = "-0-"
("-CH3" "-CH2-" "-CH<" ">C<" "-OH" "-Cl")
*** I am thinking ...
- 427-
Molweight is 108.5
7 = "-CI"
*** Suitable combinations of fragments:
1 = ("-CI" (1) "-OH" (1) ">C<" (1) "-CH2-" (1) "-CH3" (2»
2 = ("-CI" (1) "-OH" (1) "-CH<" (2) "-CH3" (2»
3 = ("-Cl" (1) "-OH" (1) "-CH<" (1) "-CH2-" (2) "-CH3" (1»
4 = ("-Cl" (1) "-OH" (1) "-CH2-" (4»
Please enter number: (3)
Number of bonds: 10
Number of fragments: 6
Connectivity OK !
*** I am thinking ...
*** 6 unique structures found
*** UNIQUE SOLUTION # 1 > Structure list:
("-CH<" ("-CI") ("-CH2-" ("-OH"» ("-CH2-" ("-CH3"»)
*** UNIQUE SOLUTION # 2 > Structure list:
("-CH<" ("-OH") ("-CH2-" ("-CI"» ("-CH2-" ("-CH3")))
*** UNIQUE SOLUTION # 3 > Structure list:
("-CH<" ("-CH3") ("-CH2-" ("-CI"» ("-CH2-" ("-OH")))
*** UNIQUE SOLUTION # 4 > Structure list:
("-CH2-" ("-CH2-" ("-CI"» ("-CH<" ("-CH3") ("-OH"»)
*** UNIQUE SOLUTION # 5 > Structure list:
("-CH2-" ("-CH2-" ("-OH"» ("-CH<" ("-CH3") ("-CI")))
*** UNIQUE SOLUTION # 6 > Structure list:
("-CH2-" ("-CH2-" ("-CH3"» ("-CH<" ("-OH") ("-CI"»)
ACKNOWLEDGEMENT
The author gratefully acknowledges valuable discussions with doctors P. Cleij, G.I.
Kleywegt, H.I. Luinge (Utrecht University), C.E. Goewie and R.C.C. Wegman (RIVM).
- 428-
REFERENCES
1. D.L. Massart, A. Dijkstra and L. Kaufman, "Evaluation and optimization of
laboratory methods and analytical procedures", Elsevier, Amsterdam, 1978.
2. P. Cleij, H.A. van 't Klooster and J.C. van Houwelingen, "Reproducibility as the
basis of a similarity index for continuous variables in straightforward library search
methods", Anal. Chim. Acta, 150 (1984) 23.
3. P. Cleij, "Reproducibility of mass spectral peak intensities as the basis of an
automated library search method for identification of organic compounds",
Dissertation, State University of Utrecht, 1984.
4. R.W. Bally, D. van Krimpen, P. Cleij and H.A. van 't Klooster, "An automated
library search system for 13C-NMR spectra based on the reproducibility of chemical
shifts", Anal. Chim. Acta, 157 (1984) 227.
5. D.S.M. Bot, P. Cleij, H.A. van 't Klooster, H. van Halbeek, G.A. Veldink, J.F.G.
Vliegenthart, "Identification and substructure analysis of oligosaccharide chains
derived from glycoproteins by computer-retrieval of high-resolution 1H-NMR
spectra", J. Chemometrics, (in the press).
6. H.J. Boessenkool, P. Cleij, H.A. van 't Klooster, C.E. Goewie and H.H. van den
Broek, "A retrieval system for combined liquid chromatographic and diode-array
ultraviolet spectral data", Microchim. Acta, 1986 II (1987) 75.
7. H.J. Luinge and H.A. van 't Klooster, "Artificial intelligence used for the
interpretation of combined spectral data", Trends Anal. Chern., 4 (1985) 242.
8. H.J. Luinge and H.A. van 't Klooster, "Interpretation of combined infrared and mass
spectral data by application of information theory and artificial intelligence", Abstracts
COBAC IV Conference (Computer-Based Methods in Anal. Chern.), Technische
Universitlit Graz (1986), p. 5.
9. G.J. Kleywegt and H.A. van 't Klooster, "Chemical applications of PROLOG.
Interpretation of mass spectral peaks", Trends in Analytical Chemistry, 6 (1987)
55-57.
10. H.J. Luinge, G.J. Kleywegt, H.A. van 't Klooster, J.H. van der Maas, "Artificial
intelligence used for the interpretation of combined spectral data. Part 3. Automated
generation of interpretation rules for infrared spectral data", J. Chern. Inf. Comput.
Sci., 27 (1987) 95.
11. G.J. K1eywegt, H.J. Luinge and H.A. van 't Klooster, "Artificial intelligence used
for the interpretation of combined spectral data. Part 2. PEGASUS: a PROLOG
program for the generation of acyclic molecular structures", Chemometrics and
Intelligent Laboratory Systems (in the press).
- 429-
SESSION VII
ENVIRONMENTAL SCENARIO
Chairman T. LA NOCE
Future environmental problems
FUTURE ENVIRONMENTAL PROBLEMS
Finn Bro-Rasmussen Laboratory of Environmental Science and Ecology Technical University of Denmark, DK-2800 Lyngby
Summary
Referring to recent experience concerned with the deterioration of marine ecosystems and degradation of their environmental qualities, it is urged that the problems of the future are identical with or they are based on the environmental problems of today. Environmental damages are caused by macropollutants as well as micropollutants, they may be irreversible by nature and irrevocable in their development. Confidence in the prevention of future probl~ms is therefore closely related to the proper protection of the environment today. This prevention should be based on analysis and problem identification in their entirety, and methods of protection should concentrate on the sources of environmental impact - rather than on restorations and remedial actions at the level of environmental target systems.
1. INTRODUCTION I wish to thank the organizers for the invitation and
for this possibility to meet and address you in your individual capacities of analytical chemists and participants of the COST 641 project on Organic Micropollutants in the Aquatic Environment.
I have been encouraged to talk about Future Environmental Problems. Like you, however, I was originally trained as an analytical chemist, and I carry no crystal ball to assist me in my present task. Therefore, I shall try to share with you some of my recent 'experiences concerned with environmental problems. From these problem areas I shall attempt an extrapolation into other environmental problems - and possibly also in some cases into the nearest future.
The Symposium is limited to deal with the aquatic environment, and I shall certainly stay within these frames. I shall, however, take the liberty to go beyond the organic micropollutants to other important chemicals in order to widen the perspectives, and with the objective to focus on the aquatic ecosystem behaviour and quality as a whole.
- 432-
2. BASIC ELEMENTS OF ENVIRONMENTAL PROBLEMS In the conceptual analysis of most environmental
problems, it is often fruitful to distinguish between three phases of problem developments, i.e. to visualize the problem as a sequence of events which describe the chemical's movement (cfr. fig. 1) from:
- the anthropogenic SOURCES of the technological world, through
- the transport and transformation processes connected to environmental DISTRIBUTION until they reach
- the TARGETS of exposure, i.e. the biological organisms or ecological systems in which EFFECTS are to be observed and evaluated.
Rele8S8 Pathways Fate
@06 0 'U OQ ~ ~ ..L&I. 10 . C: QueUc
: ! Q Q ~ Or ~:
~ \enestllc:
I'en,po(\
Figure 1. Conceptual scheme for pollution problem development.
In our practical work with environmental problems, many of us are operating within the frames limited by only one of these phases, although we as analysts do possess scientific background for serving in anyone or all of them. In such situations, of course, we concentrate on the problems connected to that specific phase only, and we tend to block mentally for or to reduce the attention which should have been devoted to the other phases as well - as if we were captured in our own world of daily duties.
My initial observation, therefore, is that we should never evaluate an environmental problem without analysing the problem in its entirety, i.e. as a result of a chain of environmental events. If we do so, we may indeed increase the problems or even create new environmental problems for the future.
- 433-
3. EFFECTS OF MACROPOLLUTANTS In retrospect, we all know that this may be correct,
because most of us have experienced examples of "future problems" which have emerged and become reality before it was anticipated or even without being forecasted. Referring to such examples we may have to accept that the "future has already started".
Since the time when I personally first witnessed a nitrate contamination of Danish drinking water, i.e. about 20 years ago, this problem has developed from a single, though disturbing observation in the sphere of human health, to become a major environmental problem. The agricultural fertilizer usage is recognized as the source, and therefore also as the key to a major political issue. The uncertainties 20 years ago have manifested themselves as widespread environmental and human health hazards of today. Obviously, we did not at that time investigate and analyse the risk of distribution and infiltration sufficiently, and we did not accept the agricultural practise as the cause to an emerging problem.
Similarly, the infiltration of herbicides in recent years into ground water aquifers is already today a matter of concern in several European countries, and it is likely to become one of our major future problems, e.g. 10-20 years from now, if we do not take immidiate precautions at the source of agricultural uses.
Turning to the fresh water and to the marine compartments of the environment, we are increasingly aware of problems which have developed in recent years, and which are likely to develop further into problems of the future. As the result of distribution of inorganic nutrients and organic waste material from runoff and urban discharges, we now witness a number of disturbances and disruptions in the natural ecosystems allover Europe. Among these, I am personally most familiar with the threatening trends in the Northern regions, i.e. the Baltic Sea area, the Danish Inland waters, and the North Sea, although I suspect that the situation at the Southern European scene, e.g. the Mediterranian Sea, is much the same.
In the Baltic Sea and throughout the Danish straits and belts, up till 25% or more of the bottom area is regularly under conditions of oxygen depletion, and even in the North Sea, espec. in its eastern parts, considerable bottom areas of anoxic conditions have been identified regularly within the last years. The individual significance and weight of various components of pollution, i.a. nitrogen, phosphorus and organic material varies geographically and according to seasons of the year. In general, however, it is accepted that individual episodes and the situation as a whole are the results of eutrophication and excessive loading with organic materials, followed by sedimentation and detrimental circulation of phosphorus and other nutrients, all of which are not sufficiently flushed out through the narrow Danish belts to the North Sea and into the oceanic waters.
- 434-
The ultimate and dramatic effects of this marine eutrophica~ion is observed by commercial fisheries, who in recent years have experienced a nearly complete failure in the late summer and autumn activities as far as benthic catches are concerned, e. g. plaice, eel and lobsters. In contrast, the potential for harvest of pelagic fish seem to have increased. These can be cod, herring and salmon, which -at least temporarily - seem to benifit in growth and in reproduction from the boosted plancton production in the open waters, and thereby give the false impression of a well-functioning, productive ecosystem.
Concealed in the fisheries statistics, however, is the reduction in the widespread communities of macroalgae, seaweeds and coastal marine flora which earlier characterized all coastal lines in the region. Among other things, they gave shelter for the breeding and reproduction of many marine species, which are now under threat of extinction, adding further to the description of the marine environment as a distorted aquatic ecosystem. The changing living conditions resulting from pollution cause a shift in the balance between an overproductive pelagic biota on the one side, and a degradation of coastal and benthic communities on the other. This endangers many individual species, and threatens the ecosystem quality as a whole.
It is a matter of course, that this situation gives rise to much public - as well as political - concern in all neighbouring countries, not the least the Scandinavian countries who are by tradition highly dependant of the lasting quality and productivity of their surrounding marine waters. It is mentioned therefore, that political measures were initiated e.g. in Denmark in 1986/87, and recent parlamentary planning has set targets for the next 5-6 years concerning reduction of Danish emissions of all macropollutants into the aquatic. environment, e. g. by 50% for nitrogen and 80% of phosphorus from all sources. These targets are not fully in accordance with the estimated pollution loads, and - although they most certainly will give a relative improvement of the present situation - it remains to be seen whether they will be sufficient to restore the disrupted environmental qualities concerned.
4. EFFECTS OF MICROPOLLUTANTS Parallel to the described situation, there is an
accumulating stock of evidence on effects of chemicals other than nutrients which are observed in several species living in locally polluted waters. These may be examples of morphological damages, e.g. spinal deformities in fish exposed to local discharges of heavy metals and pulp mill effluents, physiological disorders manifested in histological, haematological or enzymatical parameters, or possibly skeletal anomalies, all of which can be related with a high degree of probability to certain sources of pollution (2).
The future problems in these cases seem to be rather complex, i.e. connected to the needs for research and for better understanding of the ecotoxicological cause-effect relations, as well as to the necessity for legislation and
- 435-
management in order to restrict pollution and restore the ecosystem in both structure and functions.
This has already been illustrated in earlier cases, e. g. in connection with the group of chemicals on which we have the most comprehensive experience, namely organochlorines such as DDT and PCB. The chemical fate and effects of these are well known for several species of fish-eating birds and mammals, and examples are multible. Mentioned is for instance the harmful effects of DDT/PCB on white-tailed eagles with greatly reduced reproduction rates in the highly polluted regions, or the failing hatchability of seabirds, e.g. guillemots and cormorants in the Baltic proper, or the disastrous disappearance - close to extinction - of fisheating mammals, such as otters and seals in the Baltic Sea region. Some of these situations now seem to be repeated in the Wadden Sea and the German Bight environment of the North Sea.
These observations were made after an introductory period of use of the chemicals in the years following 1945. They closely correlated the findings of high concentrations and accumulative potential of the.organochlorine pollutants, and they were further followed by a period of intensive research, of international regulations, and also by a period of gradual - although far from complete - recovery.
Obviously, if we consider the number of potential and positively identified micropollutants of which we are aware today, it is neither realistic nor advisable to believe that our future environmental problems can be solved in the same manner; at least not if the requirements for time, for research and for detailed hazard evaluations has to be framed and carried out in the same scale and intensity as in the examples. In the long run, a much more predictive monitoring activity has to be installed, and precautionary measures must be taken earlier and with responsibilities placed directly and obliging at the level of primary polluter and/or polluting activity.
5. ENVIRONMENTAL DISTRIBUTION AND FATE In our study of pathways and fate of environmental
chemicals, we have today achieved a relatively good knowledge on several transport mechanisms and a fair understanding of the distribution processes for chemicals between and within the diffe~ent compartments of the environment. This is at least the case for those wellstudied chemicals, such as organochlorines (e. g. DDT and PCB) and organometallic compounds (e.g. methyl-mercurials), which are lipophilic and of high accumulative potential.
Going beyond these chemicals, however, a symposium like the present and its immidiate preceeding COST 641 meetings, has contributed with SUbstantial evidence on the fragile nature of our knowledge. A number of papers presented here deals with studies of properties of individual chemicals and of single processes, but they are few when related to the total number and needs, and they still leave us with more questions and future problems unanswered than before.
without going into details, I refer to the presented example of significant Swedish research concerned with the microbiological transformation processes of natural and
- 436-
conditioned phenols found in the effluents from pulp mills (3). The so-called Guuajacol-cycle resulting from these studies points to the environmental risk connected to differing metabolic patterns under various conditions, and to transformation of chemicals into metabolites which are more hazardous than the parent compounds.
This is due to formation of a series of compounds belonging to the polychlorinated quaiacol-catechol-veratrole family, which are produced in methylation and demethylation processes induced microbiologically, e.g. under aerobic and anaerobic conditions. The individualcompounds are free or they are found in _ conjugates in the aquatic environment, either in the water, the sediment and in plant- and animal tissues.
The immidiate problems imposed by such findings, namely the difficulties in approaching the hazard assessment in a classical way are obvious. This has already been mentioned by dr. Neilsson and his co-workers, and need not be repeated here. Seen in a wider perspective, however, they may give rise to future problems of another - and mostly unnoticedcharacter, nam-ely the uncertainty whether research results and hazard evaluations of today are still valid and applicable in coming years. The risk is that the total environment and its intrinsic conditions have changed, e.g. under the stress from pollution with other chemicals, and as we have already seen, an excess of nutrients and/or other materials can be decisive for the primary ecosystem quality and functions, as when they determine the condi tions of oxygen supply or depletion.
6. CHEMICAL EMISSIONS When we deal with the term "environmental problems",
this may refer to the damages or to the unacceptable changes - i.e. the effects - caused by pollutants, or we may equally justified from a holistic point of view identify the problems and deal with them as connected to the cause of environmental effects, or even in some cases the sources of detrimental pollution.
I do not wish to go into detailed discussion on this aspect. Personally I find, however, that we are at risk of leaving many environmental problems of today unsolved (or partly solved only), and thus to let the problems continue as problems of the future. Speaking in very generalized terms, I find reason to believe so from many examples of reluctance in the society - or even among experts - to accept environmental effects as problem areas. We also often see a lack of acceptance of the need for reactions or restrictions towards present pollution practices, or alsowhen such obstacles do not exist - in some cases there is a failing ability (or even lack of will) to interfere in the pollution process due to other consequences, e.g. economical or political.
In the case of marine pollution which has been my theme of choice today, only few people doubt that the ideal is to avoid polluting discharges, while at the same time the majority - and even many among us - are still hesitating to impose - as an example - legal restrictions on existing emissions (e.g. as those indicated in table 1) in order to
- 437-
reduce them to a practical Zero and thereby restore the ecological quality of the sea.
Table 1
POLLUTION LOAD FOR THE BALTIC SEA AREA, 1986 ().
Substance unit From land From atmosIlhere
Nitrogen, total t yr-1 530.000 413.000 Phosphorus, tot. t yr-1 42.000 6.000 BOD t yr-1 1. 640.000 Mercury t yr-1 .5 Cadmium t yr-1 60 80 Zinc t yr-1 9.000 3.200 Lead t yr-1 300 2.900 Copper t yr-1 4.200 380 oil t yr-1 36.000 Arsenic t yr-1 180 Nickel t yr-1 110 Vanadium t yr-1 290 Chromium t yr-1 0.2
The latter question may become critical as in the case of the already mentioned 50% and 80% reductions of nitrogenand phosphorus-emissions, respectively. Are these reductions sufficient to restore the ecological quality of the sea ? Scientifically, we do know that they are only partial if they are measured against the full anthropogenic load emitted via modern urban, industrial and agricultural practices. The relevant question, therefore, may rather be whether we by reducing our prese~t discharges and emissions - but not eliminating them ! - in fact are preserving our environ-mental problems instead of sol ving them ? Or, if we even create new problems, because we have insufficient experience and we can give no assurance that biological successions in still polluted and disturbed, marine ecosystems will be the same as those we normally describe for unpolluted ecosystems ?
7. CONCLUDING REMARKS In summing up, it is my belief that we have only
limited time, space and facilities available from which we can forecast - and therefore meet the challenge of future environmental problems. The obvious way of forecasting - as it has been my aim to show through few and only illustrating examples - is to describe the present state of environmental disturbances as the base for extrapolations. This is then attempted by utilizing the experiences from those developments and those processes, which have led to the present state.
- 438-
processes, which have led to the present state. Obviously, the outcome of this is bound to be uncertain
at its best. possibilities are open not only for mistakes in our forecasting, but also for damages which prove to be irreversible. The risk of frrevocable failures with unpredictable consequences is imposed on the environment by us, but the costs will be carried by future generations, who are as dependant on the lasting quality and functions of this environment as we are.
The greatest problem, therefore, for our future environment is to be found in the needs for comprehensive and interdisciplinary analyses and for planning in terms of predicting and preventing the problems - rather than our present practice of describing the polluting development, followed by trial-and-error reductions, and attempts to restore ecosystem qualities on the basis of single quality parameters.
We are left with no choice. We have to accept, that our present environmental problems are real problems, and that they are of significance for the future. We must accept, that the environmental problems are our own problems, i.e. they are of concern not only for the environment in itself, but for the human society and for us as members of this society, today and in the future. And finally, we must combat and sol ve the future problems by acting now and by preventing the problems of today instead of waiting until the future has come.
REFERENCES:
(1) Baltic Marine Environment Protection commission (1987). First Baltic Sea Pollution Load Compilation. Report No. 20 from the Helsinki Commission.
(2) ESTHER-project (1987). Test and Evaluation of Hazards of Chemical Substances. SNU report No. 3375 (in Swedish). Research project sponsored by the Swedish National Environment Protection Board, Solna, Sweden.
(3) Neilsson, A. Allard, A-S. & Remberger, M. (1985). Biodegradation and Transformation of Recalcitrant Compounds. In Handbook of Environmental Chemistry, Vol. 2, page 29-86. Hutzinger, O. (ed.). Springer Verlag, Berlin.
- 439-
LIS T OF
J. ALBAIGES Environmental Chemistry Dept. CID, CSIC C/Jorge Girona Salgado 18-26 E - 08034 BARCELONA Spain
G. ANGELETTI Commission of the European Communities DG XII/E-1 Rue de la Loi, 200 B - 1049 BRUSSELS Belgique
D. BARCELO Environmental Chemistry Dept. CID, CSIC C/Jorge Girona Salgado 18-26 E - 08034 BARCELONA Spain
G. BECHER Dept. of Toxicology National Inst. of Public Health Geitmyrsveien 75 N - 0462 OSLO Norway
M. BENEDINI IRSA - CNR Via Reno, 1 I - 00198 ROMA Italy
M.J. BENOLIEL Istituto Hidrografico R. Das Trinas 49 P - 1296 LISBOA Portugal
PAR TIC I PAN T S
J. BISCAYA Instituto Hidrografico R. das Trinas 49 P - 1296 LISBOA Portugal
A. BJ.0RSETH SCATEC P.O. BOX 147 N - 1312 SLEPENDEN Norway
S. BONACCI Istituto di Igiene dell'Universita P.le Aldo Moro I - 00161 ROMA Italy
K. BOOIJ Netherlands Institute for Sea Research P.O. box 59 NL - 1790 AB DEN BURG The Netherlands
D. BOTTA Pol itecnico Dipartimento di Chimica Industriale e Ingegneria Chimica Piazza Leonardo da Vinci, 32 I - 20133 MILANO Italy
J. BRODESSER Hygiene Institut Venusberg Sigmund-Freud Str. D - 5300 BONN 1 Deutschland
F. BRO-RASMUSSEN Lab. of Environmental Science and Ecology Technical University of Denmark Building 224, DK - 2800 LYNGBY Denmark
- 441-
A. BUCHERT National Food Agency 19 Morkhoj Bygacle DK - 2860 SOBORG Denmark
A. BUSA ESSO Ital iana V.le Castello della Magliana 25 00148 ROMA Italy
J. CAIXACH Centro de Investigacion y Desarrollo - C.S.I.C. Jorge Girona Salgado 18-26 E - 08034 BARCELONA Spain
B. CALLAN An Foras Forbartha St. Martin's House Waterloo Road IR - DUBLIN Ireland
R. CALVERLEY Analytichem International P.O. BOX 234 UK - CB2 INN CAMBRIDGE England
P. CAPEL EAWAG CH - 8600 DUBENDORF Switzerland
S. CAPRI IRSA - CNR Via Reno, 1 I - 00198 ROMA Italy
G. CARLBERG Center for Industrial Research P.O. BOX 124, BLINDERN N - 0314 OSLO Norway
J.M. CARRASCO-DORRIEN Universidad Politecnica de Valencia E.T.S.I. Agronomos Camino de Vera, 14 E - 46022 VALENCIA Spain
A. CHICOTE AYUSO Centro de Estudios Hidrografico Paseo Bajo Virgen del Puerto 3 E - MADRID Spain
B. CRATHORNE Water Research Centre Henley Road, MARLOW, UK - SL7 2HD BUCKS England
C. CREMISINI Lab. Geochimica Ambiente ENEA CRE Vi a Anguillarese I - CASACCIA - ROMA Italy
F. CUBILLO GONZALEZ Direc. General de Recursos Hidr. Consejeria de Obras Publicas y Transportes Orense, 60 E - 28020 MADRID Spain
M.A. L. DA CRUZ CAVACO Empresa Publica de Lisboa P - LISBOA Portugal
- 442-
L.F. DE ALENCASTRO Institut du Genie de L'Environnement EPEL EcubLens
CH - 1015 LAUSANNE Switzerland
A. DELOGNO Dipartimento di Chimica dell'Universita P.le Aldo Moro I - 00161 ROMA Italy
F. ·DE SMEDT Laboratory of Hydrology VUB Pleinlaan 2 B - 1050 BRUSSELS Belgium
R. DIAS PIMENTEL Estado do Ambiente e dos Recursos Naturais Rua Antero de Quental, 44 P - 1100 LI SBOA Portugal
A. 01 CARO Istituto di Igiene dell'Universita P.Le ALdo Moro I - 00161 ROMA ItaLy
C. 01 PALO ENEA CRE Casaccia PAS SCAB ECOL Via Anguillarese Casaccia I - 00100 ROMA Italy
M. DORE Universite de Poitiers 40 Avenue du Recteur Pineau F - 86022 POI TIERS Cedex France
V. DREVENKAR Inst. for Medical Research Mose Pijade 158 YU - 41000 ZAGREB Yugoslavia
J.P. DUGUET Laboratoire Central Lyonnaise des Eaux Rue du Pt Wilson 38 F - 78230 LE PECQ France
D. DUQUET RSL Alltech Europe Begoniastraat 5 B - 9731 EKE Belgium
I. ESPADALER Centro De Investigacion y Desarrollo - C.S.I.C. Jorge Girona Salgado 18-26 E - 08034 BARCELONA Spain
M. FIELDING Water Research Centre WRC, Henley Road MEDMENHAM MARLOW, UK - SL7 2HD BUCKS EngLand
S. FINGLER Inst. for Medical Research M. Pijade 158 YU - 41000 ZAGREB Yugoslavia
P. FLANAGAN An Foras Forbartha St. Martin's House, Waterloo Road, IR - DUBLIN 4 Ireland
- 443-
J. FOLKE Cowiconsult Teknikerbyen 45 DK - 2830 VIRUM Denmark
C. FORST Kernforschungszentrum Karlsruhe IHCh Postfach 3640 D - 7500 KARLSRUHE Deutschland
D. FRAISSE CNRS - SCA P.B. 22 F - 69290 VERNAl SON France
G. FRIEGE Landesamt fur Wasser und Abfall D - DUSSELDORF Deutschland
S. GALASSI IRSA - CNR Via Occhiate I - 20047 BRUGHERIO Italy
M.T. GALCERAN Department of Analytical Chemistry University of Barcelona Diagonal 647 E - 08028 BARCELONA Spain
S. GEIST Federal Environment Agency Radetzkystrasse 22 A - 1030 VIENNA Austria
W. GIGER EAWAG CH - 8600 DUBENDORF Switzerland
N.J. GONZALES Direc. General deL Medio Ambiente Ministerio de Obras Publicas E - MADRID Spain
P.S. GRIFFIOEN Institute for Inland Water Management and Waste Water Treatment NL - 8200 LELYSTAD The Netherlands
E. GJESSING National Institute of Public Health N - 0462 OSLO 4 Norway
B. HARLAND Imperial Chemical Industries PLC Freshwater Quarry, Overgang, UK - BRIXHAM, Devon England
P. HENSCHEL Umweltbundesamt Bismarckplatz 1 D - 1000 Berlin 33 Deutschland
B. HOLMBOM Abo A.kademi SF - 20500 TURKU/ABO Finland
- 444-
B. HURNI A. LORETO Amt fur Umweltschutz, Rheinstrasse 29
Kanton Basel-LandCEV Centro Studi E.V. Umana Viale Tor di quinto 58
CH - 4410 LIESTAL Switzerland
A. KLOSE Commission of the European Communities DG XII/A/5 Rue de la Loi, 200 B - 1049 BRUSSELS Belgique
KODRIC-SMIT M. MED. Centre Y - SISAK Yugoslavie
K. KOLSET Center for Industrial Research P.B. 350 Blindern 0314 N - OSLO 3 Norway
T. LA NOCE Istituto di Ricerca sulle Acque C.N.R. Via Reno, 1 I - 00198 ROMA Italy
P.G. LAUBEREAU Hess. Landesanstalt fur Umwelt Unter den Eichen 7 D - 6200 WIESBADEN Deutschland
A. LIBERATOR I Istituto di Ricerca sulle Acque C.N.R. Via Reno, 1 I - 00198 ROMA Italy
I - ROMA Italy
J. LUNA MARINO Confederacion Hidrografica del Guadiana - MOPU Avda de Santa Marina, 33 E - BADAJOZ Spain
G. MACCHI IRSA Via Reno, I - 00198 ROMA Italy
M. MAKELA National Board of Waters and Environment P.O. Box 250 SF - 00101 HELSINKI Finland
E. MANTI CA Politecnico - Dipartimento Chimica Industriale e Ingegneria Chimica Piazza Leonardo da Vinci 32 1-20133 MILANO Italy
A. MARCOMINI University of Venice Dorsoduro 2137 I - 30123 VENICE Italy
R. MASSOT Commissariat a l'Energie Atomique SEA CENG B.P. 85 X F - 38041 GRENOBLE France
- 445-
S. MATSUI Kyoto University Lab. for Control of Environmental Mi cropollutants 1-2 Yumihama, OTSU CITY, Shiga, 520 Japan
P. MIKKELSON University'of Jyvaskyla Killikinkatu 1-3 SF - 40100 JYVXSKYLX Finland
N. MIMICOS NRCNS DEMO Aghia Paraskevi Attikis GR - 15310 ATHENS Greece
A. MINDERHOUD National Institute of Public Health Laboratory for Ecotoxicology Antonie van Leeuwenhoeklaan 9 P.O. Box 1 NL - 3720 BA BILTHOVEN The Netherlands
A. MOLLER AGRI Contact Torupvejen 97 DK - 3390 HUNDESTED Denmark
S. MONARCA Dipart. di Igiene Universita di Perugia Via del Giochetto I - 06100 PERUGIA Italy
M. MULLER Swiss Fed. Research Station Wadenswi FAW CH - 8820 WADENSWIL switzerland
A. NEILSON Swedish Environmental Research Institute BOX 21060 S - 10031 STOCKHOLM Sweden
N. NYHOLM Water Quality Institute 11, Agern Alle DK - HORSHOLN Denmark
N. NUNZIA Unione Sanita Locale USL RM1 I - ROMA Italy
C. O'DONNELL An Foras Forbartha ST. Martinis House Waterloo road IR - DUBLIN 4 Ireland
H. OTT Commission of the European Communities DG XII/E-1 Rue de la Loi, 200 B - 1049 BRUSSELS Belgique
J.R. PARSONS Laboratory of Environmental and Toxicological Chemistry University of Amsterdam Nieuwe Achtergracht 166 NL - 1018 WV AMSTERDAM The Netherlands
R. PAUKKU Institute for Environmental Research University of Jyvaskyla Fellervonkatu '8 SF - 40100 JYVASKYLA Finland
-446-
N. PEPE Ufficio Stampa CNR P.le Aldo Moro, 7 I - 00185 ROMA Italy
J. PRADA ALVAREZ-BUYLLA Confederacion Hidrografica del Guadiana - MOPU C/Tinte, sIn E - CIUDAD REAL Espana
G. PREMAZZI Joint Research Centre I - 21020 ISPRA (Varese) Italy
M. PRIHA The Finnish Pulp and Paper Research Institute P.O. Box 136 SF - 00101 HELSINKI Finland
M. PRINCI USL Triestina - PMP Servizio Chimico Ambientale V. Lamarmora, 13 I - 34139 TRIESTE Italy
D. QUAGHEBEUR Instituut voor Hygiene en Epidemiologie Juliette Wytsmanstraat 14 B - 1050 BRUSSELS Belgium
S. REKOLAINEN National Board of Waters and Environment P.O. Box 250 SF - 00101 HELSINKI Finland
- 447-
L. RENBERG National Swedish Environmental Protection Board BOX 1302 S - 17125 SOLNA SwedeI'!
O. RINGSTAD Center for Industrial Research P.O. BOX 124, Blindern N - 0314 OLSO Norway
J. RIVERA Centro de Investigacion Y Desarrollo C.S.I.C. Jorge Girona Salgado, 18-26 E - 08034 BARCELONA Spain
M. RONCHETTI Dipartimento di Chimica dell'Universita P.le Aldo Moro I - 00100 ROMA Italy
R. SANCHEZ CRESPO Confederacion Hidrografica del Guadiana - MOPU C/Tinte sIn E - CIUOAD REAL Spain
M. SANNA USL Roma 10 Via Saredo, 52 I - 00100 ROMA Italy
M. SANTOR I IRSA - CNR Via Reno, 1 I - 00198 ROMA Italy
R. SAVOIR Laboratoire de la CIBE 764 Chaussee de Waterloo B - 1180 BRUXELLES Belgique
S. SCHMIDT German Standards Inst. H.-T.-V- Buttinger - Str.8 D - 5090 LEVERKUSEN 1 Deutschland
H.F. SCHOLER Hygiene Institut Venusberg Sigmund-Freud-Str. D - 5300 BONN Deutschland
G. SCHRAA Dept. of Microbiology Agricultural University H. Van Suchtelenweg 4 NL - 6703 CT WAGENINGEN The Netherlands
S. SCHRAP Lab. Environmental and Toxicological University of Amsterdam Nieuwe Achtergracht 166 NL - 1018 AMSTERDAM The Netherlands
R. SCHWARZENBACH EAWAG CH - 8600 DUBENDORF Switzerland
S.P. SCOTT Thames Water Authority 177 Rosebery Ave. UK - ECIR 4 TP LONDON England
L. SEBASTIANI Annichiarico IRSA Via Reno 1 I - 00198 ROMA Italy
Z. SMIT Medical Center Sisak Department of Sanit. Chemistry and Ecology Tomislavova 1, Y - 44000 SISAK Yugoslavia
J.A. SOARES SILVA MATOS Direccao Geral Gualidade Ambiente Apartado 155 P - 7501 SANTO ANDRE codex Portugal
S. SPADONI Unione Sanita Locale USL RM1 I - 00100 ROMA Italy
Chemistry E. STEPHANOU University OF Crete P.O. Box 1470
-448-
GR - IRAKLION Greece
E. STORR-HANSEN Environmental Protection Agency Kemikontrol 27, Morkhoj - Bygade DK - 2860 SOBORG Denmark
O. SVANBERG National Environmental Protection Board Laboratory for Aquatic Toxicolo~ Studsvik S - 61182 NYKOPING Sweden
M.L. TOSATO Universita Lab. Tossicologia Comparata Istituto Superiore di Sanita 299 P.le Aldo Moro 5 I - 00161 ROMA
F. VAN HOOF Studie Centrum von Water Mechelsesteenweg 64 B - 2018 ANTWERP Belgium
J. VENNEKENS Commission of the European Communities DG XIlA/2 Rue de la Loi, 200 B - 1049 BRUSSELS Belgique
F. VENTURA C.S.I.C. Jorge Girona Salgado 18 - 26 E - 08034 BARCELONA Spain
M.P. M. VIANA D.G.G.A. - Centro de Investiga~ao do ambiente Av. Alm. Gago Coutinho n.30 P - 1300 LISBOA PortugaL
L. VIRKKI University of Jyvaskyla, Kyllikinkatu 1-3 SF - 40100 JYVASKYLA Finland
A. WAGGOTT Water Research Centre Elder Way UK - SGI ITH STEVENAGE, HERTS England
- 449-
C. WATTS Environmental Pathways Medmenham Laboratory, Henley Road, Medmenham, P.O. Box 16 MARLOW, U.K. - BUCKS SL7 2HD EngLand
R. WEGMAN NationaL Institute of Public Health and Environmental Hygiene P.O. Box 1 NL - 3720 BILTHOVEN Netherlands
A. WEIGHTMANN Dept. of Applied Biology Univ. of WaLes UK - CARDIFF EngLand
C. WOLFF SHELL Int. Petroleum Mij B.V. P.O. Box 162 NL - 2501 AN THE HAGUE The Netherlands
J. YATES Water Research Centre ELder way UK - SGI ITH STEVENAGE, HERTS EngLand
J. ZEYER EAWAG CH - 6047 KASTANIENBAUM Switzerland
L. ZOCCOLILLO Dipartimento di Chimica dell'Universita P.le Aldo Moro I - 00100 ROMA Italy
AERTS, S., 318 ANSELME, C., 299 ALBAIGES, J., 75 ALLARD, A.-S., 228 ASCHJEM, B.F., 372 ASYEE, G.M., 176
BARCELO, D., 75 BECHER, G., 353
I N D E X
BEN AMOR, H., 338 BENOLIEL, M.J., 83 BONACCI, S., 357 BRO-RASMUSSEN, F., 432 BRODESSER, J., 69 BRUNNER, P.H., 266
CAIXACH, J., 329 CALVERLEY, R.A., 31 CAPEL, P.D., 189 CAPRI, S., 266 CARLBERG, G.E., 347 CARRASCO, J.M., 59 CARTON I , G.P., 116 CHRISTOPHERSEN, N., 372 CONTI, R., 357 CUBILLO, F., 402
DAVIES, I.W., 97 DE LAAT, J., 338 DE SMEDT, F., 387 DE TORRES, M., 329 DELOGU, A., 116 DEWAELE, C., 14 DI CARO, A., 357 DORE, M., 338 DRANGSHOLT, H., 347 DREVENKAR, V., 198, 238 DUGUET, J.P., 299 DUQUET, D., 14
o F AUT H 0 R S
ESPADALER, I., 329
FIELDING, M., 284 FIGUERAS, A., 329 FINGLER, S., 238 FORST, C., 52 FRAISSE, D., 329 FRIEGE, H., 132
GALASSI, S., 108 GALCERAN, M.T., 46 GIGER, W., 189, 266 GJOS, N., 347 GOWLING, R.W., 103 GRELLA, A., 357 GRIFFIOEN, P.S., 144 GUZZELLA, L., 108
HARLAND, B.J., 103 HEIBERG, A., 372 HELDSTAB, J., 408 HERVE, S., 91 HOLMBOM, B., 278 HOJ;lTH, H., 284 HURNI, B., 128
IMBODEN, D.M., 408
JAMES, H., 2 JANSSENS, J., 312 JOHNSEN, S., 353
KEELING, R.L., 2 KNUUTILA, M., 91 KNUUTINEN, J., 88, 344 KODRIC SMIT, M., 198
- 451-
KOLSET, K., 372 KRONBERG, L., 278
LANDMARK, L.H., 347 LINDGREN, C., 228
MALLEVIALLE, J., 299 MANNILA, P., 344 MARCOMINI, A., 266 MAZOUNIE, P., 299 MELCHIORRI, C., 357 MIKKELSON, P., 88 MIMICOS, N., 184 MOORE, K., 154 MYLONA, A., 184
NEILSON, A.H., 228 NYHOLM, N., 256
OPPERHUIZEN, A., 170, 176
PAASIVIRTA, J., 88, 91, 344 PARSONS, J.R., 176, 206 PASQUINI, R., 357 PAUKKU, R., 91 PLUYS, E., 312 PRADA ALVAREZ-BUYLLA, J., 323 PSATHAKI, M., 121
REKOLAINEN, S., 195 REMBERGER, M., 228 RENBERG, L., 244 RIVERA, J., 329 RONCHETTI, M., 116
SANCHEZ CRESPO, S., 323 SANTOS, F.J., 46 SAVINO, A., 357 SCHMIDT, S., 22 SCHOELER, H.F., 69 SCHRAA, G., 215 SCHRAP, S.M., 170 SCHWARZENBACH, R.P., 408 SCOTT, S. P ., 2 SEBASTIANI, L., 357 SHIRLEY, M.L., 83 SIJM, D.T.H.M., 206 SLINGERLAND, P., 62 SMIT, Z., 198 STEPHANOU, E., 121, 184 STIEGLITZ, L., 52 STORMS, M.C., 206 SVANBERG, 0., 244
VAN DE MEENT, D., 144 VAN HOOF, F., 312, 318 VAN 'T KLOOSTER, H.A., 416 VENTURA, F., 329 VIGERUST, B., 372 VIRKKI, J., 344
WAGGOTT, A., 2 WATTS, C.D., 154 WEGMAN, R.C.C., 62 WETZEL, J., 408 WHITTLE, P., 2
YATES, J., 97
ZOCCOLILLO, L., 116 ZOURARI, M., 121 ZWICK, G., 52
- 452-
