Fresenius J Anal Chem (1995) 351:461-466 Fresenius' Journal of

© Springer-Verlag 1995

Preliminary appraisal of a novel sampling and storage technique for the speciation analysis of lead and mercury in seawater Magnus Johansson, Hfikan Emteborg, B6rjc Glad, Fredrik Reinhoidsson, Douglas C. Baxter Department of Analytical Chemistry, Umezi University, S-901 87 Ume~, Sweden

Received: 24 February 1994/Revised: 6 April 1994

Abstract. A sampling technique suitable for the precon- centration of lead and mercury species from seawater was developed and evaluated in this preliminary study. Sea- water was first pumped through a tubular functional membrane at a flow rate of 2 ml rain- 1 for adjustment of the sample pH. The analyte species were then enriched by solid-phase extraction in a microcolumn containing a resin with immobilized dithiocarbamate (DTC) groups. Optimum recoveries of alkyl- and inorganic lead species were obtained at a pH of about 7, whereas the enrichment efficiencies of alkyl- and inorganic mercury compounds were fairly independent of pH. The buffering membrane was effective in adjusting the pH of the seawater stream from its natural value of 8.1 to 7.1 prior to analyte enrich- ment on the DTC resin. Storage stability of the analyte species on the column was also studied, but found to be poor, indicating that elution from the DTC resin and further sample processing should commence as soon as possible after loading. Suggestions for further develop- ment and improvement of the proposed sampling tech- nique are given.

Introduction

In assessing trace element concentrations in seawater, the need to transport the samples to the laboratory and store them until the analysis can be performed may lead to problems. Confining the ensuing discussion to lead and mercury species, it is apparent that adsorption losses on the sample vessel walls may occur, unless suitable mater- ials and appropriate container pre- treatment techniques are chosen [1-7]. Addition of preservative agents to pre- vent adsorption losses and volatilization of mercury are frequently required as well, increasing the risks for sample contamination. Furthermore, when speciation analysis is to be carried out, such preservatives may alter the distri- bution of chemical forms, [5-7] and irrespective of the storage conditions, certain species exhibit very limited stability in aqueous solutions, such as alkyllead com- pounds [8].

Correspondence to: D.C. Baxter

In an attempt to eliminate the aforementioned difficul- ties in the context of speciation analysis, an alternative sampling technique has been devised. Directly at the sampling site, organic and inorganic lead and mercury species can be simultaneously extracted and preconcen- trated in a flow system incorporating a solid sorbent, containing dithiocarbamate functional groups, installed in a microcolumn [9]. Thus the requirement to preserve, transport and extract the analyte species from large vol- umes of seawater is avoided, and only the microcolumn needs to be sent to the laboratory. However, there could be practical problems in accomplishing deep water samp- ling with the apparatus described. In order to provide a suitable pH for the efficient recovery of lead and mer- cury species on the dithiocarbamate resin, a tubular func- tional membrane for flow buffering [10] is included in the sampling system. This permits the pH of the seawater sample to be maintained at the required value for precon- centration without the addition of buffers, and thus elim- inates one potential source of contamination, of particu- lar importance for the determination of inorganic lead and mercury.

Here, the storage stability of lead and mercury species preconcentrated on the microcolumn, the efficiency of the flow buffering membrane for seawater sampling, and the recovery of target analytes is evaluated. Seawater was selected for this study because: (i) samples collected at sea may have to be stored for extended periods of time prior to analysis; (ii) contamination problems involved in the storage, preservation and manipulation of seawater can be severe; (iii) the concentrations of humic substances, which may complex analyte ions and reduce recoveries on the dithiocarbamate resin [91, are low; and (iv) the buffer- ing membrane requires that the sample has sufficient ionic strength to mediate the pH adjustment process [10]. These considerations made seawater an ideal test matrix for this evaluation of the sampling technique described.

Experimental

In this study, the following species are considered: trimethyllead (M%Pb+), triethyllead (Et3Pb+), diethyl- lead (Etzpb2+), inorganic lead (Pb2+), methylmercury (MeHg+), ethylmercury (EtHg +) and inorganic mercury

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Table 1. Operating conditions for the gas-chromatographic separation of butylated lead and mercury species

Parameter Analyte

Lead Mercury

Column oven Initial column temperature (°C) 30 50 Initial hold time (min) 1 1 Ramp rate ( °C min- 2) 40 40 Final column temperatue (°C) 240 180 Isothermal hold time (min) 1 1 Carrier gas flow rate (rot He min a) 18 18

Injector Sample volume a (gl) 14-16.5 5 6 Injector temperature (°C) 180 180

Detection system QTAAS MIP-AES b

a Manual injection b Column effluent initially vented to prevent solvent entering the cavity, then redirected to plasma after 2.3 min

Recirculatlon

Sample

z

~ Membrane

Buffer

Chelating Column

(Hg2+). Dimethyllead was not available and thus could not be included in this study.

Instrumentation

For the determination of lead species a coupled gas chromatography quartz tube atomic absorption spec- trometry (GC-QTAAS) instrument was used. Separation of the lead species was achieved, following butylation (see Procedure), by GC using a 15 m x 0.53 mm i.d. wide-bore fused-silica capillary column with a 1.5 btm thick non- polar DB-1 stationary phase (J & W Scientific, Rancho Cordova, CA, USA) installed in a Varian Model 3300 gas chromatograph (Palo Alto, CA, USA). A 0.5 m length of 0.25 mm i.d. de-activated, fused-silica transferred the col- umn effluent to the QT atomizer via a heated interface maintained at 240 °C. Operating conditions for the gas chromatograph are listed in Table 1. The QT atomizer consisted of a quartz T-tube (15 cm x 0.7 cm i.d.) mounted on the optical axis of a Perkin-Elmer Model 5000 spec- trometer (Uberlingen, Germany), equipped with a Pb Intensitron hollow cathode lamp as the light source. Lead was detected at an analytical wavelength of 283.3 nm. A flow of H 2 (15.5 mlmin -1) was added to the He flow from the gas chromatograph to improve the atomization efficiency of the lead species, [11, 12] and the temperature of the QT atomizer was set at 1000 °C.

The same type of gas-chromatographic system was used for the separation of butylated mercury species. However, in this case, an atmospheric pressure He micro- wave-induced plasma atomic emission spectrometry (MIP-AES) system was utilized for detection purposes [9]. The He plasma was sustained in a Beenakker TMol0 microwave cavity (AHF Ingenieurb/iro, Tfibingen, Ger- many), and the mercury emission was detected at 253.7 nm using an MPD 850 multichannel spectrometer (Applied Chromatography Systems, Luton, UK). The in- terface was maintained at 150 °C, sufficient for the transfer

Fig. 1. Schematic illustration of the sampling system. The seawater is pumped through the buffering tubular membrane. Regenerant buffer is pumped counter-currently around the outside of the membrane through the annular space between the concentric tubes, and recirculated via a vessel containing 500 ml buffer. The pH adjusted seawater then flows through the chelating column packed with 60 btl DTC resin, where the analytes are enriched

of the relatively volatile butylated mercury species. Fur- ther information on the MIP-AES system is available elsewhere [9] and the gas-chromatographic conditions are given in Table 1.

Reagents

All reagents used were of analytical reagent grade or higher purity. Acids were further purified using an all- quartz sub-boiling distillation system (Heraeus Quarz- schmelze, Hanau, Germany). To reduce reagent blanks resulting from the inorganic forms of lead and mercury, procedures described by Sperling et al. [13] and Em- teborg et al. [9] were followed.

Sampling system

Figure 1 illustrates the design of the sampling system used in this work. The buffering membrane consists of an interpenetrating polymer network (IPN) produced by photo-initiated radical polymerization of 95% (w/w) 2,3- epoxypropyl methaerylate (glycidyl methacrylate; GMA) and 5% (w/w) ethylene dimethacrylate (EDMA) in the pores of a polypropylene tube (Celgard X-20, 0.4 mm i.d., wall thickness 25 gin, porosity 40%, nominal pore size 0.03 lam, length 0.5 m; Hoechst Celanese, Charlotte, NC, USA). Functionalization of the pendant epoxy groups of the polymerized GMA was achieved using pure ethylene

diamine in a poly(tetrafluoroethylene) (PTFE) lined auto- clave at a temperature of 80°C for 24 h. The expected functionalization reaction is

O O O OH

where @ represents the polymer backbone.

The functionalized tubular membrane was installed in the configuration shown in Fig. 1. A peristaltic pump (Ventur Alitea, Stockholm, Sweden) was used to deliver the sample solution through the buffering membrane at a flow rate of 2 ml min- 1. A regenerating buffer solution (0.5moll -1 imidazole, pH adjusted to 6.5 with sub- boiling distilled HNO3) was pumped counter-currently at a flow rate of 4 ml min- 1 using the second channel of the peristaltic pump. The effluent sample solution (of pH 7.1) then passed through a microcolumn (9.7 mmx 2.8ram i.d., volume 60 gl) made from poly(chlorotrifluoroethylene) (Kel-F; Malm6 Fluorocarbon, Malm6, Sweden) with polypropylene frits (Vyon-P; PIAB, Akersberga, Sweden) containing a dithiocarbamate (DTC) resin for enrichment of the lead and mercury species. Further details on the preparation of the buffering membrane [10, 14] and DTC resin [9] may be found elsewhere.

Procedure

Seawater aliquots were spiked with five organometallic species to yield the following concentrations, expressed in terms of lead or mercury in ng l - l : 150 Me3Pb+; 150 Et3Pb+; 150 Et2Pb2+; 8.0 MeHg÷; 8.0 EtHg ÷. The sam- pies were allowed to equilibrate for 20 24 h in the dark at 4°C. Samples (100 ml) were then pumped through the DTC resin microcolumn via the buffering membrane tube. Regenerant buffer was pumped conter-currently around the outside of the membrane through the annular space between the concentric tubes, to maintain a con- stant pH in the effluent flow. The analyte species were then enriched from the pH adjusted seawater on the DTC resin. In total twelve columns were loaded with the analyte species, five being stored at room temperature and five at a temperature of 4 °C. After loading, the columns were tightly capped still wet (to avoid shrinkage of the resin) and placed in acid-washed glass containers wrap- ped in aluminium foil. Directly after sampling two of the columns were separately installed in a flow injection sys- tem (Perkin-Elmer FIAS-2000) and the analytes eluted. Lead species were selectively eluted from the columns using 1 ml of a 0.066 tool 1-1 HC1 solution at a flow rate of 3.3 ml min -1. Mercury species were then eluted using 0.9 ml of 5% (w/w) thiourea in 0.1 moll -1 HNO3 (pH 1) at the same flow rate. Initial experiments showed com- plete separation of the lead and mercury species using this elution protocol. Additional replicates were made in an initial study of the pH dependence of the analyte recoveries.

The eluent solutions and appropriate aqueous stan- dards were then adjusted to a pH value of ~ 9 with addition of a suitable volume of pH 9.0 borate buffer and

463

1 moll-1 NaOH. Then, 1 ml of 0.5 moll 1 sodium die- thyldithiocarbamate solution was added, lead and mer- cury species being extracted into 1 ml hexane and 0.5 ml toluene, respectively. After separation of the organic phase (0.7 ml hexane or 0.4 ml toluene recovered), the analytes, including the inorganic forms originally present in the seawater, were then butylated using a Grignard reagent (0.2 ml of 2.0 mol 1-1 butylmagnesium chloride in tetrahydrofuran). To the butylated extracts, 0.5 ml of 0.6 moll 1 HC1 was added to eliminate excess Grignard reagent, and then the organic phase was transferred into screw- capped glass vials using Pasteur pipettes.

Seawater sample collection. Sub-surface seawater was col- lected into four 1 1 acid-washed PTFE containers (Nal- gene, Rochester, NY, USA) from Saltholmen, Gothen- burg on the west coast of Sweden (salinity = 18.9%o, pH 8.1) in October 1993. After sampling the containers were sealed and placed into acid-washed plastic bags. No pre- servative agents were added, and the samples were stored in the dark at 4°C until subsamples were spiked with alkyllead and alkylmercury species in the laboratory, four days later.

Results and discussion

Extraction efficiency as a function of pH

Although solid-phase extraction has been widely em- ployed for the preconcentration of organic pollutants [15] and trace metals [2, 13, 16] from water samples, application to organometallic species has so far been rather limited, as discussed by Szpunar-Lobinska et al. [17]. These authors obtained essentially pH independent extraction efficiencies for phenyl- and butyltin species over the pH range 2-12 using a Cls microcolumn [173. Recoveries of tri- and disubstituted organotins were > 95%, whereas the monosubstituted species were slight-

ly less efficiently retained by Cls (80-90%), due to their increased polarity and ionic character.

As the alkyllead [18] and, in particular, the alkylmer- cury species considered here are more polar than butyl- and phenyltin compounds, better recoveries would be expected using a cation exchange resin. For this purpose the DTC resin synthesized by Emteborg et al. [9] was tested, as previous results have shown the efficiency of this material for the recovery of mercury species from sea- water. The effect of pH on the retention of lead and mercury species was investigated using seawater samples spiked with Me3Pb +, Et3Pb +, Et2Pb 2+, MeHg + and EtHg +, and adjusting the pH with HNO3 or NaOH. After selective elution to separate the lead and mercury compounds, extraction into an appropriate organic sol- vent and butylation, the recoveries were determined and the results are shown in Fig. 2. The pH range covered is representative of conditions prevaifing in both estuarine and open ocean seawaters [19]. Furthermore, studies on the liquid-liquid extraction of ionic alkyllead compounds as their diethyldithiocarbamate complexes [20-23] generally indicate an optimum recovery at pH values from 8 to 9.

As seen in Fig. 2, the optimum pH for recovery of lead species on the DTC resin is around 7. No explanation for

464

8

1 0 0

90

8 0

7 0

60

50

40 I I I ,

6 7 8 9 pH

( a )

10

100 -

90

80

8 70-

6 0 -

( b )

5 0 I I ] I

5 6 7 8 9 10 b pH

Fig. 2a, K Retention efficiencies of a Me3Pb + (e), Et3Pb + (11), EtzPb 2+ (A) and Pb 2+ (C)); and b MeHg + (O), EtHg + (11) and Hg a+ (A) from seawater on the DTC column as a function of pH. Sampling rate 2 ml min 1. Error bars are for one standard deviation (n = 3 - 5)

the differences in the pH dependences between solid- phase and liquid-liquid extraction [20-23] can be offered at present. The results for mercury species show no great pH dependence, in agreement with the work of Minagawa and Takizawa [24] and Lansens et al. [25] concerning the preconcentration of Hg 2+ and MeHg + from natural water samples.

The recoveries of alkyllead compounds are generally poorer than those obtained by other workers using liquid-liquid extraction of diethyldithiocarbamate com- plexes [20-23]. This may be partly due to the co-extrac- tion of Pb 2+, which Chakrabarti et al. [20] found to inhibit the recovery of alkyllead species. This problem was overcome by masking Pb 2+ with ethylenediaminetetra- acetic acid. Obviously, such a solution would be impracti- cal using the present sampling technique, due to sample dilution effects and the potential for contamination.

As the recoveries of mercury species were also some- what lower than previously obtained [9], an alternative explanation may be appropriate. The seawater was sam- pled from the west coast of Sweden, at a point close to the River G6ta outlet. River water contains considerable amounts of humic substances which can complex ions in solutions [26]. Humic-rich waters have been shown to severely depress the recoveries of mercury species on the DTC resin, [9] and impair the liquid-liquid extraction [27] of MeHg +. Thus sufficient organic material may be

present in the sub-surface seawater sample used to at least partly complex the added lead and mercury compounds. Due to the short contact time of the flowing sample with the DTC resin (450 ms), complete transfer of the analytes from humic complexes to the chelating groups cannot be expected.

Effect of sampling rate

Experiments were performed without the buffering mem- brane, at the optimum pH for recovery of ionic alkyllead species (Fig. 2), to study the effect of the sample flow rate on the retention efficiency of the DTC resin. No effect was observed for MeHg +, EtHg + and Hg ~+ a t sampling rates up to 8 ml min- 1. For EtzPb 2 + and Pb 2 +. the recoveries were fairly constant at flow rates up to 5 ml rain- 1. How- ever, the recoveries of both trialkyllead species decreased rapidly, by about 4% per 1 ml min- 1 increase in sampling rate, above 2 ml rain- 1.

The independence of the recoveries of mercury species on the flow rate can be attributed to the extremely high affinity mercury exhibits towards the dithiocarbamate groups on the chelating resin [9, 24, 25, 28]. Equally efficient retention of EtzPb 2 + and Pb 2 + up to flow rates of 5 ml rain- 1 can probably be ascribed to their greater polarities and more ionic character compared with the trialkyllead species. Consequently, a sampling rate of 2 mlmin 1 was chosen for all further studies, both to maximise the recoveries of Me3Pb + and Et3Pb +, and to avoid rupturing the buffering membrane.

Performance of the flow buffering tubular membrane

The amine groups on the functional membrane used here have a pKa value of ~ 9. By using a regenerant buffer solution with a similar pH value, sampled solutions vary- ing in pH from 3 to 11 can be efficiently buffered to pH 9 [10]. However the optimum pH for recovery of lead species on the DTC column was 7. To obtain a pH of about 7 in the seawater sample (pH 8.1) a more acidic buffer was required, and imidazole (pKa~ 6.9) was se- lected. The pH of the imidazole buffer was adjusted to 6.5 with HNO3 and the pH of the seawater eluting from the membrane was continuously monitored using a calib- rated pH electrode mounted in a low dead-volume flow- through cell. After a stabilization period of 2 min, the eluent pH remained constant at 7.1 + 0.02 for the next hour. Similar results were obtained for a seawater sample adjusted to pH 7.8. Thus the buffering membrane system can conveniently be applied to seawater samples in the pH range from 7.8 to 8.1 without altering the buffer composition. However, lower pH samples would require a change in the buffer composition, which could be incon- venient. It would therefore be desirable to synthesize a membrane possessing functional groups with a pKa of ~ 7 in order to buffer samples in a broader pH range without changing the buffer [10].

At the pH of the imidazole buffer, essentially all of the amine groups in the membrane will be protonated, and buffering of the influent seawater sample is a result of transfer of an ion pair (H + N O ; ) through the membrane.

Table 2. Stability of alkyllead and alkylmercury species on the DTC column Storage Recovery a (%)

time (d) M%Pb + Et 3Pb + Et2Pb 2 + MeHg + EtHg +

20°C 4°C 20°C 4°C 20°C 4°C 20°C 4°C 20°C 4°C

2 99 96 76 83 70 79 82 86 59 64 6 75 80 48 42 < 5 b < 5 _c 90 - 71 13 58 75 25 32 < 5 < 5 80 91 46 72 90 36 40 24 37 < 5 < 5 < 5 < 5 < 5 < 5

465

a Analyte concentration recovered after a given storage time relative to the concentration recovered directly after sampling b Analyte concentration recovered below the detection limit corresponding to about 5% of the initial concentration c No peaks observed in the GC-MIP-AES chromatograms

The positively charged amine functionalities will effectively exclude transport of the cationic analyte species from the sample stream to the regenerant buffer. This is obviously an important consideration in avoiding analyte losses through the membrane. Similarly, the risk for contamination of the sample due to impurities in the buffer is eliminated.

Species stability on the DTC resin

After enrichment on the chelating resin it may be advant- ageous to be able to store the microcolumn for extended time periods, for example when sampling is performed at sea and the necessary analytical instrumentation is not available directly. Therefore the stability of the analyte species considered here was studied over three months. The results are presented in Table 2. Five of the samples were stored at room temperature (~21 °C) and five in a refrigerator at 4 °C. Pairs of columns were eluted, ex- tracted, derivatized and analyte concentrations deter- mined following 2, 6 and 13 days storage, the remainder being processed after 90 days. Concentrations determined for duplicate columns processed after 90 days using both storage modes agreed to within about 15% and 8% for alkyllead and alkylmercury species, respectively.

It can be seen from Table 2 that storage at 4°C generally provides better species stabilities. Furthermore, the ethylated forms of both lead and mercury decompose more rapidly than the corresponding methyl analogues, irrespective of the storage conditions, in agreement with previous observations [8]. As the organometallic species decomposed, corresponding increases in the inorganic metal peaks were seen. From Table 2 it is obvious that the recoveries of MeHg + and EtHg-- increased over the inter- val between 2 and 13 days storage at 4°C. This may simply be a result of uncertainties in the analyses, but the possibility of transalkylation reactions cannot be ruled out. Indeed, Ebinghaus and Wilken [29] have provided experimental evidence for the transmethylation of Hg 2+ in the presence of MeBPb ÷. The much higher concentra- tions of ionic alkyllead compounds and their decomposi- tion over the time interval concerned might thus favour such transalkylation reactions.

Obviously, rigorous statistical verification of changes in analyte concentrations with storage time and mode

would require additional replicates to be performed. However, this seems unnecessary in the present case be- cause of the observed losses of species with time. Possibili- ties to improve the storage stabilities are being studied and are mentioned below (see Conclusion).

Overall performance

A summary of the overall performance of the proposed sampling system and analytical procedures used is given in Table 3. Analyte losses are divided into two categories, those occurring during the actual sampling phase (solid- phase extraction) and those resulting during the work-up following elution. No losses were observed in the buffer- ing membrane tube or in the analyte elution step under the conditions described in Procedures. The major losses of two species, EtzPb 2+ and Pb 2+, occurred during the processing of the analytes eluted from the DTC column. This is probably due to inefficient butylation of these species, and Radojevic et al. [-21] have suggested that propylation results in much better recoveries. Propylation has been the derivatization method of choice in more recent studies of lead speciation [22, 23].

Clearly, the relative detection limits for all species must be improved before this analytical method could be applied to seawater samples at natural concentration levels (although no such data are yet available for alkyl- lead species). This could be achieved by processing larger sample volumes, a time-consuming stage with the low flow rate (2 ml rain- 1) required in the present configura- tion. In the case of lead species, detection limits could be improved by over two orders of magnitude using the GC-MIP-AES system described by Lobinski and co- workers [22, 23]. Improved relative detection limits could be achieved for the mercury species by injecting larger processed sample volumes on the GC-MIP-AES system [-9]. Only two species could be detected in unspiked seawater samples, Pb 2 + at a concentration of 0.5 ~tgl-1 and Hg 2 ÷ at 2 ng 1-1

Conclusion

This preliminary appraisal of the proposed sampling tech- nique has demonstrated several promising features for the

466

Table 3. Overall performance of the sampling system and analytical pro- cedures

Analyte Analyte losses (%)

Solid-phase extraction a

Estimated detection limits

Solvent extraction Absolute ~ Relative d and derivatization b (pg) (ng 1-1)

Me3Pb + 7 10 Et3Pb + 9 10 Et2Pb 2+ 23 40 Pb 2+ 10 60 MeHg ÷ 24 < 5 EtHg + 41 < 5 Hg 2 ÷ 23 < 5

10 8 12 10 15 25 20 42 0.4 0.5 0.4 0.6 0.4 0.5

a For seawater adjusted to pH 7.1 using the buffering tubular membrane b On the basis of the analyte concentrations eluted from the DTC resin c Calculated as the analyte mass in the solution injected onto the GC column (i.e. corrected for losses and blanks) yielding a signal-to-noise ratio of 1 a Assuming a 100 ml seawater sample and injection volumes reported in Table 1

speciation analysis of lead and mercury in seawater. How- ever, a number of impor tan t problems have emerged which require further careful study:

(1) The effect of dissolved organic mat ter on the recove- ries of the analytes should be investigated. (2) Efforts should be made to increase the sampling flow to improve the rate of preconcentrat ion. In the present system this is limited by the buffering membrane tube and the small volume of the chelating column. Higher flows could be employed using a reactor designed for flat sheet membranes, similar to the configurat ion described by Stillian [30] and a longer packed-bed chelating column. (3) The stabilities of the organometal l ic species retained on the D T C resin were disappointing, indicating that the sample processing should proceed immediately after col- umn loading. However, it m ay be possible to minimize the rate of analyte decomposi t ion by drying the columns prior to freezing, or rinsing the resin with ethanol or an azide solution to prevent biodegradat ion. (4) Applicat ion of the sampling system to other organo- metallic and metallic species should be worthwhile, as m a n y other analytes exhibit high affinities towards d i th iocarbamate groups [28], including butylt in com- pounds [31]. Addit ional sample matrices, such as rain- water, which fulfil the requirements discussed in the Intro- duction, should also be tested.

Acknowledgements.This work was supported by the Centre for Environ- mental Research in Ume~i, the Swedish Natural Sciences Research Council and the Swedish Environmental Protection Board.

References

1. Griepink B (1984) Pure Appl Chem 56:1477 2. Sturgeon R, Berman SS (1987) CRC Crit Rev Anal Chem

18 : 209 3. Betti M, Papoff P (1988) CRC Crit Rev Anal Chem 19 : 271 4. Hamlin SN (1989) Water Resour Bull 25:255 5. Barley GE (1989) Collection preparation, and storage of

samples for speciation analysis. In: Batley GE (ed), Trace

element speciation: analytical methods and problems CRC Press, Boca Raton, pp 1 24

6. Leermakers M, Lansens P, Baeyens W (1990) Fresenius J Anal Chem 336 : 655

7. Ahmed R, Stoeppler M (1986) Analyst 111 : 1371 8. Van Cleuvenbergen R, Dirkx W, Quevauviller P, Adams

F (1992) Int J Environ Anal Chem 47 : 21 9. Emteborg H, Baxter DC, Frech W (1993) Analyst

118 : 1007 10. Glad B, Irgum K (1994) Anal Chim Acta (submitted) 11. Forsyth DS (1987) Anal Chem 59:1742 12. Forsyth DS, Marshall WD (1985) Anal Chem 57 : 1299 13. Sperling M, Yin X, Welz B (1991) J Anal At Spectrom

6:295 14. Glad B, Irgum K (1994) J Membrane Sci (submitted) 15. MacCarthy P, Klusman RW, Cowling SW, Rice JA (1993)

Anal Chem 65 : 244R 16. Fang Z (1991) Spectrochim Acta Rev 14:235 17. Szpunar-Lobinska J, Ceulemans M, Lobinski R, Adams FC

(1993) Anal Chim Acta 278:99 18. Hewitt CN, Metcalfe PJ, Street RA (1991) Water Res 25 : 91 19. Stumm W, Morgan JJ (1981) Aquatic Chemistry, 2 edn.

Wiley, New York 20. Chakraborti D, De Jonghe WRA, Van Mol WE, Van

Cleuvenbergen R J A, Adams FC (1984) Anal Chem 56 : 2692 21. Radojevic M, Allan A, Rapsomanikis S, Harrison RM

(1986) Anal Chem 58 : 658 22. Lobinski R, Adams FC (1992) Anal Chim Acta 262:285 23. Lobinski R, Boutron CF, Candelone J-P, Hong S, Szpunar-

Lobinska J, Adams FC (1993) Anal Chem 65:2510 24. Minagawa K, Takizawa Y (1980) Anal Chim Acta

115:103 25. Lansens P, Meuleman C, Leermaerkers M, Baeyens

W (1990) Anal Chim Acta, 234:417 26. Boggs S, Livermore DG, Seitz MG (1985) Rev Macromolec

Chem Phys C25 : 599 27. Horvat M, Liang L, Bloom NS (1993) Anal Chim Acta

282:153 28. Hulanicki A (1967) Talanta 14:1371 29. Ebinghaus R, Wilken RD (1993) Appl Organomet Chem

7 : 127 30. Stillian J (1985) LC Magazine, 3 : 802, 806, 808, 812 31. Dirkx WMR, Van Mol WE, Van Cleuvenbergen RJA,

Adams FC (1989) Fresenius Z Anal Chem 335:769