Toxicological Profiles of Chemicals of Concern at Mapua finalMapua, a small coastal village located...

Toxicological Profiles of Chemicals of Concern at Mapua A report commissioned by the Nelson Marlborough District Health Board Leo J. Schep, PhD ToxInform Ltd 13 Quarry Rd Mosgiel NEW ZEALAND 17th August 2009

1

TABLE OF CONTENTS

INTRODUCTION.......................................................................................................... 7

1.0 ATRAZINE.......................................................................................................... 9

1.1 Physical Chemical Properties.............................................................................. 9 1.1.1 Basic Properties.................................................................................. 9 1.1.2 Brief notes.......................................................................................... 9 1.1.3 Uses ................................................................................................... 9

1.2 Routes of Exposure............................................................................................ 10 1.2.1. Ingestion .......................................................................................... 10 1.2.2 Skin.................................................................................................. 10 1.2.3 Eye................................................................................................... 10 1.2.4 Inhalation ......................................................................................... 10

1.3 Toxicokinetics .................................................................................................... 10 1.3.1 Absorption ....................................................................................... 10 1.3.2 Distribution ...................................................................................... 11 1.3.3 Metabolism ...................................................................................... 11 1.3.4 Elimination....................................................................................... 11

1.4 Mechanism of Action within the Body.............................................................. 11 1.4.1 Oestrus Cycle Disruption.................................................................. 11

1.5 Summary of Clinical Signs and Symptoms Following Toxic Exposure........... 12 1.5.1 Acute Signs and Symptoms .............................................................. 12 1.5.2 Chronic Signs and Symptoms........................................................... 13

1.6 Carcinogenicity.................................................................................................. 14 1.6.1 Human Studies ................................................................................. 14 1.6.2 Animal Studies................................................................................. 15

1.7 Mutagenicity ...................................................................................................... 15

1.8 Developmental Effects and Teratogenicity....................................................... 15 1.8.1 Human studies.................................................................................. 15 1.8.2 Animal Studies................................................................................. 16

1.9 Environmental fate (metabolism and accumulation) and subsequent ecotoxicity .......................................................................................................... 16

1.9.1 Environmental Levels....................................................................... 16 1.9.2 Environmental Fate .......................................................................... 17 1.9.3 Ecological Degradation .................................................................... 17

1.10 Dose Response Information............................................................................... 18 1.10.1 Doses Necessary to Cause Acute and Chronic Toxicity .................... 18 1.10.2 Summary of Case Reports Following Human Toxicological

Exposure ......................................................................................... 19 1.10.3 Measures of Mammalian Toxicity .................................................... 19 1.10.4 Guidance Values .............................................................................. 20

2

1.11 A summary of Biological Assays that are Available to Test for Evidence of these Chemicals ............................................................................................. 20

1.12 Summary............................................................................................................ 21

1.13 References.......................................................................................................... 22

2.0 PENTACHLOROPHENOL ............................................................................. 31

2.1 Physical Chemical Properties............................................................................ 31 2.1.1 Basic Properties................................................................................ 31 2.1.2 Brief Notes....................................................................................... 31 2.1.3 Uses ................................................................................................. 32

2.2 Routes of Exposure............................................................................................ 32

2.2.1 Ingestion............................................................................................................. 32 2.2.2 Skin.................................................................................................. 32 2.2.3 Eye................................................................................................... 33 2.2.4 Inhalation ......................................................................................... 33


2.4 Mechanism of Action within the Body.............................................................. 35


2.6 Carcinogenicity.................................................................................................. 38 2.6.1 Human Studies ................................................................................. 38 2.6.2 Animal Studies................................................................................. 38

2.7 Mutagenicity ...................................................................................................... 38

2.8 Developmental Effects and Teratogenicity....................................................... 39 2.8.1 Human Studies ................................................................................. 39 2.8.2 Animal Studies................................................................................. 39

2.9 Environmental Fate (Metabolism and Accumulation) and Subsequent Ecotoxicity ......................................................................................................... 39


2.10 Dose Response Information............................................................................... 41 2.10.1 Doses Necessary to Cause Acute and Chronic Toxicity .................... 41 2.10.2 Summary of Case Reports Following Toxic Exposure ...................... 43 2.10.3 Measures of Mammalian Toxicity .................................................... 44 2.10.4 Guidance Values .............................................................................. 45

2.11 A Summary of Biological Assays That are Available to Test for Evidence of These Chemicals ............................................................................ 46

2.12 Summary............................................................................................................ 47

3

2.13 References.......................................................................................................... 48

3.0 POLYCHLORINATED BIPHENYLS............................................................. 58

3.1 Physical Chemical Properties............................................................................ 58 3.1.1 Basic Properties................................................................................ 58 3.1.2 Brief Notes....................................................................................... 58 3.1.3 Uses ................................................................................................. 59

3.2 Routes of Exposure............................................................................................ 60 3.2.1 Skin.................................................................................................. 60 3.2.2 Ingestion .......................................................................................... 60 3.2.3 Inhalation ......................................................................................... 60


3.4 Mechanism of Action within the Body.............................................................. 62


3.6 Carcinogenicity.................................................................................................. 69

3.7 Mutagenicity ...................................................................................................... 70

3.8 Developmental Effects and Teratogenicity....................................................... 70 3.8.1 Humans............................................................................................ 70 3.8.2 Animal Studies................................................................................. 71

3.9 Environmental Fate (Metabolism and Accumulation) and Subsequent Ecotoxicity ......................................................................................................... 72


3.10 Dose Response Information............................................................................... 73 3.10.1 Doses Necessary to Cause Human Acute and Chronic Toxicity ........ 73 3.10.2 Summary Case Reports Following Toxicological Exposure.............. 75 3.10.3 Measures of Mammalian Toxicity .................................................... 76 3.10.4 Guidance Values .............................................................................. 76

3.11 A Summary of Biological assays That are Available to Test for Evidence of These Chemicals in Biological Samples ........................................................ 77

3.12 Summary............................................................................................................ 78

3.13 References.......................................................................................................... 79

4.0 DIOXINS ........................................................................................................... 93

4.1 Physical Chemical Properties............................................................................ 93 4.1.1 Basic Properties................................................................................ 93 4.1.2 Brief notes........................................................................................ 93 4.1.3 Uses ................................................................................................. 95

4

4.2 Routes of Exposure............................................................................................ 96 4.2.1 Ingestion .......................................................................................... 96 4.2.2 Skin.................................................................................................. 97 4.2.3 Eye................................................................................................... 97 4.2.4 Inhalation ......................................................................................... 97


4.4 Mechanism of Action within the Body............................................................ 101

4.5 Summary of Clinical Signs and Symptoms Following Toxic Exposure......... 102 4.5.1 Signs and Symptoms ...................................................................... 102

4.6 Carcinogenicity................................................................................................ 103 4.6.1 Human Studies ............................................................................... 103 4.6.2 Animal Studies............................................................................... 104

4.7 Mutagenicity .................................................................................................... 104

4.8 Developmental Effects and Teratogenicity..................................................... 104 4.8.1 Human Studies ............................................................................... 104 4.8.2 Animal Studies............................................................................... 105

4.9 Environmental Fate (Metabolism and Accumulation) and Subsequent Ecotoxicity ....................................................................................................... 105

4.9.1 Environmental Levels..................................................................... 105 4.9.2 Environmental Fate ........................................................................ 106 4.9.3 Ecological Degradation .................................................................. 107

4.10 Dose Response Information............................................................................. 107 4.10.1 Doses Necessary to Cause Toxicity ................................................ 107 4.10.2 Summary of Case Reports Following Toxicological Exposure........ 109 4.10.3 Measures of Mammalian Toxicity .................................................. 109 4.10.4 Guidance Values ............................................................................ 110

4.11 A Summary of Biological Assays that are Available to Test for Evidence of These Chemicals in Biological Samples (e.g. blood, urine) ........................ 112

4.12 Summary.......................................................................................................... 113

4.13 References........................................................................................................ 114

5.0 TOLERABLE DAILY INTAKES FOR A RANGE OF ORGANOCHLORINE PESTICIDES AS DEFINED BY VARIOUS INTERNATIONAL BODIES......................................................................... 128

5.1 References........................................................................................................ 134

6.0 APPENDICES ................................................................................................. 137

5

LIST OF TABLES Table 1.1 Physical Chemical Properties of Atrazine .................................................. 9 Table 1.2 Table Summaries of LD50 single dose values for atrazine28 .................... 19 Table 2.1 Physical Chemical Properties of Pentachlorophenol and

pentachlorophenate.................................................................................. 31 Table 2.2 Summary of measured serum PCP concentrations of various sub-

populations after possible chronic exposure to elevated levels of PCP143..................................................................................................... 43

Table 2.3 Post-mortem analysis of body fluids and tissue samples as evidence of PCP intoxication21............................................................................... 44

Table 2.4 Summary of single dose LD50 values for PCP10....................................... 44 Table 2.5 Tolerable daily intakes for pentachlorophenol, as defined by various

regulatory authorities............................................................................... 45 Table 2.6 A summary of established MRL levels for PCP by ingestion, as

recommended ATSDR ............................................................................ 46 Table 3.1 Physical Chemical Properties of 8 Major Aroclor products..................... 58 Table 3.2 Initial signs and symptoms reported in Yusho patients following

PCB exposure.60...................................................................................... 63 Table 3.3 A summary of various doses of PCB ingested and their respective

severity of symptoms suffered by 146 Yusho patients132.......................... 74 Table 3.4 Oral toxicity of seven Aroclor mixtures148............................................... 76 Table 3.5 Reference doses for Aroclor 1016 and 1254 as defined by the

EPA223..................................................................................................... 76 Table 4.1 Fundamental physical chemical properties of the various dioxin

homologues............................................................................................. 93 Table 4.2 Toxic Equivalent Factors for the toxic dioxins and PCBs. ....................... 94 Table 4.3 Post-mortem TCDD concentrations (ng/g) in various tissues from a

55 year old woman who died of pancreatic cancer 7 months after exposure at Seveso .................................................................................. 99

Table 4.4 Dose weighted estimated first-order elimination half-lives (years) for the 17 toxic dioxin congeners, as reported by Flesch-Janys and colleagues (1996)52 and Van der Molen and colleagues (2000)53......... 100

Table 4.5 Summary of LOAEL responses, animal body burdens, and the calculated estimated daily intakes (EDI) of TCDD in humans88............. 108

Table 4.6 Summaries of international assessments of tolerable daily intakes of dioxins .................................................................................................. 111

Table 5.1 Tolerable daily intake values for aldrinf, as defined by various regulatory authorities............................................................................. 129

Table 5.2 Tolerable daily intake values for dieldrinf, as defined by various regulatory authorities............................................................................. 130

Table 5.3 Tolerable daily intake values for lindane, as defined by various regulatory authorities............................................................................. 130

Table 5.4 Tolerable daily intake values for chlordane, as defined by various regulatory authorities............................................................................. 131

Table 5.5 Tolerable daily intake values for DDT/DDE/DDDe, as defined by various regulatory authorities ................................................................ 131

Table 5.6 Tolerable daily intake values for heptachlor, as defined by various regulatory authorities............................................................................. 132

6

Table 5.7 Tolerable daily intake values for heptachlor epoxidec, as defined by various regulatory authorities ................................................................ 132

Table 5.8 Tolerable daily intake values for hexachlorobenzene, as defined by various regulatory authorities ................................................................ 132

Table 5.9 Summary and recommendations of the lowest guidance values for the respective organochlorines............................................................... 133

LIST OF FIGURES Figure 1.1 Proposed pathways for human and rodent biotransformation of

atrazine.17, 20............................................................................................ 11 Figure 1.2 The neuroendocrine control of the gonadal function. ............................... 12 Figure 2.1 Brief summary of the hepatic oxidation and conjugation of PCP ............. 34 Figure 3.1 The structural formula of PCBs; positions for attachment of either

the hydrogen or chlorine molecule can occur at 2’,3’,4’,5’,6’ and 2,3,4,5,6. Substitution with chlorine on the 2 (also 2’,6,6’) position is termed ortho, at position 3 (also 3’, 5, 5’) meta and position 4 (also 4’) is defined as para. ..................................................................... 59

Figure 4.1 The structural formule for (a) Dibenzo-para-dioxin and (b) dibenzofuran. .......................................................................................... 93

Figure 4.2 The structural formula of 2,3,7,8-tetra CDD. ........................................... 94 Figure 4.3 The (relative) formation of dioxins; each value is normalized to the

400oC case. Reproduced with permission.4 ............................................. 95 Figure 4.4 Reported estimated daily exposure to PCDDs and PCDFs, expressed

in TEQ pg/kg body weight, from the United States,12 Netherlands,15 the United Kingdom16 and Germany.17 Reproduced with permission.18 ........................................................................................... 97

Figure 4.5 Mean lipid adjusted TCDD values from populations (ppt) sampled in Europe and North America from 1970 to 2000. Reproduced with permission.54 ......................................................................................... 101

7

Introduction Mapua, a small coastal village located between Nelson and Motueka at the mouth of the Waimea Estuary, was from 1932 to 1988 the site of the Fruit Growers Chemical Company factory and their private landfill. As a result, a range of pesticide pollutants and heavy metals contaminated the soil and surrounding sediments. In 2004, remediation of the contaminated soil was undertaken and finished in 2007. Consequent to this work, concerns were raised regarding the risks to public health from the exposure from a number of these pesticides and other contaminants resulting from the remediation process. Consequently, the Nelson Marlborough District Health Board requested a toxicological profile of a range of chemicals that may still be present at the site or may have been released from the site during remediation; these may pose a risk to residents in an adjoining suburb. Requested profiles were divided into three groups: Group one. Comprehensive chemical profiles of chemicals cited in the PCE report: pentachlorophenol, polychlorinated biphenyls, and atrazine. These profiles include the following sections, summarising information published in the peer-reviewed literature:

Physicochemical properties Routes of exposure (skin, ingestion, inhalation) Toxicokinetics (absorption, distribution, metabolism, elimination) Mechanism of action within the body Summary of clinical signs and symptoms following toxic exposure (human and animal: acute and chronic) Carcinogenicity Mutagenicity Developmental Effects and Teratogenicity Environmental fate (ecological metabolism, accumulation and degradation) and subsequent ecotoxicity Dose response Information:

Doses necessary to cause acute and chronic toxicity Summary of case reports following toxicological exposure Measures of mammalian toxicity Defined exposure limits

A summary of biological assays available to test for evidence of these chemicals in biological samples

Group two. A less detailed toxicological review on dioxins; focusing on toxicity and with summaries of the other categories. Group three: Provide tabulated summaries of tolerable daily intakes, as defined by various regulatory authorities, comment and review their values with those reported in Craig Stevenson’s THI report and sunmarise if Craig’s values are acceptable for determining the respect THIs. This list comprises the organochlorines DDT/DDE, dieldrin, lindane and aldrin, chlordane, heptachlor, heptachlor epoxide and hexachlorobenzene.

8

For the purpose of these reports, unless stated otherwise, an acute exposure is defined as exposure to a chemical for duration of 14 days or less, a sub-acute exposure for one month or less, subchronic for 1 to 3 months and a chronic exposure for more than 3 months. Unless stated otherwise, the term TDI is used throughout the text except when citing research studies or names of organisations, where the original terminology was retained. (eg RfD) One limitation of the study was the accessibility of all available references. Almost all of the references were read by the author, but a small number studies such as from old government or company reports, for example, were unattainable and information obtained and summarised were from a secondary source. In such instances the cited reference will be flagged by the author.

9

1.0 Atrazine

1.1 Physical Chemical Properties

1.1.1 Basic Properties

Table 1.1 Physical Chemical Properties of Atrazine CAS number: 1912-24-9 Synonyms: Aatram; Aatrex ; Agro Atrazine 500, Atraflo; Atragranz; Atrasol;

Atranex; Atratol; Gesaprim; Primatol; Crisazine; Chemgro atrazine; Flowable atrazine; Nu-trazine

Chemical class: Triazine herbicide Chemical name: 6-chloro-N-ethyl-N'-(1-methylethyl)-triazine-2,4-diamine

(IUPAC) Empirical formula: C8H14ClN5 Chemical structure:

Molecular weight: 215.69 Melting point: 175-177 oC Vapour pressure: 3.0 x 10-7 mmHg at 20 oC Water solubility: 30 mg/L at 20 oC Henry's law constant: 2.96 x 10-9 atm-m3/mol (at 25 oC) Partition Coefficient (OW): 2.61 pKa: 1.7

1.1.2 Brief notes Atrazine consists of a triazine ring (a heterocyclic ring with three carbons replaced by nitrogen) with two attached amine groups and a chlorine atom. It is poorly soluble in water, but moderately soluble in a non-polar solvent (eg solubility in chloroform is 52,000 mg/L). It remains stable in slightly acidic, alkaline or neutral media but is hydrolysed by alkali or mineral acids at higher temperatures.

1.1.3 Uses Atrazine is a selective post-emergence herbicide that has been used commercially since 1958. It is one of the most widely used herbicides in the world, with approximately 70,000 and 90,000 tons being applied each year.1 Despite being banned by the European Union in 2003, because of its “ubiquitous and unpreventable water contamination”, it is still used in North America.2 Furthermore, it is still licensed for use in New Zealand to control some annual grasses and most broad leaf weeds.3 Atrazine is spread on crops and croplands as either a powder, liquid or in granular form. To be active, it must dissolve in water enabling it to enter the plant through the

10

roots or leaves and thereby inhibit the Hill reaction,4 which is important in the process of photosynthesis.

1.2 Routes of Exposure

1.2.1. Ingestion A recent case report describes the ingestion of atrazine that was co-ingested with ethylene glycol, formaldehyde.5 Subsequent elevated blood levels of atrazine suggest this pesticide is readily absorbed across the intestinal tract. There is no evidence, of injury to the gastrointestinal mucosa following atrazine exposure. In vitro evidence suggests atrazine does not accumulate in enterocytes but is readily absorbed across the cells.6

1.2.2 Skin There is sufficient evidence to suggest that atrazine can enter systemic circulation through the absorption by intact skin.7 Research suggested epithelial cells were metabolising atrazine. Patch tests on human volunteers have also demonstrated skin absorption by measuring detectable levels of atrazine metabolites in their urine.8 In contrast to other studies (see section on metabolism), the parent drug could not be detected. Workers exposed to atrazine via skin contact have also demonstrated dermal absorption.9, 10 In rats, investigations have demonstrated that younger animals have greater absorption.11 Age differences could not be correlated with skin thickness.

1.2.3 Eye Exposure to the eye has resulted in ocular irritation and slight abrasion to the cornea, with complete resolution of symptoms.12 To the knowledge of the author, there is no evidence of entry to systemic circulation via this route.

1.2.4 Inhalation Men engaged in weed spraying had evidence of atrazine metabolites in their urine; it was implied they had inhaled the vapour in the exercise of their employment.13 To the knowledge of the author, there are no other published reports examining the inhalation of atrazine.

1.3 Toxicokinetics

1.3.1 Absorption Atrazine readily absorbs across the intestinal tract following ingestion.5, 14, 15 Following the ingestion of atrazine, ~57% of the dose has been recovered in the urine of rats.15 There is sufficient evidence in both human and animal studies to suggest skin absorption also occurs, though bioavailability is lower than for oral ingestion.7-11 Only 0.3–4.4% of a dermal administered dose was recovered in urine.8 Approximately 16 % of an applied dose of atrazine to human skin has been reported to partition through the skin using an in vitro diffusion system.7

11

1.3.2 Distribution In the rat, atrazine will distribute to most organs of the body, with the greatest concentration in the liver and kidneys.16 The time to achieve maximum plasma concentrations following ingestion is approximately 8 to 10 hours.

1.3.3 Metabolism Following ingestion, evidence in humans and animal studies suggest the liver metabolises atrazine by phase one dealkylation followed by phase two glutathione conjugation with the subsequent predominant formation of mercapturic acids in the kidneys (see Figure 1.1 for a summary of the various metabolic pathways). Cytochrome P-450 enzymes mediate atrazine dealkylation.7, 17-19 In humans, the majority of the metabolites are excreted as dealkylated or mercapturate species.8 Similar metabolites have been recovered from the urine of workers exposed to atrazine dermally9 and by inhaling the vapour.13 Perfusion studies conducted on human tissue samples have also demonstrated atrazine metabolites in the serosal chamber.7

Figure 1.1 Proposed pathways for human and rodent biotransformation of atrazine.17, 20

1.3.4 Elimination The primary route of excretion of atrazine metabolites by urine has been demonstrated in humans, rats, mice and chickens.8, 14-16, 21, 22 There is no evidence of host accumulation. Following dermal application, some of the unchanged atrazine (less than 2%) is eliminated with the metabolites.8, 9 Similar results have been reported following oral ingestion in rats.14 The elimination half-life is rapid, with elimination occurring between 10.8-11.2 hours post-ingestion.15

1.4 Mechanism of Action within the Body

1.4.1 Oestrus Cycle Disruption The effect of atrazine (and other triazines) on cellular function in mammalian species is not well understood. In some strains of rat, atrazine, in doses ranging from 7 to 300 mg/kg/day, appears to disrupt female reproduction by altering the oestrus cycle.

12

Cessation of the cycle is possibly due to disruption of the hormones that control the cycle (Figure 1.2 summarises the gonadotrophic hormonal control).23 Disruption does not appear to be associated with any intrinsic estrogenic activity of atrazine, as atrazine does not bind to estrogen receptors in vitro.24 Rather, it may influence the cycle indirectly through the central nervous system25 via a hypothalamic site of action. Another hypothesis suggests atrazine may disrupt the cycle by binding to gamma-amino butyric acid-a central nervous system receptors that regulate the release of gonadotropin releasing hormone from the hypothalamus.26 Figure 1.2 The neuroendocrine control of the gonadal function.

Recent research on animal models, such as mice, have suggested atrazine may also act upon basal ganglia dopaminergic neurons27 with implications for increased risk of Parkinson's disease. However, in all of these studies, the doses of atrazine and the route of exposure (by ingestion) are unlikely to be representative of human exposure; it is difficult, therefore, to attribute these hypotheses as modes of action in human toxicity.

1.5 Summary of Clinical Signs and Symptoms Following Toxic Exposure

1.5.1 Acute Signs and Symptoms

1.5.1.1 Human (acute) There have been no verified cases of toxicity following the ingestion of atrazine.28 To the knowledge of the author, there has been only one peer-reviewed report of atrazine ingestion, which was coupled with the ingestion of ethylene glycol and formaldehyde.5 The patient subsequently developed systemic signs and symptoms that included renal failure, hepatic necrosis and disseminated intravascular coagulation and finally death from intractable shock. Evidence of symptoms presented by the patient are consistent with ethylene glycol and formaldehyde toxicity (e.g. severe metabolic acidosis with a large anion gap) and the authors concluded that toxicity from these two chemicals was the most likely cause of death.5 All reported case histories involving human exposure to atrazine involve skin and eye exposure. Acute skin exposure can lead to mild contact dermatitis with swelling, both

13

of which resolve within a minimal period of time. Allergic contact dermatitis has also been reported following skin exposure to a concentrated formulation.29 Symptoms following exposure to the hands have included localised pain, swelling, redness and blisters, with hemorrhagic bullae between the fingers, leading to incapacitation. Symptoms finally resolved, but persisted for several weeks. Eye exposure may also cause irritation and one report describes slight abrasion of the cornea, which required medication.12 Again, symptoms typically resolved without complications.

1.5.1.2 Animal (acute) Signs and symptoms related to animal exposure to atrazine are associated with acute, sub-acute, subchronic and chronic studies. They are all summarised together in section 1.5.2.2, according to the various organs.

1.5.2 Chronic Signs and Symptoms

1.5.2.1 Human (chronic) A retrospective study assessing workers in a chemical plant that were chronically exposed to atrazine concluded there was only a possible causal relationship with gastroenteritis.33 Another retrospective study of chemical plant workers concluded they had mortality rates equal to or less than the general population.34

1.5.2.2 Animal (acute, sub acute sub-chronic and chronic)

1.5.2.2.1 Introduction

The following paragraphs (sections 1.5.2.2.2 to 1.5.2.2.7) are a brief summary (and by no means exhaustive) of a range of EPA and other studies assessing the effects of acute, sub-acute and chronic dietary ingestion of atrazine upon various organs of the body in an assortment of different animal species. Such studies involve administering massive doses of atrazine (ranging from 10-200 mg/kg) over a range of time intervals, in order to determine the effects atrazine may have upon various organs of the host, and in some instances to ascertain the respective NOAEL values.

1.5.2.2.2 Cardiovascular and Haematology

Dogs fed 34 mg/kg atrazine for one year demonstrated mild changes in ECG measurements and pathology to the heart.35 Similar results have been demonstrated with rats and pigs being fed a range of atrazine with comparable doses.36-39 There are no reported changes to the gastrointestinal tracts of rats chronically fed doses of atrazine, ranging from 47 – 71 mg/kg. Variable haematological results have been reported following dietary ingestion of atrazine (71 mg/kg/day); effects ranged from no changes to mild changes in rat hematological indices such as decreases in erythrocyte, haemoglobin, and haematocrit levels.35-38

1.5.2.2.3 Hepatic and Renal

The liver appears to be the major target of atrazine toxicity with increased levels of serum transaminases, decreased blood glucose and in a variety mild histopathological changes to the liver following chronic ingestion of 100 mg/kg/day.36-38 The pig appears to be the most sensitive mammalian species.41 Renal effects have been noted only in studies involving rats and pig. Changes included increased urinary levels of

14

electrolyte, proteins and serum dehydrogenases as well as evidence of histopathological changes to the kidney.36-39, 40, 42

1.5.2.2.4 Endocrine and Reproduction

In some animal models, atrazine primarily appears to disrupt the female reproduction system by altering the oestrus cycle. This has been demonstrated in some strains of rat following the oral ingestion of acute, sub-acute and chronic doses of atrazine (50 – 300 mg/kg/day)25, 43-47 Atrazine (120 mg/kg/day for 7 days) may also interfere with rat male hormone regulation by decreasing the metabolism of testosterone to 5α-dihydroxytestosterone and binding to its receptor complex,48 and can lead to increased rat pituitary gland weight.49 In a two-generation rat study, doses of 2.5 and 25 mg/kg atrazine, pups of the second generation had increase in relative testis weight.124 Changes in thyroid size have also been reported, though it is uncertain if such changes are due to the direct action of atrazine or indirectly due to changes atrazine causes to the pituitary gland. Given atrazine’s ability to block endogenous hormones, this herbicide is therefore recognized as a potential endocrine disruptor. However, such disruption is dependent on both the dose and duration of exposure and the choice of animal strain.43, 44

1.5.2.2.5 Musculoskeletal, Dermal and Body weight

Rats and dogs fed diets of atrazine ranging from 34 – 72 mg/kg demonstrated no histological changes in skeletal muscles.36-38 Chronic dietary administration demonstrated no gross or histological abnormalities to skin or eyes.36-38 Mild to severe weight losses, however, were recorded following chronic exposure (e.g. 36-38, 50, 123), though in some studies rats made a full recovery when atrazine was removed from their diet.

1.5.2.2.6 Immunological

Rats fed acute dose of atrazine (100 mg/kg/day) demonstrated mild changes to immunoglobulin levels, subtle changes to the spleen and marginal decrease in thymus weight. There was no evidence of autoimmune disease or delayed-type hypersensitivity reaction.51 No effects whatsoever were observed at 20 mg/kg/day. Similar results have been found with other studies.52 Like other investigations, pigs appear to be the most sensitive species.39

1.5.2.2.7 Ocular and Neurological

Acute ingestion of atrazine (100 mg/kg/day) has yielded subtle changes in rat neurological functions.53 No behavioural changes have been noted on rats fed up to 75 mg/kg/day atrazine.54 There were no ocular changes in rats or dogs following chronic doses ranging from 52-483 mg/kg36-38

1.6 Carcinogenicity

1.6.1 Human Studies The association between atrazine contaminated drinking water (average concentration was 162.74 ng/L) and increased risk of cancer was determined in Ontario, Canada.55 There was a positive association with stomach cancer with an increased risk of 0.6 cases for men and 1.0 case for women per 100,000 person-years. Negative associations were found with other cancers. A study in Kentucky reported an association between breast cancer and atrazine-contaminated water (mean

15

concentration 0.11±0.19 µg/L).56 In contrast, however, a further study suggested there was no significant association with breast (or ovarian) cancer in this group of women57; a similar conclusion was reached in a study of Wisconsin women.58 Other population-based studies suggested positive associations with non-Hodgkin’s lymphoma,59-61 prostrate, brain and testes cancer 59 and leukemia.59 A retrospective cohort study of licensed pesticide applicators in Iowa and North Carolina showed a positive trend with lung and bladder cancers, non-Hodgkin lymphoma, and multiple myeloma.62 However, criticisms of some of these investigations have suggested the sample sizes were too small, residents were exposed to other chemicals and the routes of proposed exposure were not ascertained. Overall, there is not enough information to suggest there is a link between atrazine and cancer in humans. As such, the American Conference of Governmental and Industrial Hygienists63 determined atrazine was not classified as a human carcinogen. Furthermore, the IARC64 in 1999 downgraded atrazine from a group 2B category (i.e. possibly carcinogenic to humans) to group 3 (not classifiable as to its carcinogenicity in humans).

1.6.2 Animal Studies There is conflicting evidence in animal models that atrazine may cause cancer. Research has suggested that at high doses (70 -80 mg/kg/day), atrazine in the diet of rats may cause increased incidence and earlier onset of mammary tumours. Investigations with female Sprague-Dawley rats demonstrated increased incidence of mammary tumours (unpublished, but reviewed by Stevens and colleagues65), but this was contradicted by investigations employing Fischer 344 rats (unpublished but reviewed by Stevens and colleagues plus Wetzel and colleagues47, 65) or CD-1 mice (unpublished but reviewed by Stevens and colleagues65). Other investigations have failed to demonstrate statistically increased levels of other cancers or malignant tumours.66, 67

1.7 Mutagenicity Research tends to suggest that atrazine does not induce forward or reverse mutations in prokaryotic cells (e.g. 68). In contrast, there is some evidence of genotoxicity in a variety of eukaryotic cells; these include gene conversion (e.g. 69), forward and reverse mutation (e.g. 69, 70). With mammalian cell lines there has been conflicting evidence with some reports suggesting DNA damage may or may not occur (e.g. 71, 72).

1.8 Developmental Effects and Teratogenicity

1.8.1 Human studies Changes in reproductive and teratogenic effects were assessed in farming families that may have been exposed to atrazine. Research concluded there may have been an association with an increase in pre-term delivery.30, 31 There was no significant increase in risk of either early or late spontaneous abortions.30 Other studies from this survey reported no effect upon male spermatogenesis.32

16

A retrospective survey of Canadian farming families (n = 1898) who were exposed to atrazine, demonstrated no changes in the gender-birth expected sex ratio or increased risk for small-for-gestational-age births31; there may however be an association with increased pre-term delivery and miscarriage30, 31 There is no evidence of hypospadias in Arkansas between 1998 and 200 resulting from exposure to atrazine applications.73 There may be a weak association of atrazine in drinking water with intrauterine growth retardation (IUGR),74 though this was confounded by other factors (sources of drinking water and the actual risk of IUGR).

1.8.2 Animal Studies Developmental effects have been observed in rats and rabbits administered doses of 70 - 75 mg/kg/day atrazine during key time intervals of gestation; 6-15 days and 7 -19 days for rats and rabbits respectively. Offspring demonstrated incomplete ossification of the skull, hyoid bone, teeth, forepaw metacarpals, and hind paw distal phalanges.75 There were decreased foetal body weights, increased litter resorptions and reduced live litters. No developmental effects were observed at 5 mg/kg/day. In contrast, however, no developmental effects were found in other studies with rat species given dietary intakes ranging from 31 to 113 mg/kg/day.76, 77 Rats pre-treated sub-acutely with atrazine (0 – 120 mg/kg), to assess pregestational and lactational effects on offspring’s central nervous system and reproduction, were later mated with untreated males; litter size, weight and survival remained unchanged though there were later mild neurobehavioral effects noted.78-80 Males from the litter administered 50 mg/kg/day or higher had evidence of histopathological changes to the prostrate.80 Vaginal opening, a marker of female puberty in the rat, was delayed in rats given 30 mg/kg/day or greater of atrazine.81, 82 Delays in uterine growth were also observed at doses of 100 mg/kg/day.

1.9 Environmental fate (metabolism and accumulation) and subsequent ecotoxicity

1.9.1 Environmental Levels

Air

Due to its persistence in the environment, atmospheric levels have been determined, more often following its application to crops. Monitoring levels following such applications, it was found present in rainfall collected four months after its use.83 Concentration in rural areas within the vicinity of Paris in 1992-93 were 5–400 ng/L, with similar to values in adjacent urban areas. Assessment of agrochemical fluxes in the atmosphere over Ottawa in June 1993 and July 1994, demonstrated peak atrazine concentrations of 4.6 ng/m3.84 Increasing airborne concentrations may be associated with increasing concentrations in the vapour phase (associated with rainfall) which could reflect duration, intensity and distance from the site of atrazine application.

Water

In agricultural watershed streams, average concentrations of atrazine ranged from 1.1 to 1.6 µg/L.85 Median concentrations from the Mississippi River at Baton Rouge, Louisiana, that were sampled from 1991 to 1997, were approximately 0.45 µg/L.86

17

Differences in streams and ground water are demonstrated in a large study in the United States mid-west that showed median concentrations of 3.97 and 0.001 µg/L, respectively.87 Water in drinking wells can be contaminated by atrazine runoff. In one study, for example, levels of atrazine have been detected in 31% of wells at concentrations less than 0.6 µg/L atrazine (though two were assayed at 3 µg/L).88 A total of 82 wells in New Zealand were assayed in 1990 for a range of pesticides including atrazine89; one well had 37 µg/ L, an order of magnitude in excess of the WHO TDI in drinking water (see section 1.10.4). A second well was slightly above the WHO limit and the remaining 80 had undetectable levels. Another survey assessing 111 wells throughout New Zealand was undertaken in 2002.90 Atrazine was detected in several wells, with concentrations two orders of magnitude below the WHO TDI in drinking water.

Soil

Concentrations of atrazine in soil may vary, depending on climatic patterns and agricultural usage. Where atrazine has been applied (1.1 kg per ha), levels of atrazine in surface soil have declined from 0.83 µg/g 6 days post-application to 0.5 µg/g after 2 months to 0.08 µg/g after 12 months.98 Similar trends occur following more intensive application.91 In all instances, concentrations in the soil drop by 90% within one year.

1.9.2 Environmental Fate To be active as a herbicide, atrazine must first be dissolved in water before it can enter through the roots and act upon the plant (by preventing photosynthesis). All plants take up atrazine, but those plants not affected by its action soon break down the chemical. Any remaining atrazine in the soil is either washed off by rainwater into streams and lakes92, 93 or may leach through successive layers of soil94 where it can then enter groundwater. Overall, atrazine (and its metabolites) can persist and thereby contaminate streams, lakes and costal estuaries.95 In the United States, for example, atrazine is up to 20 times more frequently detected in ground water than any other herbicide.96 Any residues that remain within the soil are broken down, typically over the period of a growing season. Atrazine has low rates of volatility,97 and airborne dispersal favours the particulate phase (binding to dust and other small particles) rather than the vapour phase. Nevertheless, it is discharged into the air via spray droplets (typically by spray applicators) where it may react with other chemicals, pollute fog or rain98, 99 or will bind to airborne particulate matter.100 Bound chemical can then travel large distances from the site of application; in one instance it was assayed in ground water 100–300 km from the site of application.101, 102 Atrazine is removed from the atmosphere by precipitation and dry deposition.101

1.9.3 Ecological Degradation Atrazine can remain stable as a dry powder for many years.103 However, once it has been applied in the environment, typically as a mist or spray, environment, it can subsequently be degraded by one of three pathways:

1. Hydrolysis of the hydrogen carbon bond leads to the formation of hydroxyatrazine.94, 104 This pathway typically occurs in soil and aquatic environments. Indeed, this is regarded as the major (non-microbiological) pathway leading to the degradation of this herbicide.104

2. N-dealkylation of carbon number 4 and/or 6. Degradation yields deethylated atrazine, deisopropylated atrazine and 2-chloro-4,6-diaminotriazine.94, 105 The first two products are photolytic metabolites.

18

3. Micro-organism degradation – a series of metabolic steps causing the splitting of the triazine ring,103 dehalogenation, dealkylation and in some instances leading to the final formation of urea.106, 107

Rates of degradation in soil, air and water have been reported, but values are variable. A brief summary of the literature suggests the following mean rates:

Soil half-life: 14 – 109 days108, 109 Half-life in water: >200 days Half-life in the vapour phase : 14 hours110

Direct photo-degradation of vapour-phase atrazine in the atmosphere is rapid (14 hour half-life). In contrast to vapour-phase atrazine, particulate-bound atrazine will not degrade at the same rate (which may explain why the intact herbicide can be transported such substantial distances in the atmosphere). The degradation of atrazine in water, however, is slow, with mean residence times in excess of 200 days in lakes and streams. Photolysis does not occur unless there is a substantial amount of dissolved organic matter of pH below 7.111, 112 Degradation is predominantly by hydrolysis113 but aerobic bacteria may make a small contribution. No biodegradation has been observed in ground water associated with aquifer sediment systems under anaerobic conditions.114 Degradation and losses from soil occur by various mechanisms: photolysis, biodegradation, percolation into groundwater, volatilisation, and irreversible binding to soils.115 The degradation half-life is extended in dry and cold conditions. Residues of atrazine and its metabolites have remained bound to soil-based sediments two years after the initial exposure.116 Bacteria appear to play a negligible role.117

1.9.3.1 Effect on Ecological Food Chain Despite remaining in the environment, research does suggest atrazine is not accumulated and concentrated in the ecological food chain as it is rapidly metabolised and eliminated from the respective hosts (see sections 1.3.3 and 1.3.4). Nevertheless, there is a body of literature that suggest the presence of atrazine and its metabolites itself may have an influence on various aquatic species ranging from simple photosynthetic organisms through to fish and ultimately humans that consume the water.118

1.10 Dose Response Information

1.10.1 Doses Necessary to Cause Acute and Chronic Toxicity

1.10.1.1 Acute Based on animal LD50 studies (summaries are provided in section 1.10.3.1), atrazine is regarded as a compound of low acute toxicity to humans.28, 119 Death, in association with other more deadly co-ingestants, has followed the ingestion of 500 mL atrazine.5 To the knowledge of the author, no other reports of acute toxicity in humans have been reported.

19

A review of all animal livestock exposures to atrazine over a ten-year interval, reported to the European Veterinary Poison Control Center concluded the risk of atrazine toxicity is small.120

1.10.1.2 Chronic There is limited information on the chronic toxicity of atrazine and less so on defined exposures and/or blood concentrations of studied participants. For example, reproductive and developmental toxicity were assessed in three studies based solely on questionnaires.30-32 Farming families in Iowa who may have been exposed to atrazine residues in drinking water (mean atrazine concentration was 2.2 µg/L) may have a weak association with intra-uterine growth retardation.74 See section 1.8.1.

A study in Maryland observed possible associations between diagnosis of childhood leukaemia and bone cancer and residents living within 3.2 km from well water containing detectable levels of nitrate and pesticides including atrazine, though this was not demonstrated with atrazine alone.121 The authors, however, stressed such an association was tenuous, subject to further investigations and should remain as a suggested hypothesis. The maximum detectable concentration of atrazine in these wells was 1.7 µg/L; this value is almost half the defined United States EPA maximum contaminant level that is associated with little or no risk of adverse health effects or cancer (see Appendix 3).

1.10.2 Summary of Case Reports Following Human Toxicological Exposure To the knowledge of the author, there has been only one peer-reviewed report of atrazine ingestion, which was coupled with the ingestion of ethylene glycol and formaldehyde.5 This case has been adequately covered in section 1.5.1.1.

1.10.3 Measures of Mammalian Toxicity Measures of toxicity in mammals, for example, are used to determine acceptable levels of exposure in humans. For example, the US National Academy of Sciences used a LOAEL that caused low chronic toxicity122 to derive their ADI for atrazine.

1.10.3.1 LD50 Values For an explanation of LD50 values see Appendix one. A summary of various LD50 values from a range of mammalian species and routes of exposure are summarised in Table 1.2. Table 1.2 Table Summaries of LD50 single dose values for atrazine28

Species Route LD50 (mg/kg/bw) Rat Oral 1900 - 3000 Rat Dermal >3000 Mouse Oral 1750 Rabbit Oral 750

20

1.10.4 Guidance Values

Tolerable Daily Intakes

A summary of tolerable daily intakes (TDI), and how they are derived, are presented in table 1.3. For an explanation of TDI, see appendix 1 Table 1.3 Tolerable daily intakes for atrazine, as defined by various regulatory authorities

Agency EPA1 RIVM2

Type of Exposure Chronic ingestion RfD

Chronic ingestion TDI

Value 35 µg/kg/day 5 µg/kg/day

Year 1993 1999/2000

Species rat rat

Critical effect Body weight Reproductive organs

NOAEL/LOAEL

3.5 mg/kg/day NOAEL

0.5 mg/kg/day NOAEL

Uncertainty factor 100(10A, 10H) 100(10A, 10H)

Principle study 123 124 1EPA – United States Environmental Protection Agency 2RIVM - Rijksinstituut voor Volksgezondheid en Milieu (National Institute of Public Health and the Environment, the Netherlands)

Other Measures of Daily Intake

The European Union has limited the maximum allowable concentration for a single pesticide to 0.1 µg/L (Drinking Water Directive: 98/83/EC, 1998) and their presence in different foods and drinks is limited by legislation. The WHO125 has established an international guideline for TDI’s of atrazine in drinking water at 2 µg/litre. Regulations governing atrazine residues in food vary, depending on the foodstuff and the country that legislates its control. They range from 0.01 (Australian meat and dairy products) to 10 (American Pineapple fodder and forage) mg/kg atrazine residue limits.64 In some European countries, it is banned from use. Based on the possible health risks and exposure to the general population, the EPA has set the Maximum Contaminant Level for atrazine in water systems at 0.003 mg/L (see Appendix 3).

1.11 A summary of Biological Assays that are Available to Test for Evidence of these Chemicals A commonly used method for sensitive and specific analysis of atrazine has been high-pressure liquid chromatography (HPLC), linked to a detector that is either mass spectrometry (e.g. 126) or tandem mass spectrometry (e.g. 127). Other investigators

21

have employed gas chromatography for their solute separations.128 As these methods are very time consuming and expensive, other more cost effective and rapid techniques have been developed though not necessarily used widely; these have included enzyme linked immunoassays129 and immunosensors (e.g. 130). Atrazine can be assayed from biological samples including plasma,5 saliva,131 skin,132 urine8 and from hepatic tissue.133 Before analysis can be achieved, extraction of atrazine from the surrounding matrix (such as plasma) is required; such methods have included solvent extraction, solid phase extraction, supercritical water extraction, microwave-assisted extraction (MAE) and pressurised liquid extraction (e.g. 134-136). Given the complexity of these procedures, analysis can be slow, cumbersome and expensive. The determination of atrazine levels in the human body may be difficult due to its rapid excretion from the body15 (half life is between 10.8 to 11.2 hours post-ingestion). If analysis needs to be undertaken, it would be prudent to assay both atrazine and its key metabolites in urine. However, the presence of these compounds in the body would only indicate recent contact with this pesticide. High Resolution Gas Chromatography Mass Spectrometry is the most commonly recommended and used method to assay for this herbicide; analysis must be undertaken in a laboratory that has ISO/IEC 17025 registration. As with dioxin and PCP, the laboratory in New Zealand most suited to assess this pesticide, having the necessary skills and registrations for such analysis, is AsureQuality Limited in Lower Hutt. Though they presently have not assayed it in either urine or blood, they have extensive experience in measuring levels in water, food and soil. Development of an assay for either blood or urine should not be difficult.

1.12 Summary Atrazine consists of a triazine molecule with two amine groups and one chlorine atom. Due to its persistence in the environment its use has been restricted and banned in some countries though it is still widely used in the United States (and New Zealand). It can absorb across the skin, intestinal and respiratory tracts though rates and bioavailability for dermal absorption are far lower than via the intestinal route. Atrazine will readily distribute to most organs of the body followed by rapid elimination, predominantly via the kidneys. The effect of atrazine in mammalian species is not well understood, but based on animal studies, where animals were fed massive doses of this herbicide, it may function as an endocrine disruptor. In humans there are no reports of atrazine toxicity that are attributed to this herbicide alone. Skin exposure can cause mild contact dermatitis, which soon resolves. Evidence of toxicity in humans following chronic exposure to lower levels of atrazine are not as well defined; there may be statistical associations with increased risk of pre-term delivery and gastroenteritis. In contrast, toxicity has been achieved in animal studies though the administered doses necessary to achieve observable effects ranged from 10 –200 mg/kg body weight. These effects have included mild haematological changes, hepatic and renal injury, endocrine disruptions, weight losses and mild changes to neurological and immunological indices. Based on LD50 studies, atrazine has been regarded as of low acute toxicity to humans.

22

There is some evidence, though conflicting, that atrazine may cause increased incidents of mammary cancers in rats. In humans, there is not enough information to suggest there is a link between atrazine and cancer; however, based on limited evidence in animal studies, atrazine has been classified as possibly carcinogenic to humans. There is insufficient evidence to suggest atrazine causes mutagenicity. There may be a weak statistical association in humans with atrazine causing mild developmental effects. Atrazine can persist and contaminate streams and lakes due to run off from pastoral applications. It can also be dispersed large distances from the site of application, bound to airborne particulates. However, once it has been solubilised and dispersed onto crops or used in other applications, it will degrade within a year, though the rate of degradation will depend on the surrounding matrix. As a dry powder, atrazine can remain stable for many years, though once in contact with the soil, degradation will occur, albeit at a rate slower than the solubilised form. Despite remaining in the environment, it does not bioaccumulate in the ecological food chain.

1.13 References 1. Steinberg CEW, Lorenz R, Spieser OH. Effects of atrazine on swimming

behaviour of zebrafish, Brachydanio rerio. Water Research. 1995;29:981-985. 2. Sass JB, Colangelo A. European Union bans atrazine, while the United States

negotiates continued use. International Journal of Occupational and Environmental Health. 2006;12:260-267.

3. Young S. New Zealand Novachem Agrichemical Manual. Christchurch: AgriMedia Ltd; 2008.

4. Gysin H, Knuesli E. Chemistry and herbicidal properties of triazine derivatives. In: Metcalf R, ed. Advances in Pest Control Research. New York: Wiley (Interscience); 1960.

5. Pommery J, Mathieu M, Mathieu D, Lhermitte M. Atrazine in plasma and tissue following atrazine-aminotriazole-ethylene glycol-formaldehyde poisoning. Journal of Toxicology-Clinical Toxicology. 1993;31:323-331.

6. Brand RM, Cetin Y, Mueller C, Cuppett SL. Effect of fatty acids on herbicide transport across Caco-2 cell monolayers. Toxicology in Vitro. 2005;19:595-601.

7. Ademola JI, Sedik LE, Wester RC, Maibach HI. In vitro percutaneous absorption and metabolism in man of 2-chloro-4-ethylamino-6-isopropylamine-s-triazine (atrazine). Archives of Toxicology. 1993;67:85-91.

8. Buchholz BA, Fultz E, Haack KW, et al. HPLC-accelerator MS measurement of atrazine metabolites in human urine after dermal exposure. Analytical Chemistry. 1999;71:3519-3525.

9. Catenacci G, Barbieri F, Bersani M, Ferioli A, Cottica D, Maroni M. Biological monitoring of human exposure to atrazine. Toxicology Letters. 1993;69:217-222.

10. Catenacci G, Maroni M, Cottica D, Pozzoli L. Assessment of human exposure to atrazine through the determination of free atrazine in urine. Bulletin of Environmental Contamination and Toxicology. 1990;44:1-7.

11. Hall LL, Fisher HL, Sumler MR, Monroe RJ, Chernoff N, Shah PV. Dose response of skin absorption in young and adult rats. In: Mansdorf SZ, Sager R, Nielson AP, eds. Performance of Protective Clothing: Second Symposium. Philadelphia: American Society for Testing and Materials; 1988:177-194.

23

12. Technical Information. Atrazine, Pesticide Fact Sheet. Memphis, TN: Chemlawn Services Co; 1987.In Stevens JT, Sumner DD. Herbicides. In: Hayes WJ, Laws ER, eds. Handbook of Pesticide Toxicology. San Diego: Academic Press Inc; 1981:1317-1408.

13. Ikonen R, Kangas J, Savolainen H. Urinary atrazine metabolites as indicators for rat and human exposure to atrazine. Toxicology Letters. 1988;44:109-112.

14. Meli G, Bagnati R, Fanelli R, Benfenati E, Airoldi L. Metabolic profile of atrazine and n-nitrosoatrazine in rat urine. Bulletin of Environmental Contamination & Toxicology. 1992;48:701-708.

15. Timchalk C, Dryzga MD, Langvardt PW, Kastl PE, Osborne DW. Determination of the effect of tridiphane on the pharmacokinetics of [14C]-atrazine following oral administration to male Fischer 344 rats. Toxicology. 1990;61:27-40.

16. Bakke JE, Larson JD, Price CE. Metabolism of atrazine and 2-hydroxyatrazine by the rat. Journal of Agricultural & Food Chemistry. 1972;20:602-607.

17. Hanioka N, Jinno H, Tanaka-Kagawa T, Nishimura T, Ando M. In vitro metabolism of simazine, atrazine and propazine by hepatic cytochrome P450 enzymes of rat, mouse and guinea pig, and oestrogenic activity of chlorotriazines and their main metabolites. Xenobiotica. 1999;29:1213-1226.

18. Hanioka N, Jinno H, Kitazawa K, et al. In vitro biotransformation of atrazine by rat liver microsomal cytochrome P450 enzymes. Chemico-Biological Interactions. 1998;116:181-198.

19. Adams NH, Levi PE, Hodgson E. In vitro studies of the metabolism of atrazine, simazine, and terbutryn in several vertebrate species. Journal of Agricultural and Food Chemistry. 1990;38:1411-1417.

20. Hanioka N, Jinno H, Tanaka-Kagawa T, Nishimura T, Ando M. In vitro metabolism of chlorotriazines: characterization of simazine, atrazine, and propazine metabolism using liver microsomes from rats treated with various cytochrome P450 inducers. Toxicology & Applied Pharmacology. 1999;156:195-205.

21. Foster TS, Khan SU. Metabolism of atrazine by the chicken. Journal of Agricultural & Food Chemistry. 1976;24:566-570.

22. Ross MK, Filipov NM. Determination of atrazine and its metabolites in mouse urine and plasma by LC-MS analysis. Analytical Biochemistry. 2006;351:161-173.

23. Goldman JM, Stoker TE, Cooper RL. Neuroendocrine and reproductive effects of contemporary-use pesticides. Toxicology and Industrial Health. 1999;15:26-36.

24. Connor K, Howell J, Chen I, et al. Failure of chloro-s-triazine-derived compounds to induce estrogen receptor-mediated responses in vivo and in vitro. Fundamental & Applied Toxicology. 1996;30:93-101.

25. Cooper RL, Stoker TE, Tyrey L, Goldman JM, McElroy WK. Atrazine disrupts the hypothalamic control of pituitary-ovarian function. Toxicological Sciences. 2000;53:297-307.

26. Shafer TJ, Ward TR, Meacham CA, Cooper RL. Effects of the chlorotriazine herbicide, cyanazine, on GABA(A) receptors in cortical tissue from rat brain. Toxicology. 1999;142:57-68.

27. Coban A, Filipov NM. Dopaminergic toxicity associated with oral exposure to the herbicide atrazine in juvenile male C57BL/6 mice. Journal of Neurochemistry. 2007;100:1177-1187.

24

28. Stevens JT, Sumner DD. Herbicides. In: Hayes WJ, Laws ER, eds. Handbook of Pesticide Toxicology. San Diego: Academic Press Inc; 1981:1317-1408.

29. Schlicher JE, Beat VB. Dermatitis resulting from herbicide use--a case study. Journal of the Iowa Medical Society. 1972;62:419-420.

30. Arbuckle TE, Lin ZQ, Mery LS. An exploratory analysis of the effect of pesticide exposure on the risk of spontaneous abortion in an Ontario farm population. Environmental Health Perspectives. 2001;109:851-857.

31. Savitz DA, Arbuckle T, Kaczor D, Curtis KM. Male pesticide exposure and pregnancy outcome. American Journal of Epidemiology. 1997;146:1025-1036.

32. Curtis KM, Savitz DA, Weinberg CR, Arbuckle TE. The effect of pesticide exposure on time to pregnancy. Epidemiology. 1999;10:112-117.

33. Gass R. Atrazin-produktion; gesundheitsrisiko bei chemiearbeitern? Internationale Gesellschaft für Epidemiologie, Europäische Regionaltagung. Basel, Switzerland; 1991.

34. Sathiakumar N, Delzell E, Austin H, Cole P. A follow-up study of agricultural chemical production workers. American Journal of Industrial Medicine. 1992;21:321-330.

35. EPA. Chronic toxicity study in dogs. EPA MRID 404313-01. EPA Guidelines No. 83-1. Washington, DC: U.S. Environmental Protection Agency; 1987. In ATSDR, Toxicological profile for atrazine. 2003, Department of Health and Human Services, Public Health Service. Agency for Toxic Substances and Disease Registry: Atlanta, Ga.

36. EPA. Twenty-four month combined chronic oral toxicity study of rats utilizing atrazine technical. Twelve month interim report for toxigenics study 410-1102. EPA TRID 4426-010-19. Washington, DC: U.S. Environmental Protection Agency; 1984. In ATSDR, Toxicological profile for atrazine. 2003, Department of Health and Human Services, Public Health Service. Agency for Toxic Substances and Disease Registry: Atlanta, Ga.

37. EPA. Twenty-four month combined chronic oral toxicity study of rats utilizing atrazine technical. Final report for toxigenics study 410-1102. EPA TRID 4701-920-30. Washington, DC: U.S. Environmental Protection Agency; 1986. In ATSDR, Toxicological profile for atrazine. 2003, Department of Health and Human Services, Public Health Service. Agency for Toxic Substances and Disease Registry: Atlanta, Ga.

38. EPA. Supplement to two-year chronic feeding/oncogenicity study in rats administered atrazine. EPA MRID 406293-02. EPA Guidelines No. 83-5. Washington, DC: U.S. Environmental Protection Agency; 1987. In ATSDR, Toxicological profile for atrazine. 2003, Department of Health and Human Services, Public Health Service. Agency for Toxic Substances and Disease Registry: Atlanta, Ga.

39. Curic S, Gojmerac T, Zuric M. Morphological changes in the organs of gilts induced with low-dose atrazine. Veterinarski Arhiv. 1999;69:135-148.

40. Santa Maria C, Moreno J, Lopez-Campos JL. Hepatotoxicity induced by the herbicide atrazine in the rat. Journal of Applied Toxicology. 1987;7:373-378.

41. Gojmerac T, Kartal B, Zuric M, Curic S, Mitak M. Serum biochemical and histopathological changes related to the hepatic function in pigs following atrazine treatment. Journal of Applied Toxicology. 1995;15:233-236.

42. EPA. Oncogenicity study in mice. EPA MRID 404313-02. EPA Guidelines No. 83-2. Washington, DC: U.S. Environmental Protection Agency; 1987. In ATSDR, Toxicological profile for atrazine. 2003, Department of Health and

25

Human Services, Public Health Service. Agency for Toxic Substances and Disease Registry: Atlanta, Ga.

43. Aso S, Anai N, Noda S, Imatanaka N, Yamasaki K, Maekawa A. Twenty-eight-day repeated-dose toxicity studies for detection of weak endocrine disrupting effects of nonylphenol and atrazine in female rats. Journal of Toxicologic Pathology. 2000;13:13-20.

44. Cooper RL, Stoker TE, Goldman JM, Parrish MB, Tyrey L. Effect of atrazine on ovarian function in the rat. Reproductive Toxicology. 1996;10:257-264.

45. Eldridge JC, Fleenor-Heyser DG, Extrom PC, et al. Short-term effects of chlorotriazines on estrus in female Sprague-Dawley and Fischer 344 rats. Journal of Toxicology & Environmental Health. 1994;43:155-167.

46. Eldridge JC, Wetzel LT, Tyrey L. Estrous cycle patterns of Sprague-Dawley rats during acute and chronic atrazine administration. Reproductive Toxicology. 1999;13:491-499.

47. Wetzel LT, Luempert LG, 3rd, Breckenridge CB, et al. Chronic effects of atrazine on estrus and mammary tumor formation in female Sprague-Dawley and Fischer 344 rats. Journal of Toxicology & Environmental Health. 1994;43:169-182.

48. Kniewald J, Osredecki V, Gojmerac T, Zechner V, Kniewald Z. Effect of s-triazine compounds on testosterone metabolism in the rat prostate. Journal of Applied Toxicology. 1995;15:215-218.

49. Simic' B, Kniewald J, Kniewald Z. Effects of Atrazine On Reproductive-Performance in the Rat. Journal of Applied Toxicology. 1994;14:401-404.

50. Cantemir C, Cozmei C, Scutaru B, Nicoara S, Carasevici E. p53 protein expression in peripheral lymphocytes from atrazine chronically intoxicated rats. Toxicology Letters. 1997;93:87-94.

51. Liskova A, Wagnerova J, Tulinska J, Kubova J. Effect of the herbicide atrazine on some immune parameters in mice. Journal of Trace and Microprobe Techniques. 2000;18:235-240.

52. Vos JG, Krajnc EI. Immunotoxicity of pesticides. Developments in Toxicology and Environmental Science. 1983;11:229-240.

53. Podda MV, Deriu F, Solinas A, et al. Effect of atrazine administration on spontaneous and evoked cerebellar activity in the rat. Pharmacological Research. 1997;36:199-202.

54. Dési I. Neurotoxicological investigation of pesticides in animal-experiments. Neurobehavioral Toxicology and Teratology. 1983;5:503-515.

55. Van Leeuwen JA, Waltner-Toews D, Abernathy T, Smit B, Shoukri M. Associations between stomach cancer incidence and drinking water contamination with atrazine and nitrate in Ontario (Canada) agroecosystems, 1987-1991. International Journal of Epidemiology. 1999;28:836-840.

56. Kettles MA, Browning SR, Prince TS, Horstman SW. Triazine herbicide exposure and breast cancer incidence: An ecologic study of Kentucky counties. Environmental Health Perspectives. 1997;105:1222-1227.

57. Hopenhayn-Rich C, Stump ML, Browning SR. Regional assessment of atrazine exposure and incidence of breast and ovarian cancers in Kentucky. Archives of Environmental Contamination & Toxicology. 2002;42:127-136.

58. McElroy JA, Gangnon RE, Newcomb PA, et al. Risk of breast cancer for women living in rural areas from adult exposure to atrazine from well water in Wisconsin. Journal of Exposure Science & Environmental Epidemiology. 2007;17:207-214.

26

59. Mills PK. Correlation analysis of pesticide use data and cancer incidence rates in California counties. Archives of Environmental Health. 1998;53:410-413.

60. Weisenburger DD. Environmental epidemiology of non-Hodgkin's lymphoma in eastern Nebraska. American Journal of Industrial Medicine. 1990;18:303-305.

61. Zahm SH, Weisenburger DD, Saal RC, Vaught JB, Babbitt PA, Blair A. The role of agricultural pesticide use in the development of non-Hodgkins-lymphoma in women. Archives of Environmental Health. 1993;48:353-358.

62. Rusiecki JA, De Roos A, Lee WJ, et al. Cancer incidence among pesticide applicators exposed to atrazine in the agricultural health study. Journal of the National Cancer Institute. 2004;96:1375-1382.

63. ACGIH. Documentation of the threshold limit values and biological indices. Cincinnati, OH: American Conference of Governmental Industrial Hygienists.; 2000.

64. IARC. Atrazine. IARC monographs on the evaluation of carcinogenic risks to humans: Some chemicals that cause tumours of the kidney or urinary bladder in rodents and some other substances. Available at <http://monographs.iarc.fr/ENG/Monographs/vol73/mono73-2.pdf> (Last accessed on 10 Nov 2008). Lyon, France: World Health Organization, International Agency for Research on Cancer; 1999:59-114.

65. Stevens JT, Breckenridge CB, Wetzel L, et al. A risk characterization for atrazine: Oncogenicity profile. Journal of Toxicology and Environmental Health-Part a. 1999;56:69-109.

66. Pintér A, Török G, Börzsönyi M, et al. Long-term carcinogenicity bioassay of the herbicide atrazine in F344 rats. Neoplasma. 1990;37:533-544.

67. Thakur AK, Wetzel LT, Voelker R, Wakefield AE. Results of a two-year oncogenicity study in Fischer 344 rats with atrazine. In: Ballantine LG, McFarland JE, Hackett DS, eds. Traizine herbicides (ACS Symposium Series No. 683). Washington DC: American chemical Society; 1998:384-398.

68. Adler ID. A review of the coordinated research effort on the comparison of test systems for the detection of mutagenic effects, sponsored by the E.E.C. Mutation Research. 1980;74:77-93.

69. Morichetti E, Dellacroce C, Rosellini D, Bronzetti G, Soldani G. Genetic and biochemical-studies on a commercial preparation of atrazine. Toxicological and Environmental Chemistry. 1992;37:35-41.

70. Emnova EE, Mereniuk GV, Tsurkan LG. Izuchenie geneticheskoi aktivnosti gerbitsidov sim-triazinovoi gruppy na shtammakh drozhzhei Saccharomyces cerevisiae. Tsitologiia i Genetika. 1987;21:127-131.

71. Ribas G, Frenzilli G, Barale R, Marcos R. Herbicide-induced DNA damage in human lymphocytes evaluated by the single-cell gel electrophoresis (SCGE) assay. Mutation Research. 1995;344:41-54.

72. Ribas G, Surralles J, Carbonell E, Creus A, Xamena N, Marcos R. Lack of genotoxicity of the herbicide atrazine in cultured human lymphocytes. Mutation Research. 1998;416:93-99.

73. Meyer KJ, Reif JS, Veeramachaneni DNR, Luben TJ, Mosley BS, Nuckols JR. Agricultural pesticide use and hypospadias in eastern Arkansas. Environmental Health Perspectives. 2006;114:1589-1595.

74. Munger R, Isacson P, Hu S, et al. Intrauterine growth retardation in Iowa communities with herbicide-contaminated drinking water supplies. Environmental Health Perspectives. 1997;105:308-314.

27

75. Infurna R, Levy B, Meng C, et al. Teratological evaluations of atrazine technical, a triazine herbicide, in rats and rabbits. Journal of Toxicology & Environmental Health. 1988;24:307-319.

76. EPA. Two-generation reproduction study in rats. EPA MRID 404313-03. EPA Guidelines No. 83-4. Washington, DC: U.S. Environmental Protection Agency; 1987.

77. Peters JW, Cook RM. Effects of atrazine on reproduction in rats. Bulletin of Environmental Contamination & Toxicology. 1973;9:301-304.

78. Peruzovic M, Kniewald J, Capkun V, Milkovic K. Effect of atrazine ingested prior to mating on rat females and their offspring. Acta Physiologica Hungarica. 1995;83:79-89.

79. Stoker TE, Laws SC, Guidici DL, Cooper RL. The effect of atrazine on puberty in male wistar rats: an evaluation in the protocol for the assessment of pubertal development and thyroid function. Toxicological Sciences. 2000;58:50-59.

80. Stoker TE, Robinette CL, Cooper RL. Maternal exposure to atrazine during lactation suppresses suckling-induced prolactin release and results in prostatitis in the adult offspring. Toxicological Sciences. 1999;52:68-79.

81. Laws SC, Ferrell JM, Stoker TE, Schmid J, Cooper RL. The effects of atrazine on female wistar rats: an evaluation of the protocol for assessing pubertal development and thyroid function. Toxicological Sciences. 2000;58:366-376.

82. Ashby J, Tinwell H, Stevens J, Pastoor T, Breckenridge CB. The effects of atrazine on the sexual maturation of female rats. Regulatory Toxicology & Pharmacology. 2002;35:468-473.

83. Chevreuil M, Garmouma M. Occurrence of triazines in the atmospheric fallout on the catchment basin of the river Marne (France). Chemosphere. 1993;27:1605-1608.

84. Zhu T, Desjardins RL, Macpherson JI, Pattey E, St Amour G. Aircraft measurements of the concentration and flux of agrochemicals. Environmental Science & Technology. 1998;32:1032-1038.

85. Frank R, Braun HE, Holdrinet MV, Sirons GJ, Ripley BD. Agriculture and water-quality in the Canadian Great-Lakes Basin .5. Pesticide use in 11 agricultural watersheds and presence in stream water, 1975-1977. Journal of Environmental Quality. 1982;11:497-505.

86. Clark GM, Goolsby DA, Battaglin WA. Seasonal and annual load of herbicides from the Mississippi River basin to the Gulf of Mexico. Environmental Science & Technology. 1999;33:981-986.

87. Battaglin WA, Furlong ET, Burkhardt MR, Peter CJ. Occurrence of sulfonylurea, sulfonamide, imidazolinone, and other herbicides in rivers, reservoirs and ground water in the Midwestern United States, 1998. Science of the Total Environment. 2000;248:123-133.

88. Bushway RJ, Hurst HL, Perkins LB, et al. Atrazine, alachlor, and carbofuran contamination of well water in central Maine. Bulletin of Environmental Contamination & Toxicology. 1992;49:1-9.

89. Close ME. Assessment of Pesticide Contamination of Groundwater in New-Zealand .2. Results of Groundwater Sampling. New Zealand Journal of Marine and Freshwater Research. 1993;27:267-273.

90. Close ME, Flintoft MJ. National survey of pesticides in groundwater in New Zealand - 2002. New Zealand Journal of Marine and Freshwater Research. 2004;38:289-299.

91. Frank R, Sirons GJ. Dissipation of atrazine residues from soils. Bulletin of Environmental Contamination and Toxicology. 1985;34:541-548.

28

92. Hall JK, Pawlus M, Higgins ER. Losses of atrazine in runoff water and soil sediment. Journal of Environmental Quality. 1972;1:172–176.

93. Trevisan M, Montepiani C, Ragozza L, Bartoletti C, Ioannilli E, Delre AAM. Pesticides in rainfall and air in Italy. Environmental Pollution. 1993;80:31-39.

94. Schiavon M. Studies of the leaching of atrazine, of its chlorinated derivatives, and of hydroxyatrazine from soil using 14C ring-labeled compounds under outdoor conditions. Ecotoxicology & Environmental Safety. 1988;15:46-54.

95. Bester K, Huhnerfuss H. Triazines in the Baltic and North-Sea. Marine Pollution Bulletin. 1993;26:423-427.

96. Belluck DA, Benjamin SL, Dawson T. Groundwater contamination by atrazine and Its metabolites - risk assessment, policy, and legal implications. 20th National Meeting of the American Chemical Society. Washington DC: ACS Symposium Series; 1991:254-273.

97. Scheunert I, Zhang QA, Korte F. Comparative-studies of the fate of atrazine-C-14 and pentachlorophenol-C-14 in various laboratory and outdoor soil-plant systems. Journal of Environmental Science and Health Part B-Pesticides Food Contaminants and Agricultural Wastes. 1986;21:457-485.

98. Buster HR. Atrazine and other s-triazine herbicides in lakes and in rain in Switzerland. Environmental Science & Technology. 1990;24:1049-1058.

99. Glotfelty DE, Seiber JN, Liljedahl LA. Pesticides in fog. Nature. 1987;325:602–605.

100. Muller SR, Berg M, Ulrich MM, Schwarzenbach RP. Atrazine and its primary metabolites in Swiss lakes: Input characteristics and long-term behavior in the water column. Environmental Science & Technology. 1997;31:2104-2113.

101. Thurman EM, Cromwell AE. Atmospheric transport, deposition, and fate of triazine herbicides and their metabolites in pristine areas at Isle Royale National Park. Environmental Science & Technology. 2000;34:3079-3085.

102. Thurman EM, Goolsby DA, Meyer T, Cromwell AE. Evidence of long-range atmospheric transport and degradation of atrazine and deethylatrazine. ACS National Meeting. Anaheim, CA; 1995:286-287.

103. Huber W. Ecotoxicological relevance of atrazine in aquatic systems. Environmental Toxicology and Chemistry. 1993;12:1865-1881.

104. Winkelmann DA, Klaine SJ. Degradation and bound residue formation of 4 atrazine metabolites, deethylatrazine, deisopropylatrazine, dealkylatrazine and hydroxyatrazine, in a western Tennessee Soil. Environmental Toxicology and Chemistry. 1991;10:347-354.

105. Sirons GJ, Frank R, Sawyer T. Residues of atrazine, cyanazine, and their phytotoxic metabolites in a clay loam soil. Journal of Agricultural & Food Chemistry. 1973;21:1016-1020.

106. de Souza ML, Wackett LP, Boundy-Mills KL, Mandelbaum RT, Sadowsky MJ. Cloning, characterization, and expression of a gene region from Pseudomonas sp. strain ADP involved in the dechlorination of atrazine. Applied & Environmental Microbiology. 1995;61:3373-3378.

107. Nagy I, Compernolle F, Ghys K, Vanderleyden J, De Mot R. A single cytochrome P-450 system is involved in degradation of the herbicides EPTC (S-ethyl dipropylthiocarbamate) and atrazine by Rhodococcus sp. strain NI86/21. Applied & Environmental Microbiology. 1995;61:2056-2060.

108. Best JA, Weber JB. Disappearance of s-triazines as affected by soil pH Using a balance-sheet approach. Weed Science. 1974;22:364-373.

29

109. Mandelbaum RT, Wackett LP, Allan DL. Mineralization of the s-triazine ring of atrazine by stable bacterial mixed cultures. Applied & Environmental Microbiology. 1993;59:1695-1701.

110. Pelizzetti E, Maurino V, Minero C, et al. Photocatalytic degradation of atrazine and other s-triazine herbicides. Environmental Science & Technology. 1990;24:1559-1565.

111. Curran WS, Loux MM, Liebl RA, Simmons FW. Photolysis of imidazolinone herbicides in aqueous-solution and on soil. Weed Science. 1992;40:143-148.

112. Penuela GA, Barcelo D. Comparative photodegradation study of atrazine and desethylatrazine in water samples containing titanium dioxide/hydrogen peroxide and ferric chloride/hydrogen peroxide. Journal of AOAC International. 2000;83:53-60.

113. Feakin SJ, Blackburn E, Burns RG. Biodegradation of s-triazine herbicides at low concentrations in surface waters. Water Research. 1994;28:2289-2296.

114. Klint M, Arvin E, Jensen BK. Degradation of the pesticides mecoprop and atrazine in unpolluted sandy aquifers. Journal of Environmental Quality. 1993;22:262-266.

115. Koskinen WC, Clay SA. Factors affecting atrazine fate in north central U.S. soils. Reviews of Environmental Contamination & Toxicology. 1997;151:117-165.

116. Seybold CA, Mersie W, McName C, Tierney D. Release of atrazine (14C) from two undisturbed submerged sediments over a two-year period. Journal of Agricultural & Food Chemistry. 1999;47:2156-2162.

117. Sinclair JL, Lee TR. Biodegradation of atrazine in subsurface environments. Environmental Research Brief. EPA/600/S-92/001. Washington, DC: U.S. Environmental Protection Agency; 1992.

118. Graymore M, Stagnitti F, Allinson G. Impacts of atrazine in aquatic ecosystems. Environment International. 2001;26:483-495.

119. Bronstein AC. Herbicides. In: Dart RC, ed. Medical Toxicology. Philadelphia, PA: Lippincott Williams & Wilkins; 2004:1525.

120. Keck G, Bostvironnois C, Braquet E. Veterinary toxicology of a model-herbicide - atrazine. Revue De Medecine Veterinaire. 1991;142:481-488.

121. Thorpe N, Shirmohammadi A. Herbicides and nitrates in groundwater of Maryland and childhood cancers: A geographic information systems approach. Journal of Environmental Science and Health Part C-Environmental Carcinogenesis & Ecotoxicology Reviews. 2005;23:261-278.

122. HSDB. Hazardous Substance Data Bank. National Library of Medicine, National Toxicology Information Program. Atrazine. Available at <http://toxnet.nlm.nih.gov/cgi-bin/sis/search/f?./temp/~tLjMv5:1> (Last accessed on 10 Nov 2008). Bethesda MD; 2004.

123. EPA. Ciba-Geigy Corporation. MRID No. 00141874, 00157875, 00158930, 40629302. HED Doc. No. 005940, 006937. Available from EPA. Washington, DC: U.S. Environmental Protection Agency; 1986.

124. RIVM. Re-Evaluation of Human Toxicological Maximum Permissible Risk Levels. Available at <http://www.rivm.nl/bibliotheek/rapporten/711701025.pdf> (Last accessed on Nov 22 2008). Netherlands: Rijksinstituut voor Volksgezondheid en Milieu (RIVM; National Institute of Public Health and the Environment, the Netherlands); 2001.

125. WHO. Atrazine in Drinking-water. Background document for development of WHO Guidelines for Drinking-water Quality. WHO/SDE/WSH/03.04/32. Available at

30

<http://www.who.int/water_sanitation_health/dwq/atrazinerev0305.pdf> (last accessed on Oct 18 2008). Geneva, Switzerland: World Health Organization; 2003.

126. Koeber R, Fleischer C, Lanza F, Boos KS, Sellergren B, Barcelo D. Evaluation of a multidimensional solid-phase extraction platform for highly selective on-line cleanup and high-throughput LC-MS analysis of triazines in river water samples using molecularly imprinted polymers. Analytical Chemistry. 2001;73:2437-2444.

127. Ghanem A, Bados P, Perreau F, et al. Multiresidue analysis of atrazine, diuron and their degradation products in sewage sludge by liquid chromatography tandem mass spectrometry. Analytical & Bioanalytical Chemistry. 2008;391:345-352.

128. Dagnac T, Bristeau S, Jeannot R, Mouvet C, Baran N. Determination of chloroacetanilides, triazines and phenylureas and some of their metabolites in soils by pressurised liquid extraction, GC-MS/MS, LC-MS and LC-MS/MS. Journal of Chromatography. A. 2005;1067:225-233.

129. Graziano N, McGuire MJ, Roberson A, Adams C, Jiang H, Blute N. 2004 National Atrazine Occurrence Monitoring Program using the Abraxis ELISA method. Environmental Science & Technology. 2006;40:1163-1171.

130. Rodriguez-Mozaz S, Reder S, de Alda ML, Gauglitz G, Barcelo D. Simultaneous multi-analyte determination of estrone, isoproturon and atrazine in natural waters by the RIver ANAlyser (RIANA), an optical immunosensor. Biosensors & Bioelectronics. 2004;19:633-640.

131. Denovan LA, Lu C, Hines CJ, Fenske RA. Saliva biomonitoring of atrazine exposure among herbicide applicators. International Archives of Occupational and Environmental Health. 2000;73:457-462.

132. Lioy PJ, Edwards RD, Freeman N, et al. House dust levels of selected insecticides and a herbicide measured by the EL and LWW samplers and comparisons to hand rinses and urine metabolites. Journal of Exposure Analysis and Environmental Epidemiology. 2000;10:327-340.

133. Lang D, Criegee D, Grothusen A, Saalfrank RW, Bocker RH. In vitro metabolism of atrazine, terbuthylazine, ametryne, and terbutryne in rats, pigs, and humans. Drug Metabolism & Disposition. 1996;24:859-865.

134. Gatidou G, Kotrikla A, Thomaidis NS, Lekkas TD. Determination of two antifouling booster biocides and their degradation products in marine sediments by high performance liquid chromatography-diode array detection. Analytica Chimica Acta. 2004;505:153-159.

135. Gatidou G, Zhou JL, Thomaidis NS. Microwave-assisted extraction of Irgarol 1051 and its main degradation product from marine sediments using water as the extractant followed by gas chromatography-mass spectrometry determination. Journal of Chromatography. A. 2004;1046:41-48.

136. Sabik H, Jeannot R, Rondeau B. Multiresidue methods using solid-phase extraction techniques for monitoring priority pesticides, including triazines and degradation products, in ground and surface waters. Journal of Chromatography a. 2000;885:217-236.

31

2.0 Pentachlorophenol

2.1 Physical Chemical Properties

2.1.1 Basic Properties

Table 2.1 Physical Chemical Properties of Pentachlorophenol and pentachlorophenate

Characteristic Pentachlorophenol Pentachlorophenate Synonyms: Chlon; Dowicide 7; Dowicide

EC-7; Dura Treet II; EP 30; Fungifen: Grundier Arbezol; Lauxtol; Liroprem; Penta Concentrate; Penta Ready; Penta WR; Permasan; Santophen 20; Woodtreatb

Dow Dormant Fungicide; Dowicide G; Dowicide G-St; Mystox D; Napclor-G; Santobrite; Sapco25 Weedbeads

CAS number: 87-86-5 131-52-2 Chemical class: Chemical name: pentachlorophenate;

2,3,4,5,6- Empirical formula: C6HCl5OH C6Cl5ONa Chemical structure

Molecular weight: 266.35 288.34 Melting point: 190 oC - Boiling point: 309–310 oC - Vapour pressure: 0.00011 mmHg (25 oC) Water solubility: 14 mg/L (20 0C) 330,000 mg/L (25 oC) Henry's law constant:

3.4x10-6 atm-m3/mol (at 25 oC)

Partition Coefficient (OW):

5.01

pKa 4.7

2.1.2 Brief Notes Pentachlorophenol (PCP) consists of a phenol (an aromatic ring with an alcohol functional group) plus 5 attached chlorine molecules (Table 2.1). As it is a highly halogenated (non-ionic) organic chemical, it is poorly poor water-soluble but has moderate to high solubility in organic (non-polar) solvents. In contrast, the sodium salt pentachlorophenate is highly water-soluble but is poorly soluble in organic solvents.

32

2.1.3 Uses PCP and its water-soluble sodium salt pentachlorophenate have been in commercial use since approximately 1936, when they were introduced as timber preservatives against wood termites.1 Since then, they have been used throughout the world as herbicides, disinfectants, molluscicides, algicides, and constituents of antifouling paints.2 At its peak, worldwide consumption was estimated to be approximately 90,000 tons per year.3 Until 1988, it was estimated New Zealand use peaked at approximately 200 tonnes per year for antisapstain treatment4 and 100 tonnes per year for wood preservation.5 Due to its toxicity, it has become a restricted–use pesticide.5 It has been recognized that PCP was regarded as the most commonly detected environmental contaminant for humans in the United States.6 Primarily it has been used as a wood preservative and 80% of its use is in the United States. Due to its high solubility in non-polar organic solvents, and the sodium salt’s high solubility in water, both forms were also widely used in insecticidal and fungicidal formulations for non-wood applications.7 Earlier formulations of PCP were contaminated with various dioxins, which may have partly contributed to host toxicity upon exposure. Formulations sold in New Zealand, for example, were shown to contain dioxin congeners at 0.9 mg I-TEQ/kg .8 Following toxic exposures to PCP, research has shown evidence of dioxins distributed in various body compartments.9 Dioxins are systematically reviewed in the dioxin section of this report.

2.2 Routes of Exposure Pentachlorophenol is a potent skin, eye, and upper respiratory tract irritant. It is rapidly and almost completely absorbed across the skin and is readily absorbed following inhalation and ingestion.10 However, almost all reports of human toxicity have occurred after skin absorption and to a lesser extent pulmonary absorption.

2.2.1 Ingestion To the knowledge of the author, there are only two case reports of the ingestion of PCP in humans.11, 12 Following the ingestion of PCP, there is no evidence of gastrointestinal injury, though there may be some evidence of localised epigastric tenderness. This pesticide is, however, rapidly absorbed across the gastrointestinal tract, causing systemic symptoms to occur within hours of ingestion11; Peak plasma levels have been reported within this time frame. The ingestion of PCP particles can also be contributed to by ciliary clearance from the respiratory tract.

2.2.2 Skin PCP is readily absorbed across skin. Such penetration is expected given its physicochemical properties; it has a high oil:water partition coefficient (5:1) and poor aqueous solubility (maximum dissolution is 14 mg/L), enabling it to readily partition across cellular lipid membranes and enter host circulation. Dermal contact with PCP can therefore lead to elevated plasma concentrations and subsequent host toxicity. Systemic toxicity can also result from airborne exposure or from contact with contaminated soil.13, 14 Dermal absorption has predominantly occurred as a result of

33

unintentional workplace exposure (most often as a result of direct contact), resulting in toxicity and in some cases death.15-20 From these studies, it is evident that there is a potential for considerable dermal absorption following inadvertent exposure to PCP, with consequential adverse effects on human health. Post-mortem analysis following dermal exposure, for example, has revealed high concentrations of PCP and its major metabolites in all major organs of the body, with notable cytotoxicity in organs where there is increased perfusion (CNS and kidney).21 For further details on this case report, see section 2.10.2 (case reports).

2.2.3 Eye Inflammation of the optic nerve has been reported following ocular exposure to PCP. Patients complained of impaired vision, but did not suffer other signs and symptoms typically associated with systemic toxicity.22 Upon examination, the patients’ optic nerve heads were slightly congested but not swollen.

2.2.4 Inhalation Although very little has been published on human exposure via inhalation of airborne PCP, it has been recognised as potentially the dominant route of chronic low-level exposure from the environment.23 PCP has a low vapour pressure, consequently little (if any) will be exhaled and the dose inhaled is either absorbed across the alveoli into the systemic circulation or is expelled via muco-ciliary extraction.

2.3 Toxicokinetics

2.3.1 Absorption PCP is readily absorbed from the gastrointestinal and respiratory tracts; dermal absorption does occur, but to a lesser extent. Bioavailability via the inhalation route has been estimated to be approximately 70 – 80%, as assessed in workers who were brushing PCP onto wood in an enclosed space.24 Following ingestion, absorption from the intestinal tract is rapid25 with peak plasma levels occurring within 4 hours. Bioavailability of PCP formulated with corn oil, has been calculated to range from 75 – 88%, depending on the administered dose.26 The rate of dermal absorption is slower than by oral absorption, but is greater when PCP is dissolved in diesel oil rather than in water.27 Percutaneous absorption from contaminated soil reported a mean bioavailability of 24% following 24 hours exposure.14 PCP can transfer across the placenta to the foetus; PCP metabolites have been detected in amniotic fluid collected (from pregnant mothers in the general population) during amniocentesis, suggesting foetal exposure may occur as early as 15–18 weeks gestation.28 Similar evidence has also been shown with PCP and two metabolites being present in umbilical cord plasma samples from mothers living on a diet of sea food containing elevated levels of PCP.29

2.3.2 Distribution Uhl and colleagues30 noted as much as 96% of the absorbed dose of PCP was bound to plasma protein in exposed humans. Tissue distribution following long-term low levels of exposure showed PCP (in decreasing order of concentration) in liver, kidney, brain, spleen and fat.31 Post-mortem analyses following fatal acute exposure to PCP

34

have shown bile and renal tissue containing the highest concentration followed by lung, liver and blood.9, 21 Before strict regulatory controls were placed on the use of PCP, members of the general population, who were not in direct contact with this fungicide, had evidence of free PCP in the urine, which could range from 0.003–0.57 µg/mL (average 0.044 µg/mL).32 Workers with substantial dermal and respiratory exposure could average as high as 5.6 µg/mL in the urine.24 More recent studies have shown urinary PCP levels in the general population have substantially dropped to less than or equal to 0.005 µg/mL and workplace exposure levels had reduced to 0.002 – 0.363 µg/mL.33-36

2.3.3 Metabolism The metabolism of PCP may undergo two stages of metabolism, predominantly within the liver. The first phase, undertaken by cytochrome P450 isozymes, may oxidise PCP to form tetrachloro-p-hydroquinone, and this metabolite (or the parent drug) may then conjugate with glucoronide to form the respective glucoronide complexes (see Figure 2.1).37, 38 Both free and conjugated products are then eliminated from the body. They have been demonstrated in the urine of both occupationally exposed workers39 and the general public.40

Figure 2.1 Brief summary of the hepatic oxidation and conjugation of PCP

2.3.4 Elimination Evidence of enterohepatic binding in primates suggests PCP is extensively eliminated via bile41; this, however, has not been confirmed in human studies,30 where elimination predominantly occurs via the kidneys. Clearance of PCP from this organ is slow (0.07 mL/min),30 possibly due to its high protein binding (>98%) in the blood. The elimination half-life of free PCP in healthy human volunteers has been calculated to be 30.2 ± 4 hours.25 Similar rates have been calculated from investigations on other species.42, 43 However, elimination profiles determined in other human studies, both chronic and acute, were longer. In one study, elimination profiles were determined in 18 workers during their 20 days of vacation; they demonstrated half-lives that were in excess of 30 days).44 Similar differences have also been found in other studies.30, 45 Despite the differences in reported half-lives, the consensus now appears to favour the longer half life of 20-30 days. Given the long half-life and reabsorption of PCP (and metabolites) by the kidneys, resultant urinary levels will represent exposure over several weeks rather that the

35

single and momentary absorption of PCP following administration. Thus, based on the elimination half life, the time to achieve a steady state (where elimination equals ingestion) in the human body burden has been calculated to be 3 months.30 Such a value is 10-20 times higher than that calculated value from animal models; this difference between animals and humans is important when determining the TDI of PCP (see section 2.10.4).

2.4 Mechanism of Action within the Body Pentachlorophenol (PCP) acts by uncoupling oxidative phosphorylation.46 PCP is a weak acid and is highly lipophilic, enabling it to readily migrate across the mitochondrial membrane. PCP alters adenosine triphosphate (ATP) formation by dissipating the proton gradient, which is the driving force for capturing energy in the ATP molecule. Consequently, the rate of electron transport is enhanced, leading to excess metabolic heat, producing fever in the host.47 PCP also inhibits other enzymes within the mitochondria, leading to mitochondrial and ultimately cellular damage.21 Such damage occurs predominantly in well-perfused organs that depend on oxygen, such as the heart and brain. Evidence of injury in fatal cases involving neonates and adults includes the vacuolisation of hepatocytes, proximal tubular cells and myocardial tissue.21, 48 PCP can also induce cytochrome P450 3A activity and therefore may influence the bioavailability of pharmaceutical drugs.49



2.5.1.1 Human (acute) Expected systemic symptoms following PCP poisoning are exceedingly high body temperature and profuse sweating.50 Nausea, vomiting, abdominal pain and intense thirst have also been described. Symptoms may progress to rapid pulse, heart failure, coma and subsequent death, which can occur within 3 to 30h following initial exposure. A rectal temperature of 42.8oC was recorded in one patient immediately prior to death.20 Patients may experience muscle twitching and jerking limbs,11 which may lead to convulsions20; rapid and marked rigor mortis is a common feature following death.15, 20, 21 Injuries as a result of direct sub-chronic exposure are clearly defined with death or serious injuries (similar symptoms to acute exposure) occurring as a result of ingestion or skin contact. Workers who have dipped their hands in PCP over 3 to 30 days have subsequently died.19 Similarly, six days skin contact with PCP powder has caused death in a worker who failed to wear appropriate protection.51

2.5.1.2 Animal (acute) Signs and symptoms related to animal exposure to PCP are associated with three routes of exposure (ingestion, inhalation and dermal) and duration of exposure (acute,

36

sub-chronic and predominantly chronic studies). They are all summarised together in section 2.5.2.2, according to the various organs.


2.5.2.1 Human (chronic) In contrast to acute exposure, symptoms following indirect chronic exposures leading to intoxication are not as well defined, with only limited evidence of a range of possible disease features. In a study of timber workers exposed to PCP, symptoms associated with chronic exposure were identified as chronic fatigue and neuropsychiatric complaints, involving subjective symptoms such as anxiety, insomnia, confusion, changes in behaviour and feelings of depression.52 However, the authors themselves concluded these were not symptoms characteristic of PCP exposure and there were a range of confounding variables that were identified. Two further studies have identified psychiatric complaints as possible symptoms of long-term exposure,53, 54 though in one of these investigations there was also evidence of lindane exposure.54 Various skin conditions that may occur upon exposure to PCP including chloracne (a skin disfiguration characterised by postulates located on the face, trunk and proximal extremities), increased risk of developing skin infections or other unspecified skin rashes55, 56 and furunculosis.16 Some workers experienced neuralgic pains in the lower limbs as well chloracne.56 The risk of various skin diseases following chronic exposure to PCP is related to the duration of exposure.57 However, a note of caution must be made with these investigations; earlier formulations of PCP were contaminated with various dioxins58 and the skin diseases resulting from exposure are now recognised as most likely been caused by these contaminants, rather than PCP. Employees exposed to PCP have complained of various subjective respiratory symptoms, including bronchitis and other respiratory tract inflammations.16, 56, 59Other investigators have suggested an association of PCP with a susceptibility to infertility,60 repeat abortions61 and possible minor changes to endocrine function.62 Biochemical evidence of hepatic function changes have been reported in a cluster of workers employed to manufacture PCP,63 though these symptoms may be more related to the PCP precursor (hexachloroclohexane) or dioxin contaminants. Reduced creatinine clearance and impaired phosphate reabsorption was reported in workers following chronic exposure to PCP and other wood preservatives.45

2.5.2.2 Animal (chronic, sub-chronic and acute)

2.5.2.2.1 Introduction

Acute, sub-chronic and chronic investigations have been conducted on a range of mammalian species, with orally administered doses of PCP, ranging from 1 to 600 mg/kg body weight. Changes to organ systems have included dose-related hepatic injuries, mild renal toxicity, changes in thyroid function, subtle changes to the immune response and evidence of mild neuro-chemical changes. However, doses administered to achieve these responses were orders of magnitude greater than expected concentrations assayed in recent dietary intakes of pesticides including PCP (reported dietary values ranged from 0.5 to 1.9 ng/kg body weight/day).64

37

2.5.2.2.2 Cardiovascular, Haematology, Metabolic and Respiration

Rats administered sub-acute and sub-chronic doses of PCP (typically 10–30 mg/kg/day) developed extensive vascular damage and heart failure in rats, rabbits, pigs, and dogs. There are conflicting results on changes to haematological parameters such as packed cell volume, leukocyte and erythrocyte concentrations following ingestion of PCP.65-67 Elevated glucose and decreased hepatic glycogen have been reported following sub-chronic exposure (40 mg/kg/day).68


In experimental animals (rats, mice, pigs) the liver is a target organ for PCP-induced toxicity. There is evidence of elevated transaminases, increased liver weight and histopathological changes following sub-acute (41 mg/kg/day) and sub-chronic (1–40 mg/kg/day) oral exposures (e.g 66, 69-71). Increasing sub-chronic concentrations of administered PCP (10, 30 and 60 mg/kg oral gavage) led to increased severity of liver damage such as necrosis, periportal fibrosis, and hepatocellular degeneration.72 Sub-chronic administration of PCP (typically 10–30 mg/kg/day) may result in mild to moderate renal toxicity; there was evidence of increased organ weight and altered enzyme levels whereas histopathologic effects were rarely seen.65


Substantial changes in thyroid hormone levels (mink, rats, sheep) were observed following some investigations of sub-acute, sub-chronic and chronic exposures (1 – 30 mg/kg/day). 73-76 Increases in sheep insulin have also been observed (2 mg/kg/day).74 Sub-chronic studies suggest PCP may decrease fertility, although the mechanism does not appear to be through histological damage to reproductive tissue (doses ranging from 1 to 60 mg/kg/day) (e.g. rats72).

2.5.2.2.5 Musculoskeletal, Dermal and Body Weight

Chronic dose administrations to rats (400-600 mg/kg/day) may lead to weight loss.71,

77


Doses ranging from 0.5–100 mg/kg/day in mice have been shown to affect humoral and cellular immunity, susceptibility to tumor induction, and complement activity. Although the earlier studies may have been influenced by the presence of dioxin contaminants,78-80 more recent evidence still suggests the pure compound may change the immune response.70


PCP adversely affects the central nervous system, possibly due to hyperthermia induced by uncoupling of oxidative phosphorylation (for further detail on this mechanism, see section 2.4.0 - Mechanism of Action within the Body). Neurochemical evidence includes changes of some rat brain enzymes and decreased glial glutathione levels following chronic exposure (20 mg/L in drinking water).81 Changes to the morphology of rat peripheral neural cells has been reported following chronic administration (38 mg/kg/day and 120 mg/kg/day for 90 and 120 days respectively).82 However, there is a suggestion there may have been dioxin contaminants present.

38

2.6 Carcinogenicity Based on sufficient animal data and on data that is limited in humans, both the International Agency for Research on Cancer (IARC) and the U.S. Environmental Protection Agency (EPA) have classified PCP as a “possible human carcinogen” The EPA classified it into group B2 (probable human carcinogen) and the IARC in 2B (possibly carcinogenic to humans).83, 84 In their judgment on PCP, IARC noted that the most consistent findings among the studies available at that time were associated with non-Hodgkin lymphoma and soft-tissue sarcoma.

2.6.1 Human Studies Investigations on people who consumed fish from a lake contaminated with a range of chlorphenols including PCP, and who drank water from nearby wells, had a significant association of non-Hodgkin’s lymphoma with fish consumption, when compared with other control populations.85 A cohort of sawmill workers (n = 27,464 men) who were employed between 1950 and 1995 demonstrated an associated increased incidence of non-Hodgkin's lymphoma, multiple myeloma, and kidney cancer following chronic exposure to PCP.86 Concerns were raised that contaminations of dioxins in the formulations may have contributed to these cancers. However, a study preceding this investigation showed no evidence of the carcinogenic causing congener 2,3,7,8-tetrachlorodibenzodioxin being present in these fungicides.87 Non-Hodgkin’s lymphoma has been reported as significant in several other studies involving PCP and other chlorophenols.88-90 In addition, soft-tissue sarcoma has also been reported.88, 90-92

2.6.2 Animal Studies In animal studies carcinogenicity due to orally administered PCP has been tested in at least four major studies using rats and mice with various grades of PCP.93-96 From these studies it was concluded PCP in an animal model might cause hepatocellular adenoma or carcinoma, mesotheliomas and nasal squamous cell carcinomas.

2.7 Mutagenicity Evidence of damages to chromosomes in workers exposed to PCP has been variable with one report suggesting a marginal increase in chromosomal aberrations97 whereas other studies did not find significant increases. 98, 99 Sister chromatid exchange was not increased in lymphocytes derived from these workers. An increase in DNA adduct formation was noted in mice liver but not kidney or spleen.69, 100 With the exception of one study,101 there is substantial evidence to suggest PCP does not cause gene mutations in bacteria (e.g.102-104). With mammalian cells, in vitro, weak chromosomal changes have been observed,105 but no significant increases in DNA damage have been reported.106-108

39


2.8.1 Human Studies Case reports suggest exposure to PCP may lead to an increased risk of miscarriages in humans. Two women experienced a series of miscarriages following exposure to PCP vapour from treated furniture in their houses.61 Serum assays demonstrated elevated concentrations of PCP (62 µg/mL and 31 µg/mL) that declined following removal of the source of PCP. Both women later had successful pregnancies. Female employees who worked at day-care centres that contained wood treated with PCP, had a higher rate of miscarriage than those in centres with untreated timber (14.3% versus 8.7%).109 There is little evidence relating to paternal occupational exposure to chlorophenol fungicides (including PCP) and subsequent childhood cancers amongst children of sawmill workers who were exposed to these agents during the course of their work.110

2.8.2 Animal Studies Recent evidence from animal studies suggests that PCP exposure during gestation can result in foetal/neonatal mortality, malformations, decreased growth, and possibly functional deficits in rats and rabbits. No developmental effects have been observed in rabbits when pure doses were administered (up to 30 mg/kg/day) on gestational days 6–18.111 Embryo lethality, post-implantation resorptions, decreases in litter size and decreases in neonatal survival have occurred across a range of studies with doses at or greater than 30 mg/kg or when lesser doses were given (that may contain impurities) (e.g. 111, 112). Similar doses are also necessary to induce malformations as reported in the above studies. Doses necessary to cause decreases in pup weight have been reported to be lower (1-14 mg/kg/day) (e.g. 111, 112).

2.9 Environmental Fate (Metabolism and Accumulation) and Subsequent Ecotoxicity


Air

Very few studies have assessed the ambient particulate air concentrations of PCP. Cautreels and colleagues113 measured air at an elevated altitude (5200 m) near La Paz, Bolivia and in residential areas of Antwerp, Belgium. Atmospheric concentrations were reported to contain 0.25 - 0.93 ng/m3 and 5.7 - 7.8 ng/m3 air, respectively. Reported values in Switzerland were more variable but were of the same order of magnitude (0.9 - 5.1 ng/m3).114 More recent studies have shown similar airborne values of PCP,115 despite a substantial worldwide reduction in its use. In contrast, airborne levels in association with occupational exposure can be an order of magnitude greater.98, 116 One building that was of recent construction with treated timber had been calculated to release substantially high airborne concentrations of 1-160 µg PCP/m3 air (within the building).117 Eight years after construction, PCP was still evident in the house, with measured values of about 0.4 µg/m3.

40

Water

Reported levels in fresh water have varied considerably, depending on pollution that may be sourced from sewerage or industrial outlets or studies conducted before worldwide restrictions of PCP were enforced. In one study, for example, wood-treating factories have contributed substantial amounts of PCP to the water burden; reported concentrations have ranged from 49 to 10,500 µg /L,118 though the latter value was from direct measurement at the sawmill itself. However, the majority of the reports measuring PCP in water contained concentrations less than 10 µg/L, with most below 1 µg/L.119 In New Zealand, samples assayed from a stream in the vicinity of an old sawmill were 3.62 µg /L, whereas in a large water body, where nine wood processing facilities occurred, concentrations didn’t exceed 0.04 µg/L.120 Other measurements in remote regions of the country led the authors to conclude PCP was not a widespread contaminant in New Zealand.

Soil

PCP has been measured in soils from various locations, but the highest reported concentrations were within soils in the vicinity of wood processing plants. Soil in the vicinity of a Swiss PCP-producing facility (Dynamit Nobel), for example, contained elevated levels of PCP that increased with increasing depth of soil.114 Surface layers, measuring 0 - 10 cm deep contained between 25 and 140 µg PCP/kg (dry weight) whereas those at 20-30 cm depth, measured from 33 to 184 at µg/kg. Leaching from timber previously treated with PCP has also led to soil contamination. The concentrations of PCP in soil measured at distances of 2.5, 30.5, and 152.5 cm from PCP-treated poles were 658, 3.4, and 0.26 mg/kg, respectively.116

2.9.2 Environmental Fate As a consequence of its widespread use, PCP has caused widespread environmental contamination; it has been found extensively in air, water and soil. PCP is released into the environment through its application as a pesticide and during its production; other sources of this halogenated chemical are from the chlorination of organic matter during water treatment and during industrial bleaching procedures.2, 121 Furthermore, other chemicals including hexachlorobenzene, pentachlorobenzene, and benzene hexachloride isomers will also metabolise to pentachlorophenol, thereby exacerbating environmental PCP levels. Pentachlorophenol is volatilised from treated wood surfaces. One investigation, for example, estimated as much as 30 to 80% of PCP applied to wood products evaporated into the atmosphere within 12 months after application.122 It can also leach from treated wood into surrounding soil.116 Volatilisation from soil and water is, however, negligible.123, 124 Upon release into the environment, PCP will adsorb and persist in soils and sediment, predominantly when conditions are acidic125 or organic matter is present.126 In contrast, PCP becomes more mobile when environmental conditions are neutral or basic.127 This leads to more extensive environmental contamination.2, 121 Moderate bioaccumulation does occur with PCP (PCP has a bioconcentration factor of <1,000) but there is no evidence of biomagnification in the food chain.128 Chemicals with a high octanol-water coefficient (greater than 4.0) are likely to bioaccumulate in organisms and food chains.129 The un-ionised form of PCP has a

41

coefficient of 5.01 suggesting this may occur. However, given environmental and physiological pH values are greater than PCP’s pKa (eg physiological pH is typically 7.4), PCP will predominantly remain ionized. Upon uptake and distribution within aquatic species, the ionic form of PCP will readily be eliminated from the organism.130-132 Consequently, PCP has minimal bioaccumulation capacity. Increasing concentrations of PCP in terrestrial or aquatic environments has not been observed.133 In recent years, however, there is evidence that decreased use of PCP as a pesticide has also correlated with diminishing PCP blood concentration in the general public; one study reported declines from 70 ng/mL in the 1980s to ~12 ng/mL in 1996.134 Indeed, a recent FDA diet study has shown total mean dietary intakes of PCP remain at or below 1 ng/kg body weight/day.64 Remaining elevated blood levels of PCP in some populations are most likely a reflection of other metabolised organochlorines (e.g. entachlorobenzene and phexachlorobenzene) that are sourced from their respective environments.135

2.9.3 Ecological Degradation Rates of degradation in soil, air and water have been reported, but values are variable. A brief summary of the literature suggests the following mean half-lives: Half-life in soil: 2-4 weeks128 Half-life in water: 3.5 – 100 hours136 Half-life in air: 58 days128, 137

PCP is resistant to hydrolysis and oxidation.136 In surface water, it is predominantly degraded by photolysis and to a lesser extent biodegradation; biodegradation typically occurs at water-rock or -macrophyte surfaces or in surface sediments,124 Atmospheric PCP undergoes both photolysis and then free radical oxidation. Both aerobic and anaerobic bacteria are responsible for the metabolism of PCP in soil.138 A wide variety of these bacterial species dechlorinate PCP to various forms ranging from tetrachlorophenol to phenol as well as chlorinated dibenzodioxins.139 Rates of soil degradation and chemical degradative pathways are influenced by organic matter, the pH and the redox potential of the soil.139, 140


2.10.1 Doses Necessary to Cause Acute and Chronic Toxicity

2.10.1.1 Acute Animal LD50 studies, as summaries in section 2.10.3.1), suggests PCP is a compound of moderate to high acute toxicity. Acute human exposure to PCP, either by ingestion or direct dermal contact, can result in serious toxicity that may lead to death, in the absence of medical intervention. One patient ingested approximately 170 mL of a solution containing 12% PCP (approximately 20.4 grams PCP ingested). He suffered hypertension, hyperthermia, tachycardia and profuse sweating. Serum levels peaked at 153 µg/mL, 5 hours post-

42

ingestion. Following decontamination and aggressive diuresis, the patient made an eventual recover.11 A brief unintentional dermal exposure to a high concentration of PCP solution (200 mg/mL) caused exfoliation of the epidermal layer of the hand following skin contact141; recovery occurred after 5 days post-exposure. Skin contact for ten minutes with PCP (4 mg/mL) caused immediate pain and erythema. After two and four days following exposure, the patient had elevated concentrations of PCP in the urine at 0.236 and 0.080 µg/mL respectively. In contrast to the above case reports, controlled clinical trials with human volunteers have demonstrated no observable clinical effects. Volunteers were orally administered 0.1 mg/kg sodium pentachlorophenate,25 and single doses totalling 3.9, 4.5, 9 and 18.8 mg PCP.30

2.10.1.2 Chronic Doses necessary to cause PCP toxicity in animals can vary according to the species, the target organ being investigated, and the duration of exposure. Such summaries, separated by organ systems, following chronic exposure are included in section 2.5.2.2. Essentially, toxicities linked with chronic exposure ranged from oral doses of 1 mg/kg/day (reduced serum thyroxine levels)76 to 60 mg/kg/day (reduced body weight)77 Chronic direct contact with PCP can also result in toxic effects and in some instances has resulted in death of the patient. It has been estimated that the lowest dose necessary to cause human death is approximately 1 gram (17 mg/kg).142 Deaths have occurred when adult males dipped their hands into 1.5 – 2% PCP v/v solutions over an extended period of time (ranging from 3 to 30 days).19 In another fatality, death occurred with a worker who was diluting a 40% v/v concentrate without any skin protection for 6 days.51 Four families who used water from the same well displayed signs and symptoms including fever, weakness, irritation to the throat and flushed faces. It was discovered PCP had leached into the well from adjoining fields four days previously; the concentration in the well was ~12.5 mg/L. Signs and symptoms resolved after they ceased using this water. The calculated adult daily oral dose was 0.42 mg/kg/day. In contrast to direct contact with the pesticide itself, there has been ongoing concern regarding chronic exposure to low levels of PCP that have remained in the environment and especially for those who have been exposed to elevated concentrations, either in the workplace or in wooden houses that have been treated with this pesticide. There is sufficient evidence that increased plasma concentrations occur in people exposed to elevated levels of environmental PCP; the results from one comprehensive study, examining serum levels in workers exposed to this agent, residents of wooden houses that had previously been treated with PCP, and a control population, are summarised in Table 2.2.143 Though residents living in wooden houses had raised PCP blood levels as a result of exposure, an earlier paper on the studied group reported no significant health problems attributable to PCP.144 Nevertheless, on the basis of this paper, the U.S. Environmental Protection Agency issued a position document proposing regulations to reduce the human health risks by banning the treatment of logs with PCP before construction.145

43

Table 2.2 Summary of measured serum PCP concentrations of various sub-populations after possible chronic exposure to elevated levels of PCP143

Studied Group Description Mean Serum concentration (ng/mL)

Residents of log homes Children (2-7) 610 Youth (8-15) 390 Adults (over 15) 350 Workplace exposure Log home construction 83 Telephone line maintenance 110 Log museum 450 Wood preservation 490 Chemical packaging 57,900 General Population Children & youths (1-19) 3.4 Adults (over 15) 40

A recent study in New Zealand assessing sawmill workers with a history of PCP exposure, suggested a link with health problems (such as non-cancerous respiratory disease, fevers, recurrent nausea and diarrhea, heart palpitations) but there was no evidence of elevated mortality rates or a statistically significant increased risk of cancers.146 Studies of workers chronically exposed to PCP (for more than ten years) suggests there may be a weak effect on the immune response and liver function; however, subjects themselves had no clinical evidence of such diseases.34 A cohort study assessing Dow Chemical Company workers exposed to PCPs concluded there was a lower than expected total cancer mortality than compared to general population.147 There is, however, an association with non-Hodgkins lymphoma (see section 2.6).85 Workers exposed to wood treated with chemicals had significantly elevated urine (174 ppb) levels of PCP but a comprehensive clinical assessment failed to find significant differences compared to a control group.148 Elevated serum levels (in excess of 10 ppb) have also been correlated with immunological abnormalities,53 linked to a possible association with hormonal or immunological changes in women with repeat miscarriages60 and temporary reduced creatinine clearance and impaired phosphate reabsorption.45

2.10.2 Summary of Case Reports Following Toxic Exposure There are numerous reports of toxic exposure to PCP following inhalation and dermal absorption.15-21, 48 Two case reports are briefly described. A 33-year-old male worked in a chemical plant breaking up large blocks of PCP. He had no personal protection against the inhalation of dust or the absorption across the skin; the workplace environment was described as dusty and the worker was observed having powder on his clothes. After three weeks he was admitted to hospital suffering from hyperthermia (100 oF); soon after his temperature increased to 105oF and he became comatosed and later died. Analysis of his blood showed elevated concentrations of PCP (see Table 2.3). Light microscopy revealed evidence of swollen mitochondria.21

44

Table 2.3 Post-mortem analysis of body fluids and tissue samples as evidence of PCP intoxication21

Organ or fluid sample PCP Concentration (µg/mL) Blood 162 Gastric contents 8.2 Urine 29 Kidney 639 Liver 52 Lung 116 Bile 1130

Over a time interval of five months in 1967, nine neonates in a nursery developed signs and symptoms consistent with PCP toxicity (predominantly sweating and fever). The hospital laundry had been using PCP as an anti-mildew agent. Consequently, two infants died and another six had evidence of serious toxicity.48, 149 One patient had a serum level of 118 µg/mL at the time of their illness.

2.10.3 Measures of Mammalian Toxicity

2.10.3.1 LD50 values For an explanation of LD50 values see Appendix one. A summary of various LD50 values from a range of mammalian species and routes of exposure are summarised in Table 2.4. Table 2.4 Summary of single dose LD50 values for PCP10

Species Route LD50 (mg/kg/bw) Rat Oral 78 Rat Oral 184 Rat Oral 146 Rat Oral 175 Rat Oral 211 Rat Dermal 96 Rat Dermal 320 Rat Dermal 330 Rat Subcutaneous 66 Rat Intraperitoneal 34 Mouse Oral 130 Mouse Oral 74 Mouse Oral 177 Mouse Oral 117 Mouse Intraperitoneal 59 Mouse Intraperitoneal 61 Rabbit Oral 70 - 300 Rabbit Dermal 40 Rabbit Subcutaneous 70 - 100 Rabbit Intravenous 22 - 23

45

2.10.4 Guidance Values Defined guidance values, as derived by various regulatory authorities, are described in table 2.5.

Tolerable Daily Intakes Table 2.5 Tolerable daily intakes for pentachlorophenol, as defined by various regulatory authorities

Agency EPA RIVM ATSDR

Type of Exposure Chronic ingestion RfD

Chronic ingestion TDL

Chronic ingestion MRL

Value 30 µg/kg/day 3 µg/kg/day 1 µg/kg/day

Year 1993 1999/2000 2001

Species rat mink mink

Critical effect

Reduced body weight

Reduced thyroxine concentration

Reduced thyroxine concentration

NOAEL/LOAEL

3 mg/kg/day NOAEL

1 mg/kg/day LOAEL

1 mg/kg/day LOAEL

Uncertainty factor 100(10A, 10H) 300(10A,

10H,3D) 1000(10A, 10H, 10L)

Principle study 93 76 76

Other Guideline Values

Based on the possible health risks and exposure to the general population, the EPA has set PCP Maximum Contaminant Level Goal in drinking water at zero (see Appendix 2); because this measure is realistically unattainable, they have defined the Maximum Contaminant Level at 0.001 ppm (see Appendix 3). The maximum residue limits in food derived from plants in Germany must not exceed 0.01-0.03 mg/kg (IPCS, 1989). The current provisional guideline value of 9 µg/litre in drinking water (used for human consumption), has been set by the WHO.150 The Minimal Risk Levels (MRL) for PCP associated with oral exposure are presented in Table 2.6 (levels for inhalation have yet to be established). These values are very similar to the ADI value described above. For an explanation of MRL values see Appendix one; a complete summary of values is presented in Appendix four.

46

Table 2.6 A summary of established MRL levels for PCP by ingestion, as recommended ATSDR

Agency ATSDR ATSDR ATSDR

Type of Exposure Acute ingestion MRL

Sub chronic ingestion MRL

Chronic ingestion MRL

Value 5 µg/kg/day 1 µg/kg/day 1 µg/kg/day

Year 2001 2001 2001

Species rat mink mink

Critical effect Developmental Reproduction Endocrine

NOAEL/LOAEL

5 mg/kg/day LOAEL

1 mg/kg/day LOAEL

1 mg/kg/day LOAEL

Uncertainty factor 1000(10A, 10H, 10L)

1000(10A, 10H, 10L)

1000(10A, 10H, 10L)

Reference 128 128 128

In regard to the treatment of timber, the Code of Federal Regulations (Code of Federal Regulations: 21CFR178.3800) has permitted the use of PCP concentrations not exceeding 50 ppm.

2.11 A Summary of Biological Assays That are Available to Test for Evidence of These Chemicals Gas chromatography with electron-capture detection (GC-ECD) has been the most sensitive and commonly used method to assay chlorinated compounds including PCP.153-155 GC-ECD has been employed to assay PCP in serum,135, 151 plasma134 and urine.134 To improve sensitivity when assaying biological fluids, samples are typically derivatised with reagents containing halogens making analysis by electron-capture detection more readily detectable.152 Other form of detection have included mass spectrometry (e.g. 153), supercritical fluid chromatography154 and capillary electrophoresis.155 Samples requiring analysis are usually complexed in organic matrices (ranging from micro-organisms to human body fluids); they must first be isolated from these media before they can be assayed. Typically, sample preparation involves either organic extraction (e.g. 156) or use of solid phase cartridges (e.g. 157). The elimination half-life of PCP from the body is 20-30 days (see section 2.3.4). Recent exposure to this pesticide can be determined by blood or urine analysis for both the parent compound and its main metabolites. Gas chromatography would be the most appropriate method of analysis. However, since PCP degradation in the soil has been rapid, with half-lives between two to four weeks upon deposition in the soil (see section 2.9.3), any recent exposure to historic contaminants within the soil is unlikely. Another approach to assessing for PCP is assaying for the octa-dioxin congeners that can be present as contaminants in the PCP formulation and may persist in the environment and remain within the body following exposure due to their long half-lives. This approach has been used by AsureQuality Limited, who are the most suited laboratory in New Zealand to undertake such analysis.

47

2.12 Summary Pentachlorophenol (PCP) is a highly chlorinated phenolic pesticide that was widely used as a timber preservative until the mid 1980s. As a result of its toxicity, and its persistence in the environment, it was subsequently banned. Due to its lipophilic nature, PCP will readily absorb across the skin, respiratory and pulmonary tracts. Most incidents of human toxicity have occurred as a result of dermal contact. Workplace exposure due to direct contact with PCP has resulted in elevated plasma levels leading to serious toxicity and death. Bioavailability is high (70-85%) following inhalation or ingestion, but is lower via the dermal route. PCP readily crosses the placenta to the foetus. Upon absorption, PCP is bound to plasma proteins and distributed throughout the body. Metabolism is slowly eliminated in humans, occurring predominantly via the kidneys; half-life following acute exposure is typically 15 days. PCP acts on the body by uncoupling oxidative phosphorylatin leading to excessive metabolic heat and subsequent hyperthermia; other symptoms following acute exposure may include tachycardia, coma, convulsions, heart failure that may lead to death. Deaths have also been reported following direct sub-chronic exposure. In contrast, signs and symptoms associated with chronic exposure to lower levels of PCP are not as well defined. Reported symptoms have ranged from non-specific neurological complaints, choracne, pain, subjective respiratory symptoms, susceptibility to infertility, spontaneous abortions and subtle changes to endocrine function, and some subtle biochemistry parameters. As these reports are typically retrospective, the authors have cautioned their interpretations because of other confounding variables.52, 54, 58, 63 Administered oral doses to various animals have demonstrated injuries to various organs including hepatic and renal injury, mild neurochemical and thyroid effects, as well as subtle immunological changes. Doses necessary to achieve these morbidities are orders of magnitude greater that expected concentrations both in the workplace and in the general population. Nevertheless, in the absence of symptoms, patients exposed to high doses (such as residents of PCP-treated log homes) may still have concerning elevated serum and urine concentrations. Reported acute LD50 values that can vary, depending on the route of exposure and species investigated, from 22 to 330 mg/kg body weight; NOEL values are lower but are reported at similar orders of magnitude to the LD50 studies. From the NOEL levels, TDI values have been established, which vary, depending on the regulatory authority and the NOEL study that underpins the respective value. The TDI allows a defined intake of PCP per day for the duration of a subject’s life without appreciable risk. PCP is regarded as a possible human carcinogen. Evidence from cohort studies suggests there may an association between elevated chronic exposure to PCP and non-Hodgkin lymphoma and soft tissue sarcoma. In animal studies, research suggests evidence of hepatocellular carcinoma, mesotheliomas and nasal squamous cell carcinomas. Evidence for PCP as a mutagenic agent is inconclusive. Exposure to elevated PCP may lead to increased incidence of miscarriages in humans and teratogenic effects in animal studies.

48

Widespread use of PCP has resulted in the contamination of air, water and soil, though with more recent decreased use, corresponding environmental (and biological) levels have also decreased. Though PCP will be bioaccumulated in a wide range of species it does not biomagnify within the food chain. The environmental fate of PCP will depend where PCP has complexed; degradative half-lives are estimated to be 2-4 weeks, 48 hours and 58 days in soil, water and air respectively, though these values can vary greatly.

2.13 References 1. Tomlin CDS. The Pesticide Manual. 13 ed. Alton, Hanpshire: British Crop

Protection Council; 2003. 2. Ahlborg UG, Thunberg TM. Chlorinated phenols: occurrence, toxicity,

metabolism, and environmental impact. Critical Reviews in Toxicology. 1980;7:1-35.

3. Detrick RS. Pentachlorophenol: possible sources of human exposure. Forest Products Journal. 1977;27:13-16.

4. Shaw CP. Pentachlorophenol and Chlordane: Timber Preservation Chemicals in the New Zealand Environment. Report prepared by the Cawthron Institute, Nelson. Wellington: Ministry for the Environment; 1990.

5. Bingham AG. National Task Group on Site Contamination, Task Brief 6 – Background Paper on PCP Related Issues. NECAL service report S92/465. Wellington, New Zealand: Department of Health; 1992.

6. Kutz FW, Cook BT, Carter-Pokras OD, Brody D, Murphy RS. Selected pesticide residues and metabolites in urine from a survey of the U.S. general population. Journal of Toxicology & Environmental Health. 1992;37:277-291.

7. Mannsville Chemical Products Synopsis: Pentachlorophenol. Asbury Park, NF; 1987.

8. Bingham AG. PCDD and PCDF Impurities in Sodium Pentachlorophenate based Antisapstains. NECAL Service Report S91/328. Wellington, New Zealand: Department of Health; 1991.

9. Ryan JJ, Lizotte R, Lewis D. Human-tissue levels of PCDDs and PCDFs from a fatal pentachlorophenol poisoning. Chemosphere. 1987;16:1989-1996.

10. Gasiewicz TA. Nitro Compounds and Related Phenolic Pesticides. In: Hayes WJ, Laws ER, eds. Handbook of Pesticide Toxicology. San Diego: Academic Press Inc; 1981:1206-1217.

11. Haley TJ. Human poisoning with pentachlorophenol and its treatment. Ecotoxicology & Environmental Safety. 1977;1:343-347.

12. Cretney MJ. Pentachlorophenol death. Bulletin of the International Association of Forensic Toxicologists. 1976;12:10.

13. Qiao GL, Brooks JD, Riviere JE. Pentachlorophenol dermal absorption and disposition from soil in swine: effects of occlusion and skin microorganism inhibition. Toxicology & Applied Pharmacology. 1997;147:234-246.

14. Wester RC, Maibach HI, Sedik L, Melendres J, Wade M, DiZio S. Percutaneous absorption of pentachlorophenol from soil. Fundamental & Applied Toxicology. 1993;20:68-71.

15. Bergner H, Constant.P, Martin JH. Industrial pentachlorophenol poisoning in Winnipeg. Canadian Medical Association Journal. 1965;92:448-451.

16. Baader EW, Bauer HJ. Industrial intoxication due to pentachlorophenol. Industrial Medicine & Surgery. 1951;20:286-290.

49

17. Mason MF, Wallace SM, Foerster E. Pentachlorophenol poisoning: report of two cases. Journal of Forensic Sciences. 1965;10:136-147.

18. Gordon D. How dangerous is pentachlorphenol? The Medical Journal of Australia. 1956;43:485-488.

19. Menon JA. Tropical hazards associated with the use of pentachlorophenol. British Medical Journal. 1958;1:1156-1158.

20. Wood S, Rom WN, White GL, Logan DC. Pentachlorophenol poisoning. Journal of Occupational and Environmental Medicine. 1983;25:527-530.

21. Gray RE, Gilliland RD, Smith EE, Lockard VG, Hume AS. Pentachlorophenol intoxication: report of a fatal case, with comments on the clinical course and pathologic anatomy. Archives of Environmental Health. 1985;40:161-164.

22. Jindal HR. Bilateral retrolobullar neuritis due to insecticides. 1968;44:341-342. 23. Wilson NK, Chuang JC, Morgan MK, Lordo RA, Sheldon LS. An observational

study of the potential exposures of preschool children to pentachlorophenol, bisphenol-A, and nonylphenol at home and daycare. Environmental Research. 2007;103:9-20.

24. Casarett LJ, Bevenue A, Yauger WL, Jr., Whalen SA. Observations on pentachlorophenol in human blood and urine. American Industrial Hygiene Association Journal. 1969;30:360-366.

25. Braun WH, Blau GE, Chenoweth MB. The metabolism/pharmacokinetics of pentachlorophenol in man, and a comparison with the rat and monkey. Developments in Toxicology and Environmental Science. 1979;7:289-296.

26. Pu X, Carlson G, Lee L. Oral bioavailability of pentachlorophenol from soils of varying characteristics using a rat model. Journal of Toxicology & Environmental Health Part A. 2003;66:2001-2013.

27. Horstman SW, Rossner A, Kalman DA, Morgan MS. Penetration of pentachlorophenol and tetrachlorophenol through human-skin. Journal of Environmental Science and Health Part A-Environmental Science and Engineering & Toxic and Hazardous Substance Control. 1989;24:229-242.

28. Bradman A, Barr DB, Claus Henn BG, Drumheller T, Curry C, Eskenazi B. Measurement of pesticides and other toxicants in amniotic fluid as a potential biomarker of prenatal exposure: a validation study. Environmental Health Perspectives. 2003;111:1779-1782.

29. Sandau CD, Ayotte P, Dewailly E, Duffe J, Norstrom RJ. Pentachlorophenol and hydroxylated polychlorinated biphenyl metabolites in umbilical cord plasma of neonates from coastal populations in Quebec. Environmental Health Perspectives. 2002;110:411-417.

30. Uhl S, Schmid P, Schlatter C. Pharmacokinetics of pentachlorophenol in man. Archives of Toxicology. 1986;58:182-186.

31. Grimm HG, Schellmann B, Schaller KH, Gossler K. Pentachlorphenolkonzentrationen in Geweben und Korperflussigkeiten von Normalpersonen. Zentralblatt fur Bakteriologie, Mikrobiologie und Hygiene - 1 - Abt - Originale B, Hygiene. 1981;174:77-90.

32. Bevenue A, Wilson LJ, Casarett LJ, Klemmer HW. A survey of pentachlorophenol content in human urine. Bulletin of Environmental Contamination & Toxicology. 1967;2:319-332.

33. Pekari K, Luotamo M, Jarvisalo J, Lindroos L, Aitio A. Urinary excretion of chlorinated phenols in saw-mill workers. International Archives of Occupational & Environmental Health. 1991;63:57-62.

50

34. Colosio C, Maroni M, Barcellini W, et al. Toxicological and immune findings in workers exposed to pentachlorophenol (PCP). Archives of Environmental Health. 1993;48:81-88.

35. Treble RG, Thompson TS. Determination of pentachlorophenol in commercially prepared lyophilized human urine control samples. Chemosphere. 1996;33:693-697.

36. Meissner T, Schweinsberg F. Pentachlorophenol in the indoor environment: evidence for a correlation between pentachlorophenol in passively deposited suspended particulate and in urine of exposed persons. Toxicology Letters. 1996;88:237-242.

37. Reigner BG, Gungon RA, Hoag MK, Tozer TN. Pentachlorophenol toxicokinetics after intravenous and oral administration to rat. Xenobiotica. 1991;21:1547-1558.

38. Ahlborg UG, Larsson K. Metabolism of tetrachlorophenols in the rat. Archives of Toxicology. 1978;40:63-74.

39. Edgerton TR, Moseman RF. Determination of pentachlorophenol in urine: the importance of hydrolysis. Journal of Agricultural & Food Chemistry. 1979;27:197-199.

40. Juhl U, Witte I, Butte W. Metabolism of pentachlorophenol to tetrachlorohydroquinone by human liver homogenate. Bulletin of Environmental Contamination & Toxicology. 1985;35:596-601.

41. Rozman T, Ballhorn L, Rozman K, Klaassen C, Greim H. Effect of cholestyramine on the disposition of pentachlorophenol in rhesus monkeys. Journal of Toxicology & Environmental Health. 1982;10:277-283.

42. Braun WH, Young JD, Blau GE, Gehring PJ. The pharmacokinetics and metabolism of pentachlorophenol in rats. Toxicology & Applied Pharmacology. 1977;41:395-406.

43. Braun WH, Sauerhoff MW. The pharmacokinetic profile of pentachlorophenol in monkeys. Toxicology & Applied Pharmacology. 1976;38:525-533.

44. Bevenue A, Emerson ML, Casarett LJ, Yauger WL, Jr. A sensitive gas chromatographic method for the determination of pentachlorophenol in human blood. Journal of Chromatography. A. 1968;38:467-472.

45. Begley J, Reichert EL, Rashad MN, Klemmer HW. Association between renal function tests and pentachlorophenol exposure. Clinical Toxicology. 1977;11:97-106.

46. Farquharson ME, Gage JC, Northover J. The biological action of chlorophenols. British Journal of Pharmacology and Chemotherapy. 1958;13:20-24.

47. Byard JL. Mechanisms of acute human poisoning by pesticides. Clinical Toxicology. 1979;14:187-193.

48. Robson AM, Elvick NH, Pundavela L, Shirkey HC. Pentachlorophenol poisoning in a nursery for newborn infants. I. Clinical features and treatment. Journal of Pediatrics. 1969;75:309-316.

49. Dubois M, Plaisance H, Thome JP, Kremers P. Hierarchical cluster analysis of environmental pollutants through P450 induction in cultured hepatic cells. Ecotoxicology & Environmental Safety. 1996;34:205-215.

50. Jorens PG, Schepens PJC. Human pentachlorophenol poisoning. Human & Experimental Toxicology. 1993;12:479-495.

51. Hayes WJ. Pesticides studied in man. Baltimore, MD: Waverly Press; 1982. 52. Gorman D, Monigatti J, Glass B, Gronwall D, Beasley M. Assessment of

pentachlorophenol-exposed timber workers using a test-of-poisoning model.

51

International Journal of Occupational & Environmental Health. 2001;7:189-194.

53. Daniel V, Huber W, Bauer K, et al. Association of elevated blood levels of pentachlorophenol (PCP) with cellular and humoral immunodeficiencies. Archives of Environmental Health. 2001;56:77-83.

54. Peper M, Ertl M, Gerhard I. Long-term exposure to wood-preserving chemicals containing pentachlorophenol and lindane is related to neurobehavioral performance in women. American Journal of Industrial Medicine. 1999;35:632-641.

55. Sangster B, Wegman RC, Hofstee AW. Non-occupational exposure to pentachlorophenol: clinical findings and plasma-PCP-concentrations in three families. Human Toxicology. 1982;1:123-133.

56. Klemmer HW, Wong L, Sato MM, Reichert EL, Korsak RJ, Rashad MN. Clinical findings in workers exposed to pentachlorophenol. Archives of Environmental Contamination & Toxicology. 1980;9:715-725.

57. Hryhorczuk DO, Wallace WH, Persky V, et al. A morbidity study of former pentachlorophenol-production workers. Environmental Health Perspectives. 1998;106:401-408.

58. Crosby DG. Environmental chemistry of pentachlorophenol. Journal of Applied Chemsitry. 1981;53:1051-1080.

59. Walls CB, Glass WI, Pearce NE. Health effects of occupational pentachlorophenol exposure in timber sawmill employees: a preliminary study. New Zealand Medical Journal. 1998;111:362-364.

60. Gerhard I, Daniel V, Link S, Monga B, Runnebaum B. Chlorinated hydrocarbons in women with repeated miscarriages. Environmental Health Perspectives. 1998;106:675-681.

61. De Maeyer J, Schepens PJC, Jorens PG, Verstraete R. Exposure to pentachlorophenol as a possible cause of miscarriages. British Journal of Obstetrics and Gynaecology. 1995;102:1010-1011.

62. Gerhard I, Frick A, Monga B, Runnebaum B. Pentachlorophenol exposure in women with gynecological and endocrine dysfunction. Environmental Research. 1999;80:383-388.

63. Cheng WN, Coenraads PJ, Hao ZH, Liu GF. A health survey of workers in the pentachlorophenol section of a chemical manufacturing plant. American Journal of Industrial Medicine. 1993;24:81-92.

64. Gunderson EL. FDA total diet study, July 1986 to April 1991, dietary intakes of pesticides, selected elements, and other chemicals. Journal of AOAC International. 1995;78:1353-1363.

65. Johnson RL, Gehring PJ, Kociba RJ, Schwertz BA. Chlorinated dibenzodioxins and pentachlorophenol. Environmental Health Perspectives. 1973;5:171-175.

66. Greichus YA, Libal GW, Johnson DD. Diagnosis and physiologic effects of pentachlorophenols on young pigs. Part I. Effects of purified pentachlorophenol. Bulletin of Environmental Contamination & Toxicology. 1979;23:418-422.

67. Knudsen I, Verschuuren HG, den Tonkelaar EM, Kroes R, Helleman PF. Short-term toxicity of pentachlorophenol in rats. Toxicology. 1974;2:141-152.

68. Nishimura H, Nishimura N, Oshima H. Experimental studies on the toxicity of pentachlorophenol. Journal of the Aichi Medical University Association. 1980;8:203-209.

69. Umemura T, Sai-Kato K, Takagi A, Hasegawa R, Kurokawa Y. Oxidative DNA damage and cell proliferation in the livers of B6C3F1 mice exposed to

52

pentachlorophenol in their diet. Fundamental & Applied Toxicology. 1996;30:285-289.

70. Blakley BR, Yole MJ, Brousseau P, Boermans H, Fournier M. Effect of pentachlorophenol on immune function. Toxicology. 1998;125:141-148.

71. Kimbrough RD, Linder RE. The effect of technical and purified pentachlorophenol on the rat liver. Toxicology & Applied Pharmacology. 1978;46:151-162.

72. Bernard BK, Hoberman AM, Brown WR, Ranpuria AK, Christian MS. Oral (gavage) two-generation (one litter per generation) reproduction study of pentachlorophenol (penta) in rats. International Journal of Toxicology. 2002;21:301-318.

73. Jekat FW, Meisel ML, Eckard R, Winterhoff H. Effects of pentachlorophenol (PCP) on the pituitary and thyroidal hormone regulation in the rat. Toxicology Letters. 1994;71:9-25.

74. Rawlings NC, Cook SJ, Waldbillig D. Effects of the pesticides carbofuran, chlorpyrifos, dimethoate, lindane, triallate, trifluralin, 2,4-D, and pentachlorophenol on the metabolic endocrine and reproductive endocrine system in ewes. Journal of Toxicology & Environmental Health Part A. 1998;54:21-36.

75. Beard AP, Bartlewski PM, Rawlings NC. Endocrine and reproductive function in ewes exposed to the organochlorine pesticides lindane or pentachlorophenol. Journal of Toxicology & Environmental Health Part A. 1999;56:23-46.

76. Beard AP, Rawlings NC. Reproductive effects in mink (Mustela vison) exposed to the pesticides Lindane, Carbofuran and Pentachlorophenol in a multigeneration study. Journal of Reproduction & Fertility. 1998;113:95-104.

77. Chhabra RS, Maronpot RM, Bucher JR, Haseman JK, Toft JD, Hejtmancik MR. Toxicology and carcinogenesis studies of pentachlorophenol in rats. Toxicological Sciences. 1999;48:14-20.

78. Kerkvliet NI, Baecher-Steppan L, Schmitz JA. Immunotoxicity of pentachlorophenol (PCP): increased susceptibility to tumor growth in adult mice fed technical PCP-contaminated diets. Toxicology & Applied Pharmacology. 1982;62:55-64.

79. Kerkvliet NI, Brauner JA, Matlock JP. Humoral immunotoxicity of polychlorinated diphenyl ethers, phenoxyphenols, dioxins and furans present as contaminants of technical grade pentachlorophenol. Toxicology. 1985;36:307-324.

80. White KL, Jr., Anderson AC. Suppression of mouse complement activity by contaminants of technical grade pentachlorophenol. Agents & Actions. 1985;16:385-392.

81. Savolainen H, Pekari K. Neurochemical effects of peroral administration of technical pentachlorophenol. Research Communications in Chemical Pathology & Pharmacology. 1979;23:97-105.

82. Villena F, Montoya G, Klaasen R, Fleckenstein R, Suwalsky M. Morphological changes on nerves and histopathological effects on liver and kidney of rats by pentachlorophenol (PCP). Comparative Biochemistry & Physiology. C, Comparative Pharmacology & Toxicology. 1992;101:3533-3563.

83. IARC. Polychlorophenols and their sodium salts. IARC monographs for the evaluation of the carcinogenic risk of chemicals to humans. 71 Pt 2. Lyon, France: World Health Organization, International Agency for Research on Cancer; 1999:769-816.

53

84. EPA. (U.S. Environmental Protection Agency) Integrated Risk Information System (IRIS) on Pentachlorophenol. Cincinnati, OH: Environmental Criteria and Assessment Office, Office of Health and Environmental Assessment, Office of Research and Development; 1993.

85. Lampi P, Hakulinen T, Luostarinen T, Pukkala E, Teppo L. Cancer incidence following chlorophenol exposure in a community in southern Finland. Archives of Environmental Health. 1992;47:167-175.

86. Demers PA, Davies HW, Friesen MC, et al. Cancer and occupational exposure to pentachlorophenol and tetrachlorophenol (Canada). Cancer Causes & Control. 2006;17:749-758.

87. Teschke K, Hertzman C, Fenske RA, et al. A history of process and chemical changes for fungicide application in the western Canadian lumber industry: What can we learn? Applied Occupational & Environmental Hygiene. 1994;9:984-993.

88. Hardell L, Eriksson M, Degerman A. Meta analysis of 4 Swedish case-control studies on exposure to pesticides as risk-factor for soft-tissue sarcoma including the relation to tumor-localization and histopathological type. International Journal of Oncology. 1995;6:847-851.

89. Pearce NE, Smith AH, Howard JK, Sheppard RA, Giles HJ, Teague CA. Non-Hodgkins-lymphoma and exposure to phenoxyherbicides, chlorophenols, fencing work, and meat works employment - a case-control study. British Journal of Industrial Medicine. 1986;43:75-83.

90. Kogevinas M, Kauppinen T, Winkelmann R, et al. Soft-tissue sarcoma and non-Hodgkins-lymphoma in workers exposed to phenoxy herbicides, chlorophenols, and dioxins - 2 nested case-control studies. Epidemiology. 1995;6:396-402.

91. Smith AH, Pearce NE, Fisher DO, Giles HJ, Teague CA, Howard JK. Soft-tissue sarcoma and exposure to phenoxyherbicides and chlorophenols in New-Zealand. Journal of the National Cancer Institute. 1984;73:1111-1117.

92. Smith JG, Christophers AJ. Phenoxy herbicides and chlorophenols - a case control study on soft-tissue sarcoma and malignant-lymphoma. British Journal of Cancer. 1992;65:442-448.

93. Schwetz BA, Quast IF, Keeler PA, Humiston CG, Kociba RJ. Results of two-year toxicity and reproduction studies on pentachlorophenol in rats. In: Rao KR, ed. Pentachlorophenol: chemistry, pharmacology, and environmental toxicology. New York: Plenum Press; 1978:301-309.

94. NTP. NTP technical report on the toxicology and carcinogenesis studies of two pentachlorophenol technical grade mixtures (CAS no 87-86-5) in B6C3F1 mice (feed studies). NTP TR 349. NIH publication No. 89-2804. Research Triangle Park, NC: National Institutes of Health, Public Health Service, U.S. Department of Health and Human Services, National Toxicology Program; 1989. In ATSDR, Toxicological profile for pentochlorophenol. 2001, Department of Health and Human Services, Public Health Service. Agency for Toxic Substances and Disease Registry: Atlanta, Ga.

95. NTP. Toxicology and carcinogenesis studies of pentachlorophenol (CAS no 87-86-5) in F344/N rats (feed studies). NIH publication no. 99-3973. NTP TR 483. Research Triangle Park, NC: National Institutes of Health, Public Health Service, U.S. Department of Health and Human Services, National Toxicology Program; 1999. In ATSDR, Toxicological profile for pentochlorophenol. 2001, Department of Health and Human Services, Public Health Service. Agency for Toxic Substances and Disease Registry: Atlanta, Ga.

54

96. NCI. Evaluation of carcinogenic, teratogenic, and mutagenic activities of selected pesticides and industrial chemicals. Volume I: Carcinogenic study. NTIS PB-223-159. Bethesda, MD: National Cancer Institute; 1968. In ATSDR, Toxicological profile for pentochlorophenol. 2001, Department of Health and Human Services, Public Health Service. Agency for Toxic Substances and Disease Registry: Atlanta, Ga.

97. Bauchinger M, Dresp J, Schmid E, Hauf R. Chromosome changes in lymphocytes after occupational exposure to pentachlorophenol (PCP). Mutation Research. 1982;102:83-88.

98. Wyllie JA, Gabica J, Benson WW, Yoder J. Exposure and contamination of the air and employees of a pentachlorophenol plant, Idaho--1972. Pesticides Monitoring Journal. 1975;9:150-153.

99. Ziemsen B, Angerer J, Lehnert G. Sister chromatid exchange and chromosomal breakage in pentachlorophenol (PCP) exposed workers. International Archives of Occupational & Environmental Health. 1987;59:413-417.

100. Sai-Kato K, Umemura T, Takagi A, Hasegawa R, Tanimura A, Kurokawa Y. Pentachlorophenol-induced oxidative DNA damage in mouse liver and protective effect of antioxidants. Food & Chemical Toxicology. 1995;33:877-882.

101. Waters MD, Sandhu SS, Simmon VF, et al. Study of pesticide genotoxicity. Basic Life Sciences. 1982;21:275-326.

102. Andersen KJ, Leighty EG, Takahash MT. Evaluation of Herbicides For Possible Mutagenic Properties. Journal of Agricultural and Food Chemistry. 1972;20:649-656.

103. Donnelly KC, Claxton LD, Huebner HJ, Capizzi JL. Mutagenic interactions of model chemical mixtures. Chemosphere. 1998;37:1253-1261.

104. Markiewicz KV, Howie LE, Safe SH, Donnelly KC. Mutagenic potential of binary and complex mixtures using different enzyme induction systems. Journal of Toxicology & Environmental Health. 1996;47:443-451.

105. Fahrig R. Comparative mutagenicity studies with pesticides. IARC Scientific Publication. 1974;10:161-181.

106. Dahlhaus M, Almstadt E, Henschke P, Lüttgert S, Appel KE. Oxidative DNA lesions in V79 cells mediated by pentachlorophenol metabolites. Archives of Toxicology. 1996;70:457-460.

107. Ehrlich W. The effect of pentachlorophenol and its metabolite tetrachlorohydroquinone on cell growth and the induction of DNA damage in Chinese hamster ovary cells. Mutation Research. 1990;244:299-302.

108. Wang YJ, Lin JK. Estimation of selected phenols in drinking water with in situ acetylation and study on the DNA damaging properties of polychlorinated phenols. Archives of Environmental Contamination & Toxicology. 1995;28:537-542.

109. Karmaus W, Wolf N. Reduced birthweight and length in the offspring of females exposed to PCDFs, PCP, and lindane. Environmental Health Perspectives. 1995;103:1120-1125.

110. Heacock H, Hertzman C, Demers PA, et al. Childhood cancer in the offspring of male sawmill workers occupationally exposed to chlorophenate fungicides. Environmental Health Perspectives. 2000;108:499-503.

111. Bernard BK, Hoberman AM. A study of the developmental toxicity potential of pentachlorophenol in the rat. International Journal of Toxicology. 2001;20:353-362.

55

112. Schwetz BA, Keeler PA, Gehring PJ. The effect of purified and commercial grade pentachlorophenol on rat embryonal and fetal development. Toxicology & Applied Pharmacology. 1974;28:151-161.

113. Cautreels W, Van Couwenberghe K, Gusman LA. Comparison between the organic fraction of suspended matter at a background and an urban station. The Science of The Total Environment. 1977;8:79-88.

114. Bundesamt fur Umweltschutz. [Polychlorinated organic compounds in Rheinfelden (Switzerland). Investigation of dust, soil, and grass samples: results and evaluation.]. Bern: Schriftenreihe Umweltschutz; 1983:32.

115. Waite DT, Gurprasad NP, Cessna AJ, Quiring DV. Atmospheric pentachlorophenol concentrations in relation to air temperature at five Canadian locations. Chemosphere. 1998;37:2251-2260.

116. Arsenault RD. Pentachlorophenol and contained chlorinated dibenzodioxins in the environment: a study of environmental fate, stability, and significance when used in wood preservation. Annual Meeting of the American Wood-Preservers' Association. Atlanta (Ga.); 1976.

117. Gebefügi I, Parlar H, Korte F. Occurrence of pentachlorophenol in enclosed environments. Ecotoxicology & Environmental Safety. 1979;3:269-300.

118. Fountaine JE, Joshipura PB, Keliher PN. Some observations regarding pentachlorophenol levels in Haverford Township, Pennsylvania. Water Research. 1976;10:185-188.

119. IPCS. Pentochlorophenol. International Programme on Chemical Safety. Health and Safety Guide No. 19. Geneva: World Health Organisation; 1989.

120. Gifford JS, Buckland SJ, Judd MC, McFarlane PN, Anderson SM. Pentachlorophenol (PCP), PCDD, PCDF and pesticide concentrations in a freshwater lake catchment. Chemosphere. 1996;32:2097-2113.

121. Pentachlorophenol: Wood Preservative. CAPCO Note 87:02. Agriculture Canada Food Production and Inspection Branch. Ottowa: Agriculture Canada; 1987.

122. Bunce NJ, Nakai JS. Atmospheric chemistry of chlorinated phenols. Journal of Air Polution and Control Association. 1989;39:820-823.

123. Kilzer L, Scheunert I, Geyer H, Klein W, Korte F. Laboratory screening of the volatilization rates of organic chemicals from water and soil. Chemosphere. 1979;8:751-761.

124. Pignatello JJ, Martinson MM, Steiert JG, Carlson RE, Crawford RL. Biodegradation and photolysis of pentachlorophenol in artificial freshwater streams. Applied & Environmental Microbiology. 1983;46:1024-1031.

125. Schellenberg K, Leuenberger C, Schwarzenbach RP. Sorption of chlorinated phenols by natural sediments and aquifer materials. Environmental Science & Technology. 1984;18:652-657.

126. Chang NI, Choi J. Studies on the adsorption of pentachlorophenol ( PCP ) in soils. Hanguk Touang Bilyo Hakkhoe Chi. 1974;7:197-220.

127. Kuwatsuka S, Igarashi M. Degradation of PCP in soil: II. The relationship between the degradation of PCP and the properties of soils and the identification of the degradation products of PCP. Soil Science and Nutrition. 1975;21:405-414.

128. ATSDR. Toxicological profile for pentochlorophenol. Atlanta, Ga: Department of Health and Human Services, Public Health Service. Agency for Toxic Substances and Disease Registry; 2001.

56

129. Veith GD, Defoe D, Knuth M. Structure Activity Relationships For Screening Organic-Chemicals For Potential Ecotoxicity Effects. Drug Metabolism Reviews. 1985;15:1295-1303.

130. Makela TP, Petanen T, Kukkonen J, Oikari AO. Accumulation and depuration of chlorinated phenolics in the freshwater mussel (Anodonta anatina L.). Ecotoxicology & Environmental Safety. 1991;22:153-163.

131. Pruitt GW, Grantham BJ, Pierce RH. Accumulation and elimination of pentachlorophenol by the bluegill, Lepomis macrochirus. Transactions of the American Fisheries Society. 1977;106:462-465.

132. Basack SB, Oneto ML, Verrengia Guerrero NR, Kesten EM. Accumulation and elimination of pentachlorophenol in the freshwater bivalve Corbicula fluminea. Bulletin of Environmental Contamination & Toxicology. 1997;58:497-503.

133. Niimi AJ, Cho CY. Laboratory and field analysis of pentachlorophenol (PCP) accumulation by salmonids. Water Research. 1983;17:1791-1795.

134. Heudorf U, Angerer J, Drexler H. Current internal exposure to pesticides in children and adolescents in Germany: blood plasma levels of pentachlorophenol (PCP), lindane (gamma-HCH), and dichloro(diphenyl)ethylene (DDE), a biostable metabolite of dichloro(diphenyl)trichloroethane (DDT). International Journal of Hygiene & Environmental Health. 2003;206:485-491.

135. Carrizo D, Grimalt JO, Ribas-Fito N, Torrent M, Sunyer J. Pentachlorobenzene, hexachlorobenzene, and pentachlorophenol in children's serum from industrial and rural populations after restricted use. Ecotoxicology and Environmental Safety. 2008;71:260-266.

136. Wong AS, Crosby DG. Photodecomposition of pentachlorophenol in water. Journal of Agricultural and Food Chemistry. 1981;29:125-130.

137. Meylan WM, Howard PH. Computer estimation of the atmospheric gas-phase reaction-rate of organic-compounds with Hydroxyl radicals and ozone. Chemosphere. 1993;26:2293-2299.

138. Frisbie AJ, Nies L. Aerobic and anaerobic biodegradation of aged pentachlorophenol by indigenous microorganisms. Bioremediation Journal. 1997;1:65-75.

139. McAllister KA, Lee H, Trevors JT. Microbial degradation of pentachlorophenol. Biodegradation. 1996;7:1-40.

140. Dolfing J, Harrison BK. Redox and reduction potentials as parameters to predict the degradation pathway of chlorinated Benzenes in anaerobic environments. Fems Microbiology Ecology. 1993;13:23-29.

141. Nomura S. Studies on chlorophenol poisoning. Rep. 1. A clinical examination of workers exposed to pentachlorophenol. The Journal of science of labour. 1953;29:474-483.

142. Dreisbach RH. Prevention, diagnosis, and treatment. Handbook of poisonings. 10th ed. Los Altos, CA: Lange Medical Publications; 1980.

143. Cline RE, Hill RH, Jr., Phillips DL, Needham LL. Pentachlorophenol measurements in body fluids of people in log homes and workplaces. Archives of Environmental Contamination & Toxicology. 1989;18:475-481.

144. CDC. Pentachlorophenol in log homes - Kentucy. Morbidity and Mortality Weekly Report. 1980;29:431-437.

145. CDC. Current trends follow-up on pentachlorophenol in log homes. Morbidity and Mortality Weekly Report. 1982;31:170-171.

146. DoL. Health Outcomes in Former New Zealand Timber Workers Exposed To Pentachlorophenal. Available at

57

<http://www.dol.govt.nz/PDFs/Final_PCP_Report_Jan_2008.pdf > (Last accessed on March 08 2009). Wellington: Department of Labour; 2008.

147. Ramlow JM, Spadacene NW, Hoag SR, Stafford BA, Cartmill JB, Lerner PJ. Mortality in a cohort of pentachlorophenol manufacturing workers, 1940-1989. American Journal of Industrial Medicine. 1996;30:180-194.

148. Gilbert FI, Jr., Minn CE, Duncan RC, Wilkinson J. Effects of pentachlorophenol and other chemical preservatives on the health of wood-treating workers in Hawaii. Archives of Environmental Contamination & Toxicology. 1990;19:603-609.

149. Armstrong RW, Eichner ER, Klein DE, et al. Pentachlorophenol Poisoning in a Nursery For Newborn Infants .2. Epidemiologic and Toxicologic Studies. Journal of Pediatrics. 1969;75:317-325.

150. WHO. Pentachlorophenol in drinking-water. Background document for development of WHO guidelines for drinking-water quality. WHO/SDE/WSH/03.04/62. Available at <http://www.who.int/water_sanitation_health/dwq/chemicals/pentachlorophenol.pdf> (Last accessed on Oct 17 2008). Geneva, Switzerland: World Health Organization; 2003.

151. Triebig G, Csuzda I, Krekeler HJ, Schaller KH. Pentachlorophenol and the peripheral nervous-system - a longitudinal-study in exposed workers. British Journal of Industrial Medicine. 1987;44:638-641.

152. Needham LL, Cline RE, Head SL, Liddle JA. Determining pentachlorophenol in body fluids by gas chromatography after acetylation. Journal of Analytical Toxicology. 1981;5:283-286.

153. Diserens JM. Rapid determination of nineteen chlorophenols in wood, paper, cardboard, fruits, and fruit juices by gas chromatography/mass spectrometry. Journal of Aoac International. 2001;84:853-860.

154. Bernal JL, Nozal MJ, Toribio L, et al. Determination of phenolic compounds in water samples by online solid-phase extraction–supercriticalfluid chromatography with diode-array detection. Chromatographia. 1997;46:295-300.

155. Zhou Q, Liu J, Jiang G, Liu G, Cai Y. Sensitivity enhancement of chlorinated phenols by continuous flow liquid membrane extraction followed by capillary electrophoresis. Journal of Separation Science. 2004;27:576-580.

156. Gabelish CL, Crisp P, Schneider RP. Simultaneous determination of chlorophenols, chlorobenzenes and chlorobenzoates in microbial solutions using pentafluorobenzylbromide derivatization and analysis by gas chromatography with electron-capture detection. Journal of Chromatography A. 1996;749:165-171.

157. Campillo N, Penalver R, Hernandez-Cordoba M. A sensitive solid-phase microextraction/gas chromatography-based procedure for determining pentachlorophenol in food. Food Additives & Contaminants. 2007;24:777-783.

58

3.0 Polychlorinated Biphenyls

3.1 Physical Chemical Properties As there are 209 different PCBs, each with its own distinct properties, a summary of each is beyond the scope of this report, however, brief summaries of the 8 major Monsanto products (described as Aroclor) are presented in Table 3.1. See section 3.1.3 for further details of theses formulations.

3.1.1 Basic Properties Table 3.1 Physical Chemical Properties of 8 Major Aroclor products Property Aroclor 1016 Aroclor 1221 Aroclor 1232 Aroclor 1242 Molecular weight 257.9 200.7 232.2c 266.5 Melting point, oC No data 1 No data No data Boiling point, oC 325–356 275–320 290–325 325–366 Density, g/cm3 at 25 oC 1.37 1.18 1.26 1.38 Solubility : Water, mg/L 0.42 (25 oC) 0.59 (24 oC) 0.45 (25 oC) 0.24 (25 oC) Organic solvent(s) Very soluble Very soluble Very soluble Very soluble Partition coefficients Log Kow 5.6 4.7 5.1 5.6 Log Koc No data No data No data No data Vapor pressure, mm Hg at 25 oC

4 x 10-4 6.7 x 10-3 4.06 x 10-3 4.06 x 10-4

Henry’s law constant, atm-m3/mol at 25 oC

2.9 x 10-4 3.5 x 10-3 No data 5.2 x 10-4

Property Aroclor 1254 Aroclor 1260 Aroclor 1262 Aroclor 1268 Molecular weight 328 357.7 389 453 Melting point, oC No data No data No data No data Boiling point, oC 365–390 385–420 390–425 435–450 Density, g/cm3 at 25 oC 1.54 1.62 1.64 1.81 Solubility : Water, mg/L at 24 oC 0.012 0.027 0.052 0.3 Organic solvent(s) Very soluble Very soluble No data Soluble Partition coefficients Log Kow 6.5 6.8 No data No data Log Koc No data No data No data No data Vapor pressure, mm Hg at 25 oC

7.7 x 10-5 4.05 x 10-3 No data No data

Henry’s law constant, atm-m3/mol at 25 oC

2.0 x 10-3 4.6 x 10-3 No data No data

3.1.2 Brief Notes The term polychlorinated biphenyls (PCB) describe a class of compounds in which a biphenyl molecule can have from 1 to 10 chlorine atoms attached to the two aromatic rings (Figure 3.1). A total of 209 possible compounds are known; each is called a congener.1 However, commercial preparations only contain 130 congeners at concentrations exceeding 0.05%.2 PCBs with the same degree of chlorine saturation

59

are defined as homologs; for example, PCBs with three chlorine molecules are defined as homologs and are termed trichlorobiphenyls. Different substitutions of chlorine within a homolog are known as isomers. Figure 3.1 The structural formula of PCBs; positions for attachment of either the hydrogen or chlorine molecule can occur at 2’,3’,4’,5’,6’ and 2,3,4,5,6. Substitution with chlorine on the 2 (also 2’,6,6’) position is termed ortho, at position 3 (also 3’, 5, 5’) meta and position 4 (also 4’) is defined as para.

PCBs can also be grouped into coplanar and non-coplanar congeners, depending on how the two benzene rings are orientated with respect to each other. Chlorines at the ortho position (see Figure 3.1 - positions at 2, 2’, 6, 6’) will introduce some constraint to the rotational freedom of the aromatic ring and are termed non-coplanar (i.e. they don’t lie on the same plane). In contrast, those therefore without substitution at the ortho position are known as co-planar PCBs (ie they lie in the same plane). These co-planar PCBs (without chlorine substitution on the ortho position) are known to interact with and activate the aryl hydrocarbon receptor (AhR), which is important in stimulating gene transcription in response to xenobiotics. Of the 209 congeners 12 possess such properties and are therefore termed dioxin-like PCBs. They have been assigned toxic equivalent factors, relative to the most toxic dioxin, 2,3,7,8-tetra CDD.3 A full explanation is provided in section 3.1.1 of the dioxin document. PCBs are relatively insoluble in water, and such solubility decreases further with increasing chlorination. Conversely, they are very soluble in organic (non-polar) solvents.4 PCBs are combustible at high temperatures and the products of combustion are more toxic than the parent molecules. By-products of combustion include hydrogen chloride, and the two major classes of dioxin, polychlorinated dibenzodioxins (PCDDs), and polychlorinated dibenzofurans (PCDFs).5, 6 These compounds are also formed during the manufacture of PCBs, and the products themselves contain dioxin impurities.7

3.1.3 Uses A distinguishing feature of PCBs is their inertness; they are resistant to both acid and alkali and have thermal stability. Consequently they were widely employed in various applications such as dielectric fluids in transformers and capacitors, heat transfer fluids, lubricants, paints, pesticide extenders and plasticisers.8-10 From 1930 to 1977, PCBs were marketed in the United States by Monsanto Corporation as mixtures of PCBs under trade names such as Aroclor 1016, 1242, 1254 and 1260; the second two digits of the name represents the chlorine content by weight (for example 1242 means a PCB mixture of mono- through heptachlorinated biphenyl congeners containing approximately 42% chlorine by weight). Other nations marketed PCBs in formulations very similar to the Monsanto products. By 1972, the United States alone

60

used 340 million kilograms annually.11 Following the Toxic Substance Control Act, use in that country was banned in 1977. New Zealand has also banned the production and use of PCBs. Controls were placed for PCBs under the repealed Toxic Substances Act 1979 and associated regulations such as the prohibition on importation, storage and use of PCB. The Hazardous Substances and New Organisms Act (1996) was also amended, in accordance with the Stockholm convention on persistent organic pollutants, thereby totally banning its use in New Zealand.216

3.2 Routes of Exposure A constant and recurring theme in this report on PCBs is differing properties of the individual congeners and the difficulty in summarising the information available, which more often focuses on commercial formulations. As such, the report will, where possible, focus on key studies identified in the various categories of the toxicological profile, and attempt to summarise that data for the wide range of PCBs most likely to be involved in human exposure and subsequent risk of toxicity. The majority of human studies have been attained from two major outbreaks of disease resulting from contaminated rice oil. The first, now called ‘Yusho’, occurred in Western Japan where it was estimated approximately 1700 people were poisoned.12,

13 The second outbreak affected ~2000 people in central Taiwan; this illness is now called ‘Yu-Cheng’.14, 15 These two incidents will be summarised in section 3.10.2 (case reports). Human studies are also derived from incidents of workplace and environmental exposure.

3.2.1 Skin Skin absorption is a major contributor to the overall intake of human exposure to PCB. Although less efficient than gut absorption, experimental studies suggest as much as 40 to 50% of an applied dose is ultimately absorbed into systemic circulation.16, 17 With humans, exposure to PCBs occurs primarily by direct dermal contact in a workplace environment.18, 19 This is confirmed by numerous studies correlating blood PCB levels of workers in related industries who were exposed to these contaminants in the course of their work.20, 21

3.2.2 Ingestion Another important route of exposure is by ingestion, typically through the ingestion of contaminated meat and dairy products, as well as processed foods and fish.22-24 Absorption is efficient and rapid. Studies with animal models suggest as much as 90% or more of a single or multiple dose of various PCB congeners may be actively absorbed.25, 26

3.2.3 Inhalation Evidence from workplace exposure suggests inhalation may also be an important route of exposure. Based on data summarised by Wolfe and colleagues (1985),27 indirect evidence has suggested that high levels commonly seen in adipose tissue from exposed capacitor workers may have been due to inhalation rather than dermal or oral exposure.

61

3.3 Toxicokinetics

3.3.1 Absorption PCB is readily absorbed into the body via skin contact, inhalation and ingestion. Absorption is influenced by the lipid solubility of the PCB congener. Increasing the number of chlorine molecules to the biphenyl structure will increase its lipid solubility.28 Passive movement across the enterocytes into host blood circulation will increase with increasing lipophilicity.29 As such, intestinal absorption is high, though net absorption can vary considerably, depending on the concentration of the measured PCB already present in blood lipids and the congener being ingested.30 Almost complete absorption for some congeners has been reported though lower values were obtained for others and some volunteers experienced negative values for differing congeners (more excreted than absorbed).30 Dietary intake will also influence PCB uptake; reduced diet intake will lead to an increased percentage of excretion.31 PCBs will readily transfer from the exposed mother to the baby, both in utero and post partum.32 To the knowledge of the author, little is known on rates of absorption consequent to inhalation. Skin absorption is less efficient than the intestinal route (see section 3.2.1). With animal models, rates of delivery are slower and are dependent on the age of the animal; following acute exposure in one study, for example, 35% and 26% bioavailability occurred after 72 hours with young and adult rats respectively.17 Percentages were elevated to 45% and 43% after 120 hours.

3.3.2 Distribution PCBs have very poor water solubility (see section 3.1.1) and once absorbed by the mammalian host depend on blood carrier molecules for transport. Upon absorption the initial target for PCBs are the highly perfused organs (predominantly the brain, kidney, lung and liver) with later redistribution to organs of high lipid content.33, 34 Irrespective of the dose ingested or the studied animal species, the predominant distribution of PCBs occurs in the adipose tissue.35 The skin is the dominant reservoir whereas other tissues having similar composition of lipids, such as the brain, have significantly lower PCB concentrations.36-38 Assessing body tissues from a Yu-Cheng patient demonstrated accumulations of PCB almost exclusively in adipose tissue; small concentrations were also evident in the liver (the adipose:liver ratio was ~20:1).37 PCB concentration in adipose tissue has been reported to be proportional to that in plasma, with a partition for total PCBs of approximately 190:1.18 Similar hepatic results were found with a Yu-Cheng baby.39 There is some suggestion from this study there may have been some sequestration in the liver, and the contaminants in the toxic dose may have been responsible for the patient’s death. However, the majority of studies have shown that liver sequestration of PCBs and other like chlorinated contaminants does not occur.40-42

62

3.3.3 Metabolism The metabolic fate of PCBs is influenced by the respective congener. PCBs with lower chlorine saturation, those that have two adjacent carbons without chlorine atoms and those where the para position is chlorine free are all readily metabolised by cytochrome P450 isoenzymes to form polar metabolites.43, 44 Those PCBs that are co-planar or partially substituted with chlorine at the ortho position are not so readily metabolised. In humans, phase I metabolism of PCB results in the formation of an hydroxylated PCB followed by phase II conjugation with either glucoronidation43 or sulfonation occurring.45 Retained hydroxylated PCB in the plasma can, in animal models, interact with thyroid hormone transport protein.46 The conjugated glucoronide metabolite is excreted into the bile and released into the gastrointestinal tract. Within the large intestine, bacteria cleave the sulfur-carbon bond with subsequent methylation and the resultant molecule is reabsorbed back into the body.47 It undergoes a final oxidation to yield a methylsulfone metabolite48 that accumulates in various organs of the body, predominantly in the adipose tissue. Forty such methyl-derivatives of PCB have, for example, been detected in the body of a patient with Yusho syndrome.49

3.3.4 Elimination Rates of elimination will depend primarily on the individual congener,29 and be influenced by the degree of chlorine saturation50 and the position on the biphenyl molecule.51 Variation may be further influenced by duration of exposure and variability in the host’s ability to metabolise the respective congeners.52. Therefore, there is a very wide variety of reported half-lives in the literature following single and short term exposure associated with both workplace exposure and in the general population. Thus, it is virtually impossible to provide a definitive PCB half-life. A brief summary of the literature is provided below. One investigation monitored isotope-labelled PCB (in a mixture similar to Aroclor 1254) in a human volunteer. Three isomers were identified in the body and subsequently displayed elimination half-lives ranging from 124 to 338 days, depending on the individual isomer.53 Similar results were reported in firemen exposed to Aroclor 1254 contamination following a fire in a transformer.54 Workers exposed for less than 5 years had PCB elimination half-lives of several months whereas values for those exposed for greater than 5 years were 2 – 3 years.56 Other studies have reported half-lives ranging from 1.4 to 12 years following long-term workplace exposure.52 In female Rhesus monkeys, half-life estimations for nine PCB congeners ranged from 0.3–7.6 years.55 Twenty years after Yu-Cheng, intoxicated patients still had elevated levels of PCBs, 3.7 times greater than the general population.57

3.4 Mechanism of Action within the Body PCB isomers and congeners that are recognised as acutely toxic are those in which the chlorine molecule is present in the meta and para positions, but not the ortho position (see section 3.1.2). They have a higher binding affinity for the Ah-receptors in the cell

63

cytosol. This receptor is responsible for mediating the induction of a range of cytochrome P450 isozymes.58 Interaction with these receptors may consequently lead to improper gene expression; chronic studies in rats, for example, have shown the expression of liver tumors.99 The mechanism for the development of chloroacne and other dermatological complaints may be associated with the failure of vitamin A utilisation.59

3.5 Summary of Clinical Signs and Symptoms Following Toxic Exposure Symptoms associated with PCB toxicity in animals have included reduced rates of growth and weight loss, changes in organ weight, acnegenesis and porphyria. Observations following human toxicity (Yu-Cheng and Yusho) have shown similar conditions, though there were other contaminants in the oil including dioxins and naphthalenes (see section 3.10.2), that may have influenced the presented signs and symptoms.

However, the evaluation of PCB toxicity in humans is complex since they usually occur as a mixture of various congeners. Some components remain stable in the environment whereas other in the mixture readily degrades (see section 3.9.3).


3.5.1.1 Human (acute) Signs and symptoms following acute exposure to PCBs have been predominantly reported in Yu-Cheng and Yusho patients. Most symptoms are related to the skin and eyes. A full summary of the reported signs and symptoms following the incident in Japan are summarised in Table 3.2. Further details are provided in sections 3.5.1.1.1, 3.5.1.1.2 and 3.10.2. Table 3.2 Initial signs and symptoms reported in Yusho patients following PCB exposure.60

Dermal Neurological Ophthalmic Miscellaneous Dark brown pigmentation of nails

Feeling of weakness Increased eye discharge

Fever

Distinctive hair follicles Hearing difficulties Hyperemia of conjunctiva

Vomiting

Increased sweating at palms

Numbness in limbs Transient visual disturbance

Diarrhoea

Acne-like skin eruptions Headache Swelling of upper eyelids

Red plaques on limbs Spasm of limbs Itching Pigmentation of skin Swelling of limbs Stiffened soles in feet and palms of hands

Pigmented mucous membrane

Jaundice

64

3.5.1.1.1 Dermal

The principal symptom following human acute exposure is dermal injury, the most severe and easily recognised form being chloroacne.61 Symptoms of chloroacne occur following exposure to high doses of PCBs and other chlorinated organic compounds. Initial symptoms of Yusho and Yu-Cheng patients following the ingestion of tainted rice oil were generalised chloroacne and hyperpigmentation of the skin. Hyperpigmentation also extended to conjunctival regions of the eye (see section 3.5.1.1.2), and the nail beds. Children exposed to PCBs in utero also suffered hyperpigmented skin and were defined as ‘cola-coloured babies’. This pigmentation faded within three months post-birth. See table 3.2 for further details of dermal symptoms associated with the Yusho patients..

3.5.1.1.2 Eye

The typical characteristic of PCB poisoning is brown pigmentation of the skin and nails; similar pigmentations have occurred to the palperbral and bulbar conjunctiva regions of the eye, as well as swelling and pigmentation of the eyelids and ocular discharges (especially hypersecretions from the Meibomian glands).62, 63 Similar symptoms have been found in children born from mothers with this condition.64 There is no evidence of damage to the optic nerve heads or fundi and visual fields were normal. Any visual acuity problems arose from hypersecretions from the Meibomian glands.63 Although PCB blood levels were decreasing, abnormalities to the glands and associated discharges were still evident in these patients.65, 66

3.5.1.1.3 Hepatic

Yusho and Yu-Cheng patients experienced elevated serum liver enzymes67, 68 with some evidence of significant increases reported for some patients, suggesting liver damage may have occurred.69

3.5.1.2 Animal (acute) Signs and symptoms related to animal exposure to atrazine are associated with acute, sub-acute, subchronic and chronic studies. These are all summarised together in section 3.5.2.2, according to the various organs. The various PCB individual congeners and mixtures will be described, where appropriate.


3.5.2.1 Human (chronic)

3.5.2.1.1 Skin

Chronic human skin exposure to PCBs can lead to various skin diseases. Following a clinical survey of capacitor manufacturers, who were exposed to airborne concentrations ranging from 0.007 to 11 mg/m3 of various Aroclors, a substantial number of employees had a history of dermatological conditions; these included conjunctival redness, oedema and hyperpigmentation.70 Workers exposed to 0.1 mg/m3 Aroclors (unspecified mixture) for an average of 14 months developed mild forms of chloroacne on the face, especially the cheeks, forehead, and ears.71 As the Aroclors were used for heat exchange, it is possible the workers were exposed to thermal degradation by-products.

65

3.5.2.1.2 Inhalation

Capacitor workers (n = 326) were exposed to mean air concentrations of Aroclors at 0.007–11 mg/m3 for five years.70, 72 Symptoms experienced included upper respiratory tract or eye irritation (48%), cough (14%), and tightness of the chest (10%). There was no control group to compare these symptoms with, however the high percentage (48%) that suffered respiratory and eye irritations suggests these symptoms are related to airborne exposure. Other workers exposed to elevated airborne concentrations of PCB experienced more frequent chest pains while mildly exercising when compared to an unexposed control group.73

3.5.2.1.3 Hepatic

A large number of epidemiological and clinical studies have shown PCB-exposed workers have elevated serum transaminases (where reported, workers and the public were exposed to PCBs for time intervals that spanned from less than one years to 45 years).74-76 Such elevations have, in some studies, been correlated with serum PCB levels.74, 75, 77 In one study, there was evidence the patients may have also experienced asymptomatic hepatomegaly,20, 78 which may be indicative of microsomal induction. Levels of PCB in the workplace ranged from 48 to 275 µg/m3 and workers had elevated blood PCB levels between 41–1,319 µg/kg. There is conflicting evidence that serum cholesterol and triglyceride levels change following chronic exposure to PCBs.74

3.5.2.1.4 Endocrine

There is conflicting epidemiological evidence of changes to thyroid hormone status as a result of environmental exposure. Some studies find an overall negative association,79 no association80, 81 and in other studies it has been reported as positive.82 In animal models, however, the relationship is clearer with evidence that PCB mixtures may cause disruption of thyroid hormone (see section 3.5.2.2.4), though administered doses were orders of magnitude greater that values expected from environmental exposure. Similar conflicts can be found with occupational studies.


There is some suggestion there may be subtle immunological changes occurring as a result of workplace and environmental exposure. These have included increased susceptibility to respiratory tract infections,67, 68 increased prevalence of ear infections in children83 and changes in markers of host immunity (serum IgA and IgM antibody levels, and/or changes in T lymphocyte subsets).84 The evidence is subtle,85 conflicting20, 74, 76 and confounded by exposures to other contaminants such as various other chlorinated organic molecules.86 However, evidence following PCB exposure in animals is more compelling (see section 3.5.2.2.6).

3.5.2.1.6 Neurological

Most of the research conducted on the neurological effects of PCB has been on neonates and young children. A summary of this research is present in section 3.8.1.

3.5.2.1.7 Reproduction

Evidence on reproduction changes in humans as a result of environmental exposure to PCBs is not compelling. The number of pregnancies in female capacitor workers exposed to a range of PCBs (Aroclors 1254, 1242, and/or 1016) for at least three months showed no apparent decrease.87 Eating fish contaminated with PCB may lead to a statistically significant reduction in mean menstrual cycle length.88 However such

66

a change, according to the authors, is not likely to be clinically significant or be of major public health concern. Further studies also suggest no statistical changes occur in fecundity and conception delays following such a diet.89

3.5.2.2 Animal (sub chronic and chronic)

The following paragraphs in this section are a brief (and by no means exhaustive) summary of a range of studies assessing the effects of dietary acute, sub-chronic and chronic ingestion of PCB mixtures upon various organs of the body in an assortment of different animal species. For the sake of brevity, minimal examples from each section will be used to reinforce each important point. This section is a brief summary of the ATSDR report summarising PCB toxicity 90; further information can be sought from this exhaustive publication.

3.5.2.2.1 Cardiovascular and Haematology

There is conflicting evidence that PCBs may cause cardiovascular injury. One investigation suggested pericardial oedema can occur following chronic administration of Aroclor 1248 (12 mg/kg/day) to primates over three months.91 Both chronic (up to 0.08 mg/kg/day for 25 months)92 and acute (4,000 mg/kg)93 studies have shown no effect which has also been reproduced in a number of other studies involving rhesus monkeys and rats respectively.


Hepatic

There is sufficient evidence to suggest that hepatotoxicity of PCBs can occur in animals exposed to commercial mixtures or single congeners for acute, sub-chronic and chronic durations by various routes of exposure. Hepatic toxic effects following exposure to PCBs include induction of liver microsomes, enlargement of the liver, plasma elevation of hepatic transaminases and lipids, altered metabolism of porphyrin and vitamin A that can lead to degenerative lesions (non-neoplastic) which may progress to tumours, depending on the dose and duration of exposure. Mice, rabbits, and guinea pigs exposed to Aroclor 1254 for 7 hours/day for 150 days (inhalation, 1.5 mg/m3) developed hepatic histopathologic lesions,94 with severity varying for each species. Such severe lesions have not been reproduced in other similar studies. Sub-chronic oral exposure (0.03 – 25 mg/kg/day for 4 weeks) of various Aroclors showed induction of liver microsomes95-97 and some suggestion liver enlargement might have occurred. At the highest concentration administered, there were increased liver triglycerides evident. Histological evidence of injury has been achieved following oral acute doses in excess of 4,000 mg/kg93; there was evidence of fatty degeneration and necrotic changes. Subacute doses at or above 50 mg/kg or doses between 6.5 and 50 mg/kg/day chronic may cause similar changes to the liver.93, 98, 99 Long term chronic studies (doses ranging from 2 – 11 mg/kg/day) employing PCB mixtures that specifically target the liver have demonstrated liver changes such as increases in relative liver weight, microsomal induction, elevated serum

67

transaminases, non-neoplastic lesions, and/or tumors.100, 101. Other studies with different species have reported similar hepatic effects.

Renal

Chronic ingestion of PCBs (1.0-10 mg/kg/day; Aroclor 1254) has caused increased kidney weight, biochemical evidence of compromised renal function and cortical tubular protein casts.102, 103 Sub-acute studies suggest histopathological changes (100 mg/kg/day; Aroclor 1242),93 though further subchronic and chronic studies employing doses ranging from 1.5 – 11 mg/kg did not demonstrate histopathological changes.96,

100


Thyroid

There is some evidence that PCBs may induce thyroid toxicity in animal models as well as a variety of changes in thyroid hormone levels. Evidential effects upon both the thyroid and thyroid hormone system are consistent across most animal studies. The intensity of effects is dose dependent. Acute studies have shown doses (2.3 mg/kg/day for 7 days; Aroclor 1254) caused depression of serum T4 and depression of T3.

104 At 18 mg/kg/day (for 7 days) there was no change in thyroid structure.105 Sub-chronic administration of Aroclor 1254 (44 mg/kg/day for 6 weeks) caused thyroid lesions, with evidence of recovery when the treatment ceased.106 In chronic investigations (28 months), enlarged thyroid glands and follicles with desquamated cells were noted in Rhesus monkeys107 (0.2 mg/kg/day; Aroclor 1254). Numerous other studies have demonstrated similar results, though variability may occur with strain and species. The effect of PCB on neonates’ thyroids, while in gestation, has been reported and evidence suggests pups are susceptible to thyroid injury. Pregnant rats exposed to Aroclor 1254 (2.5 or 25 mg/kg/day) for the duration of their pregnancy have borne pups that had thyroidal lesions and depressed levels of serum T4 and T3.

108

Adrenal Gland

Studies have demonstrated elevated (8.1 mg/kg/day; Aroclor 1254; 2 weeks),109 unchanged93 (1.5 mg/kg/day ;Aroclor 1242; 2–6 months ) and depressed levels110 (0.25 mg/kg/day; Aroclor 1254, 5 months) of serum corticosterone. There is also an evidence of degenerative change to the gland.

Reproduction

In contrast to human investigations, exposure to PCB in animal models has resulted in reproductive toxicity, though doses necessary to achieve this are large (ranging from0.4 to 10 mg/kg/day). Observed effects described in a wide range of reports in mice, rats, mink and monkeys, include prolonged oestrus cycle,111 decreased sexual receptivity, impaired ability to conceive,112 partial or total reproductive failure113 and prolonged menstruation and decreased fertility. A full summary of this research is described in the toxicological profile for polychlorinated biphenyls.90

68

3.5.2.2.4 Musculoskeletal, Dermal and Body weight

Musculoskeletal

There is minimal information on muscular skeletal changes occurring in animal models. Most studies suggest there are no histological changes following chronic exposure (4 – 11 mg/kg/day of Aroclor 1016, 1242, 1254, or 1260, for 24 months) of a range of PCB mixtures114.

Dermal and Occular

Numerous studies have shown chronic oral ingestion of PCB in monkeys and hairless mice causes similar symptoms of chloroacne to those observed in humans. Doses as low as 0.1 – 0.2 mg/kg/day (Aroclor 1248; 2 months) have caused these symptoms.112 They include progressive alopecia, periorbital oedema, acne, fingernail loss, and gingival hyperplasia and necrosis of varying severity.107 Topical application of 50 mg Arochlor 1254 to hairless mice resulted in death within 24 hours, whereas animals remained asymptomatic following chronic exposure of 1-2 mg doses.115 Applications of Phenclor 54 resulted in chloroacne and moderate metabolic disturbances.115 Ocular exudate, inflamed and prominent Meibomian glands, distorted growth of finger and toe nails where demonstrated in monkeys chronically exposed to 0.005 mg Aroclor 1254 per kg body weight.217, 218

Body weight

Studies in rats have demonstrated decreases in body weight following oral exposure to various PCB mixtures. Acute doses at or above 25 mg/kg are required for effects on weight117 whereas pronounced decreases occur in sub-chronic doses ranging from 0.1 to 25 mg/kg body weight.102 The birth weights of the infants born to females receiving the 18 mg Aroclor 1016 per kg body weight were significantly less than those of the control infants.116

3.5.2.2.5 Immunological and Haematological

Haematological

Chronic investigations have suggested various PCBs may cause aenemia; for example, administration of 0.2 mg/kg/day to rats for 12–28 months118 caused decreases in decreased haemoglobin content, decreased haematocrit, and hypocellularity of erythrocytic cells. However, such changes do not occur throughout all species of animals tested such as in guinea pigs119 or in rabbits120.

Immunological

Various animal models have been exposed to a variety of PCB mixtures and congeners; there is substantial evidence that sub-chronic and chronic oral dose of PCB, for example 4 mg/kg/day, can lead to host immunity changes. Such changes may occur to various immunological indicators including decreased thymus weight,91 reduced weight of the spleen121 and reduced antibody responses to specific antigens,122 increased susceptibility to infection123 and increased natural killer cell activity.124

Ocular exudate, inflamed and prominent Meibomian glands, distorted growth of finger and toe nails where demonstrated in monkeys chronically exposed to 0.005 mg Aroclor 1254 per kg body weight.217, 218

69

Decreased antibody (IgG and IgM) responses to sheep erythrocytes where shown in monkeys chronically exposed to 0.005 mg Aroclor 1254 per kg body weight.219-221


Ocular

Ocular effects resulting from PCB exposure were documented in monkeys following sub-chronic exposure112 (as low as 0.1 mg/kg/day; Aroclor 1248; 2 months) and are similar to those observed for human exposure. Reported signs and symptoms have included swelling and reddening of the eyelid and eyelid discharge.

Neurobehavioural

The neurobehavioural effects resulting from PCB exposure in various animal models can be summarised into two categories90: the effects on motor activity and the effects on higher function such as learning and memory. The effects on motor function are wide and varied and difficult to summarise. Sub-acute doses of PCB (Aroclor 1254) can lead to decreased spontaneous motor activity.125 Post-natal exposure via breast milk also suppressed this activity but in utero exposure did not126; however results vary, depending on the animal species and PCB mixture. The effects on higher function were also varied, depending on the test performed, formulation and species. Most were undertaken on animals perinatally. Various degrees of modifications have been reported and conclusions are difficult to make. It is suggested ortho-substituted PCB congeners are more active than coplanar PCBs in modifying cognitive processes.127

Neurochemical

The most consistent finding in animal models for neurochemical effects is decreased dopamine concentrations in different areas of the brain. For example, at high doses of PCB (500 or 1,000 mg/kg; Aroclor 1254 and 1260; acute) there was a dose-dependent decrease in the levels of dopamine in the caudate nucleus.128 Similar results were found with chronic administration of PCBs. In contrast, there appear to be no changes in brain neurochemistry of noradrenaline, adrenaline, or serotonin.129

3.5.2.2.7 Gastrointestinal

Acute doses of PCB (100 mg/kg/day Aroclor 1254) can cause ulcers in the pig stomach.130 Similar evidence was found in a series of sub-chronic studies with monkeys.131 Severity of the histopathological changes was dependent on both exposure length and dose.

3.6 Carcinogenicity There is adequate information to suggest PCBs can cause carcinomas in rats. Studies have shown, for example, increased numbers of hepatocarcinomas,100, 150, 151 thyroid tumours,100 and gastrointestinal tract152 tumours during exposures to PCBs. Based on investigations on animal studies with some supporting epidemiological evidence both the United States EPA153 and IARC154 have classified PCBs as probable human carcinogens. Both the National Toxicology Program155 and the

70

American Conference of Governmental Industrial Hygienists156 have concluded there is enough evidence to classify PCBs as known animal carcinogens. Research in these animal models also suggests that the position and number of chlorine atoms on the biphenyl molecule may be an important determinant of carcinogenicity.100 Evidence from human studies, unfortunately, has not been conclusive. There is a slight increased risk of melanomas and brain cancer deaths following chronic workplace exposure to PCBs.157, 158 A retrospective cohort study following mortality of all forms of cancer amongst 2567 workers who had at least three months exposure to PCBs in capacitor factories showed increases in liver and rectal cancers, but neither was statistically significant.159 Another cohort study investigating 14,458 employees (male and female) concluded there was a significant relationship between PCB exposures and increased mortality of biliary liver and gall bladder cancers.160 Among women, there were also significant increases in intestinal cancer and diseases of the nervous system. An update from Brown and Jones159 investigations concluded that the small number of deaths related to liver and intestinal cancers, myeloma and nervous system diseases could not be clearly associated with PCB exposure.161 Other investigations have suggested relationships with non-Hodgkin’s lymphoma,157, 162 and mortalities associated with pancreatic carcinoma,163 brain carcinoma and melanoma.157 Nevertheless, although human studies have not been conclusive, PCBs should be considered potential human carcinogens.154

3.7 Mutagenicity There is no evidence PCBs are mutagenic. PCBs have been tested both in vitro and in vivo with little evidence of genotoxic activity. There is no evidence of chromosomal abnormalities or dominant lethal mutations in animal models.164, 165 PCBs are also not mutagenic in in vitro cell lines or in the bacteria S. typhimurium.166, 167 However, there is some experimental evidence suggesting mechanistic involvement in the potential development of cancer; metabolites of lower chlorinated congeners may bind to and covalently modify cellular constituents including proteins and DNA.168, 169


3.8.1 Humans It was only following the ingestion of large doses of PCBs present in contaminated oil, as what occurred in Taiwan and Japan (Yusho and Yu-Cheng), that evidence of adverse developmental effects became apparent64 (estimated concentrations of PCBs in the oil in Japan were approximately 0.1%133, for example). Babies subsequently born to mothers, who were exposed to these PCBs while pregnant, displayed features defined as ‘foetal PCB syndrome’.11 In Taiwan, pre-natal exposure led to evidence of low birth weight, delays in developmental milestones and reduced IQs when compared with control children who were not severely contaminated.170 A further follow-up study confirmed these observations noting continuing growth impairment and reduced IQ and clumsy movements.171 The event in Japan, though similar to Taiwan, caused less pronounced neurological inhibitions.172

71

However, as summarised in section 3.10.1.3, it has been suggested that contaminated oils also contained PCDFs and polychlorinated quaterphenyls, which may have contributed to these symptoms. In contrast to the Asian incidents, other studies assessing the impact of consuming contaminated food products have resulted in conflicting evidence on pregnancy outcomes. For example, one cohort study reported a slight increase in decreased birth weight, gestational age, head circumference and behaviour in the neonate whose mothers ingested contaminated fish during the duration of their gestation,173 whereas a similar investigation linked to contaminated fish from another source suggested there is no risk.173 Other studies looking at pre-natal exposure to contaminated food have concluded there is an association with slight neurological defects on new-born and developing infants; reported syndromes have included average IQ values being marginally lower than less contaminated groups174 and observed sub-clinical decrements in cognitive functions.24, 175 However, according to Ross,176 the test results were within normal ranges, most of the differences reported were within the variability of the testing tools, and variations in results were not clinically significant. In addition, in all of the studies assessed by Ross, the mother showed no signs of PCB toxicity. A comprehensive summary of the arguments both for and against developmental deficits following pre-natal exposure to PCB are beyond the scope of this report and any further details can be sought by reading the cited papers.

3.8.2 Animal Studies Maternal toxicity can lead to significant teratogenic effects in animals, though high doses are required to achieve this phenomenon. For example, induction of cleft palate177 required an acute dose of 244 mg/kg (Aroclor 1254) in mice. Such effect on the foetus can occur in the absence of maternal signs and symptoms.178 PCB transfer via the milk to the pup is also a major contributor, though again doses necessary to achieve high concentrations in the milk are high.179 Animal studies, using doses that are orders of magnitude greater than values expected from environmental exposure, have measured distinct changes including significant increases in foetal mortality rates, decreased birth weights, reduced survival to weaning, incidents of hydronephrosis and cleft palate, and gross malformations.111, 222 In contrast to these animal studies, teratogenic effects in humans are subtle and difficult to correlate with exposures making it difficult to form a clear hypothesis on the role of environmental PCBs.176

72



Air

PCBs in the atmosphere have been detected throughout all regions of the world. Air concentrations are typically higher in the winter than the summer months.180, 181. The concentration increases with increasing chlorination and decreases with increasing ortho-substitution. Atmospheric levels of PCBs throughout the world generally appear to be decreasing over time. However, higher levels of PCBs have been determined in urban sites compared to rural locations. Atmospheric concentrations of PCBs assayed in urban and rural Baltimore locations, for example, in 1996 were 0.38–3.36 and 0.02–0.34 ng/m3, respectively.182 In remote areas, such as the Antarctica (1981-1982), concentrations ranged from 0.02–0.18 ng/m3.183

Water

Like air levels, concentrations of PCB in water have been decreasing with time. Lake Michigan water concentrations, for example, have decreased from 1.2 ng/L in 1980 to 0.47 ng/L in 1991.184. PCBs have been detected in seawater throughout the world (concentrations range from 0.02 to 0.59 ng/L). Despite having low water solubility, high concentrations have been measured in water tables surrounding industrial sites. The maximum total PCB concentration measured in groundwater on-site at the General Motors Foundry Operation hazardous waste site was 1,200 µg/L.4

Soil

Soil and sediment levels of PCB have also decreased with time. Sediment cores from various regions of the United States demonstrated decreasing levels of PCB between 1962 and 1997, after sediment concentrations peaked around 1970.185 Surface sediments from Lake Ontario, for example, have decreased from 1.3-1.9 µg/g in 1982/1983 down to 0.08-0.29 µg/g in 1995/1996.186 Concentrations in sediment may also vary with depth with an initial increase but at greater depths decreasing.187 Sediment levels downstream from polluted sites such as industry or large urban areas may also influence residue levels.188 Soil levels are also influenced by adjacent sites of pollution.189

3.9.2 Environmental Fate Widespread and increasing use of PCBs from the 1940s to the 1970s has resulted in extensive contamination of the environment. They are stable molecules that not only resist environmental degradation, but due to their lipophilicity (see section 3.1.1) and long elimination biological half-lives, they have biomagnified in both aquatic and terrestrial food chains.190 Accumulations of PCBs have been reported in a wide range of higher vertebrate species.191-193 Nevertheless, since being banned in 1977, environmental and biological levels are slowly decreasing. This is evident amongst fish stock in the Great Lakes region, a site of high PCB contamination. The degree of contamination in fish has sharply declined since production of PCB was halted in 1977.194 In Canada, for example, PCB levels have been decreasing in both dairy milk and human serum (see section 3.10.1.2). Concentrations in the milk of lactating mothers have also been steadily declining.195, 196 In recent years there is some

73

suggestion the predominant PCB contribution to the next generation is determined more by maternal milk than environmental contributions.197

3.9.3 Ecological Degradation The environmental degradation of chlorinated residues such as PCBs is dependent on three pathways: hydrolysis, biodegradation and photodegradation.198 The ecological fate of the various PCBs will determine which of these mechanisms are dominant for their degradation. Typically, photodegradation will occur with atmospheric PCBs, whereas those complexed in soil will undergo microbiological metabolism and those in aquatic environments will be transformed by hydrolysis. The degree of chlorine saturation as well as the respective isomeric patterns will also influence their degradation. PCBs are sufficiently volatile to cycle between air, water and soil, though the degree of cycling and movement is dependent on the number of attached chlorine atoms.199 The dominant mechanism for their global movement is by atmospheric spread. Those that are heavily chlorinated tend to condense from the atmosphere.199, 200 Those remaining in the atmosphere undergo photodegradation by sunlight-generated oxidants, predominantly OH, O3 and NO3.

201, 202 The calculated atmospheric half- lives for the various PCB congeners in the atmosphere can range from 3 to 500 days, depending on the degree of chlorine saturation203, 204; increasing number of substituted chlorine atoms leads to increasing half-lives. In water, the dominant pathway for PCB degradation is by hydrolysis. This occurs by photolytic cleavage of the bond between the carbon and chlorine, which leads to its replacement with hydrogen.205 With highly saturated biphenyls, there is preferential cleavage of the chlorine atoms at the ortho position205 that not only decreases the number of highly chlorinated PCBs, but increases the number of toxic coplanar (non-ortho) congeners (PCBs 77 and 126). Half-lives of degradation can range from ~60 days to 27 years,206 depending on the degree of chlorine substitution, but also on the water depth and the season. Microbial biodegradation may also occur, but again this is dependent on the degree of chlorination; such degradation occurs primarily by aerobic processes. PCBs strongly absorb to sediment and soil where they can persist, resisting uptake into plants through their roots207; degradation is dominated by aerobic and anaerobic metabolism.208, 209 Both classes of bacteria appear to act in a complementary manner; anaerobic organisms tend to dechlorinate the halogen-saturated biphenyls to form lightly chlorinated ortho-enriched congeners, which are in turn readily degradable by a wide range of different aerobic populations. Aerobic activity degrades PCBs to chlorinated benzoic acid followed by mineralisation to carbon dioxide and inorganic chlorides.210, 211 PCB half-lives in soil can range from 3 years through to 37 years198; such a wide range is influenced by chlorine substitution and position, amongst other physico-chemical parameters.


3.10.1 Doses Necessary to Cause Human Acute and Chronic Toxicity Documentation of doses in humans necessary to cause toxicity is limited to the outbreaks of PCB toxicity in Japan (1968) and Taiwan (1978). Further human

74

exposures are a result of chronic workplace and environmental exposure where researchers have sought to make correlations between exposures (typically by measuring PCB blood levels) and a range of subtle signs and symptoms.

3.10.1.1 Acute Apart from the Yusho and Yu-Cheng incidents, there is little information on acute exposure to PCBs. Dr. Masayasu Goto, a dermatologist, examined 146 Yusho patients and graded their cases according to their clinical severity.132 A summary of their symptoms, the estimated amounts of oil ingested and their calculated doses of PCB ingested are presented in Table 3.3. The authors did not provide information on their respective symptoms, when classifying their severity. Estimated doses ingested are based on a final 0.1% PCB concentration in the oil.133 Table 3.3 A summary of various doses of PCB ingested and their respective severity of symptoms suffered by 146 Yusho patients132

Volume ingested Number of patients

Estimated total dose Severity

12% no symptoms 49% light symptoms

Less than 720 mL

80

Less than 720 mg 39% severe symptoms

31% light symptoms 720-1440 mL

45

720-1440 mg 69% severe symptoms

Greater than 1440 mL

21

Greater than 1440 mg

Majority suffered severe symptoms

Quite clearly, acute exposure to high doses of PCB can lead to adverse effects.

3.10.1.2 Chronic Measures of chronic toxicity due to workplace or environmental exposures are predominantly based on evidence of elevated serum PCB concentrations that are correlated with markers of tissue injury. Patients exposed to PCB have average plasma concentrations that are typically one to three orders of magnitude greater than expected background levels, which for the general population have steadily declined from 6.4 ng/mL134 to 0.9 – 1.5 ng/mL135, 136 over the last 20-30 years. Some reports have suggested no statistically significant adverse effects following chronic exposure to PCBs, whereas other studies have suggested chronic toxicity based on elevated liver function tests, which may possibly reflect liver disease. There are also reports of chloroacne, though this is most likely a result of the dioxin contaminants present in the PCB formulations. An important source of PCB has been the consumption of contaminated fish. Ingestion of fish from Lake Michigan has resulted in higher PCB exposures, when compared with the general population. The mean PCB plasma concentration for one study group was 25.8 µg/L, with one recreational fisherman consuming approximately 40 kg of fish per year that resulted in elevated serum PCB values peaking at 360 µg/L. No adverse effects were identified within this cluster.137 Workplace exposure resulting in elevated mean serum PCB concentrations of 337 and 292 µg/L for current and former workers respectively showed elevated liver transaminases in four workers78 (from a total of 80) and chloroacne. Liver tests were,

75

however, within the range of normal and reported incidents of chloroacne may have been caused by other chlorinated contaminants. The result of another similar study reported elevated urinary porphyrin levels in workers with mean blood PCB levels of 386 µg/L138; again, elevations were within the normal laboratory range.139 In contrast, Smith and associates (1982) reported a correlation between hepatic transaminases and PCB blood concentrations (mean serum PCB values ranged from 17 – 824 µg/L). The author suggests the increases in liver enzymes are possibly a result of microsomal enzyme induction rather than chronic liver disease.

3.10.2 Summary Case Reports Following Toxicological Exposure

Documentation of doses in humans necessary to cause toxicity is limited to the outbreaks of PCB toxicity in Japan (1968) and Taiwan (1978). In both outbreaks of PCB intoxication, the oil that contaminated the rice oil is thought to have originated from a heat-exchange fluid that leaked into the rice oil-refining chamber. When PCBs are heated above 300oC, in the presence of oxygen, PCDFs are formed133; PCDDs, polychlorinated dibenzofurans (PCDFs) and chlorinated quaterphenyls have been measured in contaminated oil sourced from the Yusho outbreak.145 Following this analysis, it has been suggested that these observed signs and symptoms were different from those reported following other occupational exposures and secondly that PCDFs and possibly other contaminants were responsible for the reported morbidity and mortality.145, 146

3.10.2.1 ‘Yusho’ Outbreak in Japan In 1968 approximately 1300 people in Fukuoka, Nagasaki as well as other areas of Western Japan ingested thermally degraded PCB-heat-transfer liquid140 that had contaminated rice oil. The disease was called ‘Yusho’, namely oil disease. It was estimated that PCB concentration in the oil was approximately 0.1%.133 Victims soon developed swelling of upper eyelids, increased eye discharge, chloracne (a severe and persistent form of acne), swelling of limbs and other less reported symptoms such as nausea, fatigue; others developed liver disorders. Babies that were exposed during pre-term later developed neurological symptoms and newborns of exposed mothers displayed discolouration of nails, skin and eyes and later had premature eruption of teeth.60 Five years after the event, 22 deaths had been reported; 9 from malignant neoplasms which included 2 cases of liver cancer.141

3.10.2.2 ‘Yu-Cheng’ Outbreak in Taiwan In February 1979 rice bran oil contaminated with heat degraded PCBs and dibenzofurans (concentrations contaminating the oil ranged from 4.8 – 204.9 µg/mL) was sold in two prefectures in Taiwan (Tai Chung and Chang Hua). Approximately 2000 people subsequently become poisoned as a result of ingesting this oil. PCB blood levels in these patients ranged from 3 to 1156 ng/mL with a mean value of 89.14 ± 6.9.142 Observed acute symptoms amongst these sufferers were similar to those reported in Yusho victims67, 143; developmental and other neurological symptoms were also similar to Yusho contaminated offspring. Increased deaths from non-carcinogenic liver diseases were detected within the space of three years post-exposure.15, 144 Subsequent studies of mortality rates for cancer by the end of 1991 were not statistically elevated when compared to the general population.144

76


3.10.3.1 LD50 values Reported acute LD50 values for PCBs have a relatively wide range of values, but overall indicate a relatively low level of toxicity. This will however depend on the mammalian species and strain, the dilution, and the congener or isomer composition being investigated. Rat oral LD50 values range from 1,001 mg/kg147 (Aroclor 1254) to 11,300 mg/kg148 (Aroclor 1262). A summary of LD50 values for seven mixtures are presented in table 3.4. Table 3.4 Oral toxicity of seven Aroclor mixtures148

Aroclors 1221 1232 1242 1248 1260 1262 1268 Oral LD50 mg/kg (rats) 3,980 4,470 8,650 11,000 10,000 11,300 10,900

To the knowledge of the author, there are no published LD50 values for dermal exposure. Reported minimal lethal doses ranged from in excess of 794 mg/kg for Aroclors 1242 and 1248 to at or below 19,200 mg/kg for Aroclor 5460.148 For an explanation of minimal lethal dose values see Appendix one. Both sets of values (acute oral ingestion and acute dermal exposure) suggest that increasing chlorine saturation leads to decreased toxicity.

3.10.4 Guidance Values

Reference dose

Reference doses for the Monsanto products Aroclor 1016 and 1254, and how they were derived, are presented in table 3.5. Table 3.5 Reference doses for Aroclor 1016 and 1254 as defined by the EPA223

Agency EPA EPA

Type of Exposure Aroclor 1016 Chronic ingestion RfD Aroclor 1254 Chronic ingestion RfD

Value 0.07 µg/kg/day 0.02 µg/kg/day

Year 1996 1996

Species monkey monkey

Critical effect Reduced birth weights Clinical and Immunologic Studies

NOAEL/LOAEL 0.007 mg/kg/day NOAEL 0.005 mg/kg/day LOAEL

Uncertainty factor 100(10A, 10H) 300(10A, 10H, 3L)

Principle study 116 217-221

Other Measures of Daily Intake

In 1973, the United States FDA concluded there is unavoidable PCB contamination of the environment, and therefore have provided limits of exposure in food products for

77

human consumption. They have defined this as ‘temporary tolerances’ for PCB residues ‘to permit the elimination of such contaminants at the earliest practicable time’149. The various categories of food including products for packing of food are presented in Table 3.6. Table 3.6 Temporary tolerances of PCBs in food destined for human consumption, as defined by the FDA149

1.5 parts per million in milk (fat basis). 1.5 parts per million in manufactured dairy products (fat basis). 3 parts per million in poultry (fat basis). 0.3 parts per million in eggs. 0.2 parts per million in finished animal feed for food-producing animals (except the

following finished animal feeds: feed concentrates, feed supplements, and feed premixes).

2 parts per million in animal feed components of animal origin, including fishmeal and other by-products of marine origin and in finished animal feed concentrates, supplements, and premixes intended for food producing animals.

2 parts per million in fish and shellfish (edible portion). The edible portion of fish excludes head, scales, viscera, and inedible bones.

0.2 parts per million in infant and junior foods. 10 parts per million in paper food-packaging material intended for or used with human

food, finished animal feed and any components intended for animal feeds. The tolerance shall not apply to paper food-packaging material separated from the food therein by a functional barrier which is impermeable to migration of PCB's.

Based on the possible health risks and exposure to the general population, the EPA has set PCBs Maximum Contaminant Level Goal in drinking water at zero (see Appendix 2), and the Maximum Contaminant Level in drinking water at 0.0005 ppm (see Appendix 3).

3.11 A Summary of Biological assays That are Available to Test for Evidence of These Chemicals in Biological Samples PCBs, like dioxins, have been assayed in a wide range of human tissues and fluids; these include plasma,212 whole blood,213 adipose tissue,214 milk,213 follicular fluid,215 and serum.215. Determination of PCB follows the same protocols as described for dioxins. Predominantly, only the 12 PCB dioxin-like congeners are determined from serum. More often they are assayed in conjunction with the 17 toxic congeners. For further details, see the dioxin document, section 3.11.

78

3.12 Summary The term polychlorinated biphenyls (PCBs) describes a class of 209 compounds in which a biphenyl molecule can have up to 10 attached chlorine atoms. Those without substitution at the ortho position are known as co-planar PCBs and may activate the aryl hydrocarbon receptor in a manner similar to dioxins. PCBs are relatively insoluble in water which decreases with increasing chlorination. They are stable to thermal and corrosive degradation and thus had wide use in industry. PCBs readily absorb across both skin and the intestinal tract; human exposure is predominantly from processed fish and food resulting in high bioavailability. The percentage uptake will however be influenced by factors such as chlorine saturation of the biphenyl molecule and dietary intake. Upon systemic absorption, PCBs initially target highly perfused organs with later redistribution to organs of high lipid content, overwhelmingly to adipose tissue. Initial metabolism and conjugation within the liver will occur followed by enterohepatic recirculation and accumulation within adipose tissue. Host elimination will vary according to the degree of chlorine saturation; elimination half-lives can therefore vary from 0.3 to 12 years. Signs and symptoms following human acute exposure are predominantly reported in Yu-Cheng and Yusho patients, who live in Taiwan and Japan respectively; these patients in ingested cooking oil contaminated with various congeners of PCB and residues of dioxins. They experienced dermal, mild neurological, hepatic and ophthalmological signs and symptoms as well as fever, vomiting and diarrhoea. Chronic symptoms, resulting from workplace exposure, are not as well defined but include mild dermal, respiratory and hepatic signs and symptoms as well as subtle changes in endocrine, immunological, neurological and reproductive functions. These changes however are often of no clinical consequence. Evidence of these changes in animal models is more compelling, though they were administered excessively large doses that were orders of magnitude greater than what human subjects would be exposed to. As with signs and symptoms, knowledge of toxicity to humans has been limited to Yu-Cheng and Yusho patients. They were exposed to total doses ranging from 750 mg to an excess of 1400 mg; severity of symptoms increased with increasing doses. In animal models, reported LD50 values indicate a relatively low level of toxicity (oral values in the rat ranged from 3.9 to 10.9 g/kg body weight). There is adequate information to suggest PCBs can cause carcinomas in animal models. The US EPA (and IARC), based on investigations on animal studies with some supporting epidemiological evidence, have classified PCBs as probable human carcinogens. In contrast, the National Toxicology Program and the American Conference of Governmental Industrial Hygienists have concluded there is enough evidence to classify PCBs as known animal carcinogens. There is some evidence from human studies that PCBs may cause cancer though this is not as conclusive as has been demonstrated in other species. Exposure to high doses of PCB, as what occurred in Japan and Taiwan, can lead to defined developmental and teratogenic effects. Such effects have also been clearly demonstrated in animal models. However where toxicity occurs from workplace

79

exposure or pollution such effects on the newborn and their development are inconclusive and studies are contradictory. Widespread use of PCBs until the 1970s has resulted in extensive contamination of the environment. They have bioaccumulated and biomagnified in both aquatic and terrestrial food chains. Despite being banned for over 30 years, PCBs are still present in the environment though levels are slowly decreasing. Analysis of PCBs is conducted in a similar manner to dioxins.

3.13 References 1. Ballschmiter K, Zell M. Analysis of Polychlorinated-Biphenyls (Pcb) By Glass-

Capillary Gas-Chromatography - Composition of Technical Aroclor-Pcb and Clophen-Pcb Mixtures. Fresenius Zeitschrift Fur Analytische Chemie. 1980;302:20-31.

2. Safe SH. Polychlorinated biphenyls (PCBs): environmental impact, biochemical and toxic responses, and implications for risk assessment. Critical Reviews in Toxicology. 1994;24:87-149.

3. van den Berg M, Birnbaum L, Bosveld ATC, et al. Toxic equivalency factors (TEFs) for PCBs, PCDDs, PCDFs for humans and wildlife. Environmental Health Perspectives. 1998;106:775-792.

4. EPA. (Environmental Protection Agency). Ambient water quality criteria for polychlorinated biphenyls. Washington, DC: U.S. Environmental Protection Agency, Criteria and Standards Division, Office of Water Regulations and Standards. EPA 440/5-80-068. Washington, DC; 1980.

5. Schecter A, Charles K. The Binghamton State office building transformer incident after one decade. Chemosphere. 1991;23:1307-1321.

6. Rappe C, Buser HR, Bosshardt HP. Dioxins, dibenzofurans, and other polyhalogenated aromatics: Production, use, formation, and destruction. Annals of the New York Academy of Sciences. 1979;320:1-18.

7. Van Den Berg M, Olic K, Hutzinger O. Polychlorinated dibenzofurans (PCDFs): Environmental occurrence and physical, chemical and biological properties. Toxicological and Environmental Chemistry. 1985;9:171-217.

8. Afghan BK, Chau ASY. Analysis of Trace rganics in the Aquatic Environment. Boca Raton, FL: CRC Press, Inc; 1989.

9. Safe S. Polychlorinated-biphenyls (PCBS) and polybrominated biphenyls (PBBS) - biochemistry, toxicology, and mechanism of action. Crc Critical Reviews in Toxicology. 1984;13:319-395.

10. Welsh MS. Extraction and gas chromatography/electron capture analysis of polychlorinated biphenyls in railcar paint scrapings. Applied Occupational and Environmental Hygiene. 1995;10:175-181.

11. Jones GRN. Polychlorinated biphenyls: where do we stand now? the Lancet. 1989;334:791-792.

12. Nakanishi Y, Shigematsu N, Kurita Y, et al. Respiratory involvement and immune status in yusho patients. Environmental Health Perspectives. 1985;59:31-6.

13. Urabe H, Asahi M. Past and current dermatological status of Yusho patients. Environmental Health Perspectives. 1985;59:11-15.

14. Lu YC, Wu YC. Clinical findings and immunological abnormalities in Yu-Cheng patients. Environmental Health Perspectives. 1985;59:17-29.

80

15. Hsu ST, Ma CI, Hsu SK, et al. Discovery and epidemiology of PCB poisoning in Taiwan: a four-year followup. Environmental Health Perspectives. 1985;59:5-10.

16. Wester RC, Bucks DA, Maibach HI, Anderson J. Polychlorinated biphenyls (PCBs): dermal absorption, systemic elimination, and dermal wash efficiency. Journal of Toxicology and Environmental Health. 1983;12:511-519.

17. Fisher HL, Shah PV, Sumler MR, Hall LL. Invivo and Invitro Dermal Penetration of 2,4,5,2',4',5'-Hexachlorobiphenyl in Young and Adult-Rats. Environmental Research. 1989;50:120-139.

18. Wolff MS, Thornton J, Fischbein A, Lilis R, Selikoff IJ. Disposition of polychlorinated biphenyl congeners in occupationally exposed persons. Toxicology and Applied Pharmacology. 1982;62:294-306.

19. Phillips DL, Smith AB, Burse VW, Steele GK, Needham LL, Hannon WH. Half-life of polychlorinated biphenyls in occupationally exposed workers. Archives of Environmental Health. 1989;44:351-4.

20. Maroni M, Colombi A, Cantoni S, Ferioli E, Foa V. Occupational exposure to polychlorinated biphenyls in electrical workers. I. Environmental and blood polychlorinated biphenyls concentrations. British Journal of Industrial Medicine. 1981;38:49-54.

21. Lees PSJ, Corn M, Breysse PN. Evidence for dermal absorption as the major route of body entry during exposure of transformer maintenance and repairmen to PCBs. American Industrial Hygiene Association Journal. 1987;48:257-264.

22. Baars AJ, Bakker MI, Baumann RA, et al. Dioxins, dioxin-like PCBs and non-dioxin-like PCBs in foodstuffs: occurrence and dietary intake in The Netherlands. Toxicology Letters. 2004;151:51-61.

23. Liem AK, Fürst P, Rappe C. Exposure of populations to dioxins and related compounds. Food Additives and Contaminants. 2000;17:241-259.

24. Patandin S, Dagnelie PC, Mulder PGH, et al. Dietary exposure to polychlorinated biphenyls and dioxins from infancy until adulthood: A comparison between breast-feeding, toddler, and longterm exposure. Environmental Health Perspectives. 1999;107:45-51.

25. Albro PW, Fishbein L. Intestinal-absorption ofpolychlorinated biphenyls in rats. Bulletin of Environmental Contamination and Toxicology. 1972;8:26-31.

26. Allen JR, Norback DH, Hsu IC. Tissue modifications in monkeys as related to absorption, distribution, and excretion of polychlorinated biphenyls. Archives of Environmental Contamination and Toxicology. 1974;2:86-95.

27. Wolff MS. Occupational exposure to polychlorinated biphenyls (PCBs). Environmental Health Perspectives. 1985;60:133-138.

28. van Birgelen APJM, van den Berg M. Toxicokinetics. Food Additives and Contaminants. 2000;17:267-273.

29. Matthews HB, Dedrick RL. Pharmacokinetics of PCBs. Annual Review of Pharmacology & Toxicology. 1984;24:85-103.

30. Schlummer M, Moser GA, McLachlan MS. Digestive tract absorption of PCDD/Fs, PCBs, and HCB in humans: mass balances and mechanistic considerations. Toxicology and Applied Pharmacology. 1998;152:128-37.

31. Moser GA, McLachlan MS. The influence of dietary concentration on the absorption and excretion of persistent lipophilic organic pollutants in the human intestinal tract. Chemosphere. 2001;45:201-211.

32. Yakushiji T, Watanabe I, Kuwabara K, et al. Rate of decrease and half-life of polychlorinated biphenyls (PCBs) in the blood of mothers and their children

81

occupationally exposed to PCBs. Archives of Environmental Contamination & Toxicology. 1984;13:341-5.

33. Hashimoto K, Akasaka S, Takagi Y, et al. Distribution and Excretion of [C-14]Polychlorinated Biphenyls After Their Prolonged Administration to Male Rats. Toxicology and Applied Pharmacology. 1976;37:415-423.

34. Clevenger MA, Roberts SM, Lattin DL, Harbison RD, James RC. The Pharmacokinetics of 2,2',5,5'-tetrachlorobiphenyl and 3,3',4,4'-tetrachlorobiphenyl and Its relationship to toxicity. Toxicology and Applied Pharmacology. 1989;100:315-327.

35. IPCS. (International Program on Chemical Safety) Polychlorinated biphenyls and terphenyls. 2nd ed. Environmental Health Criteria 140. Geneva: World Health Organisation; 1993.

36. Chia LG, Chu FL. Neurological studies on polychlorinated biphenyl (PCB)-poisoned patients. American Journal of Industrial Medicine. 1984;5:117-26.

37. Chen PH, Wong CK, Rappe C, Nygren M. Polychlorinated biphenyls, dibenzofurans and quaterphenyls in toxic rice-bran oil and in the blood and tissues of patients with PCB poisoning (Yu-Cheng) in Taiwan. Environmental Health Perspectives. 1985;59:59-65.

38. Dewailly E, Mulvad G, Pedersen HS, et al. Concentration of organochlorines in human brain, liver, and adipose tissue autopsy samples from Greenland. Environmental Health Perspectives. 1999;107:823-829.

39. Masuda Y, Kuroki H, Haraguchi K, Nagayama J. PCB and PCDF congeners in the blood and tissues of yusho and yu-cheng patients. Environmental Health Perspectives. 1985;59:53-8.

40. Thoma H, Mucke W, Kauert G. Comparison of the Polychlorinated Dibenzo-Para-Dioxin and Dibenzofuran in Human Tissue and Human Liver. Chemosphere. 1990;20:433-442.

41. Kreuzer PE, Csanady GA, Baur C, et al. 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) and congeners in infants. A toxicokinetic model of human lifetime body burden by TCDD with special emphasis on its uptake by nutrition. Archives of Toxicology. 1997;71:383-400.

42. Guvenius DM, Hassanzadeh P, Bergman A, Noren K. Metabolites of polychlorinated biphenyls in human liver and adipose tissue. Environmental Toxicology & Chemistry. 2002;21:2264-9.

43. Schnellmann RG, Volp RF, Putnam CW, Sipes IG. The hydroxylation, dechlorination, and glucuronidation of 4,4'-dichlorobiphenyl (4-DCB) by human hepatic microsomes. Biochemical Pharmacology. 1984;33:3503-9.

44. Safe S. Polychlorinated biphenyls (PCBs), dibenzo-p-dioxins (PCDDs), dibenzofurans (PCDFs), and related compounds: environmental and mechanistic considerations which support the development of toxic equivalency factors (TEFs). Critical Reviews in Toxicology. 1990;21:51-88.

45. Schuur AG, van Leeuwen-Bol I, Jong WM, et al. In vitro inhibition of thyroid hormone sulfation by polychlorobiphenylols: isozyme specificity and inhibition kinetics. Toxicological Sciences. 1998;45:188-94.

46. Lans MC, Klasson-Wehler E, Willemsen M, Meussen E, Safe S, Brouwer A. Structure-dependent, competitive interaction of hydroxy-polychlorobiphenyls, -dibenzo-p-dioxins and -dibenzofurans with human transthyretin. Chemico-Biological Interactions. 1993;88:7-21.

47. Brandt I, Lund J, Bergman A, Klassonwehler E, Poellinger L, Gustafsson JA. Target-cells for the polychlorinated biphenyl metabolite 4,4'-

82

Bis(Methylsulfonyl)-2,2',5,5'-tetrachlorobiphenyl in lung and kidney. Drug Metabolism and Disposition. 1985;13:490-496.

48. Chu SG, Covaci A, Jacobs W, Haraguchi K, Schepens P. Distribution of methyl sulfone metabolites of polychlorinated biphenyls and p,p '-DDE in human tissues. Environmental Health Perspectives. 2003;111:1222-1227.

49. Haraguchi K, Kuroki H, Masuda Y. Synthesis and characterization of tissue-retainable methylsulfonyl polychlorinated biphenyl isomers. Journal of Agricultural and Food Chemistry. 1987;35:178-182.

50. Steele G, Stehr-Green P, Welty E. Estimates of the biologic half-life of polychlorinated biphenyls in human serum. New England Journal of Medicine. 1986;314:926-927.

51. Chen PH, Luo ML, Wong CK, Chen CJ. Comparative rates of elimination of some individual polychlorinated biphenyls from the blood of PCB-poisoned patients in Taiwan. Food & Chemical Toxicology. 1982;20:417-25.

52. Brown JF, Lawton RW, Ross MR, Feingold J, Wagner RE, Hamilton SB. Persistence of PCB congeners in capacitor workers and Yusho patients. Chemosphere. 1989;19:829-834.

53. Bühler F, Schmid P, Schlatter C. Kinetics of PCB elimination in man. Chemosphere. 1988;17:1717-1726.

54. Schecter A, Stanley J, Boggess K, et al. Polychlorinated biphenyl levels in the tissues of exposed and nonexposed humans. Environmental Health Perspectives. 1994;102 Suppl 1:149-58.

55. Mes J, Arnold DL, Bryce F. The elimination and estimated half-lives of specific polychlorinated biphenyl congeners from the blood of female monkeys after discontinuation of daily dosing with Aroclor 1254. Chemosphere. 1995;30:789-800.

56. Hara I, Harada H, Kimura S, Endo T, Kawano K. Follow up health examination in an electric condenser factory after cessation of PCBs usage (1st report) [in Japanese]. Japanese Journal of Industrial Health. 1974;16:365-366.

57. Lung SC, Guo YL, Chang HY. Serum concentrations and profiles of polychlorinated biphenyls in Taiwan Yu-cheng victims twenty years after the incident. Environmental Pollution. 2005;136:71-9.

58. Parkinson A, Safe SH, Robertson LW, et al. Immunochemical quantitation of cytochrome-P-450 isozymes and epoxide hydrolase in liver-microsomes rrom polychlorinated or polybrominated biphenyl-treated rats - a study of structure-activity-relationships. Journal of Biological Chemistry. 1983;258:5967-5976.

59. McConnell EE. Acute and chronic toxicity and carcionogenesis in animals. In: Kimbrough RD, Jensen AA, eds. Halogenated Biphenyls, Terphenyls, Naphthalenes, Dibenzodioxins and Related Products. 2 ed. Amsterdam: Elsevier; 1989:161-193.

60. Kuratsune M, Yoshimura T, Matsuzaka J, Yamaguchi A. Yusho, a poisoning caused by rice oil contaminated with polychlorinated biphenyls. HSMA health reports. 1972;86:1083-1091.

61. Rice RH, Cohen DE. Toxic responses of the skin. In: Klaassen CD, ed. Casarett and Doull’s toxicology: The Basic Science of Poisons. New York, NY: McGraw-Hill; 1996:529-546.

62. Allen JR, Norback DH. Polychlorinated biphenyl-induced and triphenyl-induced gastric mucosal hyperplasia in primates. Science. 1973;179:498-499.

63. Fu YA. Ocular manifestation of polychlorinated biphenyl (PCB) intoxication. Its relationship to PCB blood concentration. Archives of Ophthalmology. 1983;101:379-81.

83

64. Yamashita F, Hayashi M. Fetal PCB syndrome: clinical features, intrauterine growth retardation and possible alteration in calcium metabolism. Environmental Health Perspectives. 1985;59:41-5.

65. Kohno T, Ohnishi Y, Hironaka H. Ocular manifestations and polychlorinated biphenyls in the tarsal gland contents of yusho patients. Fukuoka Igaku Zasshi. 1985;76:244-247.

66. Nakamura T, Miyazaki W, Ohnishi Y, Ishibashi T. Ophthalmic findings in Yusho. Journal of Dermatological Science. 2005:S57-S63.

67. Rogan WJ. Yu-Cheng. In: Kimbrough RD, Jensen AA, eds. Halogenated Biphenyls, Terphenyls, Naphthalenes, Dibenzodioxins, and Related Products. Amsterdam: Elsevier; 1989.

68. Kuratsune M. Yusho, with reference to Yu-Cheng. In: Kimbrough RD, Jensen AA, eds. Halogenated biphenyls, terphenyls, naphthalenes, dibenzodioxins and related products. Amsterdam: Elsevier Science Publishers; 1989:381-400.

69. Masuda Y. The Yusho rice oil poisoning incident. In: Schecter A, ed. Dioxins and health. New York, NY: Plenum Press; 1994:633-659.

70. Fischbein A, Wolff MS, Lilis R, Thornton J, Selikoff I J. Clinical findings among PCB-exposed capacitor manufacturing workers. Annals of the New York Academy of Sciences. 1979;320:703-715.

71. Meigs JW, Albom JJ, Kartin BL. Chloracne from an unusual exposure to Aroclor. JAMA. 1954;154:1417-1418.

72. Warshaw R, Fischbein A, Thornton J, Miller A, Selikoff IJ. Decrease in vital capacity in PCB-exposed workers in a capacitor manufacturing facility. Annals of the New York Academy of Sciences. 1979;320:277-283.

73. Emmett EA, Maroni M, Schmith JM, Levin BK, Jefferys J. Studies of transformer repair workers exposed to PCBs: I. Study design, PCB concentrations, questionnaire, and clinical examination results. American Journal of Industrial Medicine. 1988;13:415-427.

74. Chase KH, Wong O, Thomas D, Berney BW, Simon RK. Clinical and metabolic abnormalities associated with occupational exposure to polychlorinated biphenyls (PCBs). Journal of Occupational Medicine. 1982;24:109-114.

75. Emmett EA, Maroni M, Jefferys J, Schmith J, Levin BK, Alvares A. Studies of transformer repair workers exposed to PCBs: II. Results of clinical laboratory investigations. American Journal of Industrial Medicine. 1988;14:47-62.

76. Smith AB, Schloemer J, Lowry LK, et al. Metabolic and health consequences of occupational exposure to polychlorinated biphenyls. British Journal of Industrial Medicine. 1982;39:361-369.

77. Baker EL, Jr., Landrigan PJ, Glueck CJ, et al. Metabolic consequences of exposure to polychlorinated biphenyls (PCB) in sewage sludge. American Journal of Epidemiology. 1980;112:553-563.

78. Maroni M, Colombi A, Arbosti G, Cantoni S, Foa V. Occupational exposure to polychlorinated biphenyls in electrical workers. II. Health effects. British Journal of Industrial Medicine. 1981;38:55-60.

79. Osius N, Karmaus W, Kruse H, Witten J. Exposure to polychlorinated biphenyls and levels of thyroid hormones in children. Environmental Health Perspectives. 1999;107:843-849.

80. Longnecker MP, Gladen BC, Patterson DG, Rogan WJ. Polychlorinated biphenyl (PCB) exposure in relation to thyroid hormone levels in neonates. Epidemiology. 2000;11:249-254.

84

81. Nagayama J, Okamura K, Iida T, et al. Postnatal exposure to chlorinated dioxins and related chemicals on thyroid hormone status in Japanese breast-fed infants. Chemosphere. 1998;37:1789-1793.

82. Langer P, Tajtakova M, Fodor G, et al. Increased thyroid volume and prevalence of thyroid disorders in an area heavily polluted by polychlorinated biphenyls. European Journal of Endocrinology. 1998;139:402-9.

83. Smith BJ. P.C.B. levels in human fluids: Sheboygan case study. Technical report WIS-SG-83-240. Madison, WI: University of Wisconsin Sea Grant Institute; 1984.

84. Lawton RW, Ross MR, Feingold J, Brown JF, Jr. Effects of PCB exposure on biochemical and hematological findings in capacitor workers. Environmental Health Perspectives. 1985;60:165-184.

85. Svensson BG, Hallberg T, Nilsson A, Schutz A, Hagmar L. Parameters of immunological competence in subjects with high consumption of fish contaminated with persistent organochlorine compounds. International Archives of Occupational & Environmental Health. 1994;65:351-358.

86. Dewailly E, Ayotte P, Bruneau S, Gingras S, Belles-Isles M, Roy R. Susceptibility to infections and immune status in Inuit infants exposed to organochlorines. Environmental Health Perspectives. 2000;108:205-211.

87. Taylor PR, Stelma JM, Lawrence CE. The relation of polychlorinated biphenyls to birth weight and gestational age in the offspring of occupationally exposed mothers. American Journal of Epidemiology. 1989;129:395-406.

88. Mendola P, Buck GM, Sever LE, Zielezny M, Vena JE. Consumption of PCB-contaminated freshwater fish and shortened menstrual cycle length. American Journal of Epidemiology. 1997;146:955-960.

89. Buck GM, Sever LE, Mendola P. Consumption of contaminated sport fish from Lake Ontario and time-to-pregnancy. American Journal of Epidemiology. 1997;146:949-954.

90. ATSDR. Toxicological Profile for Polychlorinated Biphenyls. Atlanta, Ga: Department of Health and Human Services, Public Health Service. Agency for Toxic Substances and Disease Registry; 2000.

91. Allen JR, Abrahamson LJ, Norback DH. Biological effects of polychlorinated biphenyls and triphenyls on the subhuman primate. Environmental Research. 1973;6:344-354.

92. Arnold DL, Nera EA, Stapley R, et al. Toxicological consequences of Aroclor 1254 ingestion by female rhesus (Macaca mulatta) monkeys and their nursing infants. Part 3: post-reproduction and pathological findings. Food & Chemical Toxicology. 1997;35:1191-1207.

93. Bruckner JV, Khanna KL, Cornish HH. Biological responses of the rat to polychlorinated biphenyls. Toxicology and Applied Pharmacology. 1973;24:434-448.

94. Treon JF, Cleveland FP, Cappel JW, Atchley RW. The toxicity of the vapours of Aroclor 1242® and Aroclor® 1254. American Industrial Hygiene Association journal. 1956;Q17:204-213.

95. Litterst CL, Farber TM, Baker AM, Van Loon EJ. Effect of polychlorinated biphenyls on hepatic microsomal enzymes in the rat. Toxicology and Applied Pharmacology. 1972;23:112-122.

96. Bruckner JV, Khanna KL, Cornish HH. Effect of prolonged ingestion of polychlorinated biphenyls on the rat. Food and Cosmetics Toxicology. 1974;12:323-330.

85

97. Bruckner JV, Jiang WD, Brown JM, Putcha L, Chu CK, Stella VJ. The influence of ingestion of environmentally encountered levels of a commercial polychlorinated biphenyl mixture (Aroclor 1254) on drug metabolism in the rat. The Journal of Pharmacology and Experimental Therapeutics. 1977;202:22-31.

98. Koller LD. Enhanced polychlorinated biphenyl lesions in Moloney leukemia virus-infected mice. Clinical Toxicology. 1977;11:107-116.

99. Kimbrough RD, Gaines TB, Linder RE. Morphological changes in livers of rats fed polychlorinated biphenyls: Light microscopy and ultrastructure. Archives of Environmental Health. 1972;25:354-&.

100. Mayes BA, McConnell EE, Neal BH, et al. Comparative Carcinogenicity in Sprague-Dawley Rats of the Polychlorinated Biphenyl Mixtures Aroclors 1016, 1242, 1254, and 1260. Toxicological Sciences. 1998;41:62-76.

101. Fish KM, Mayes BA, Brown JFJea. Biochemical measurements on hepatic tissues from SD rats fed Aroclors 1016, 1242, 1254, and 1250. Toxicologist. 1997;36:87.

102. Andrews JE. Polychlorinated biphenyl (Aroclor 1254) induced changes in femur morphometry calcium metabolism and nephrotoxicity. Toxicology. 1989;57:83-96.

103. Gray LE, Jr., Ostby J, Marshall R, Andrews J. Reproductive and thyroid effects of low-level polychlorinated biphenyl (Aroclor 1254) exposure. Fundamental and Applied Toxicology. 1993;20:288-294.

104. Hood A, Hashmi R, Klaassen CD. Effects of microsomal enzyme inducers on thyroid-follicular cell proliferation, hyperplasia, and hypertrophy. Toxicology and Applied Pharmacology. 1999;160:163-170.

105. Price SC, Ozalp S, Weaver R, Chescoe D, Mullervy J, Hinton RH. Thyroid hyperactivity caused by hypolipodaemic compounds and polychlorinated biphenyls: The effect of coadministration in the liver and thyroid. Archives of Toxicology. 1988;12S:85-92.

106. Collins WT, Jr., Capen CC. Ultrastructural and functional alterations of the rat thyroid gland produced by polychlorinated biphenyls compared with iodide excess and deficiency, and thyrotropin and thyroxine administration. Virchows Archiv. B, Cell Pathology Including Molecular Pathology. 1980;33:213-231.

107. Tryphonas L, Arnold DL, Zawidzka Z, Mes J, Charbonneau S, Wong J. A pilot study in adult rhesus monkeys (M. mulatta) treated with Aroclor 1254 for two years. Toxicologic Pathology. 1986;14:1-10.

108. Collins WT, Jr., Capen CC. Fine structural lesions and hormonal alterations in thyroid glands of perinatal rats exposed in utero and by the milk to polychlorinated biphenyls. American Journal of Pathology. 1980;99:125-42.

109. Sanders OT, Zepp RL, Kirkpatrick RL. Effect of PCB ingestion on sleeping times, organ weights, food consumption, serum corticosterone and survival of albino mice. Bulletin of Environmental Contamination & Toxicology. 1974;12:394-399.

110. Byrne JJ, Carbone JP, Pepe MG. Suppression of serum adrenal cortex hormones by chronic low-dose polychlorobiphenyl or polybromobiphenyl treatments. Archives of Environmental Contamination and Toxicology. 1988;17:47-53.

111. Brezner E, Terkel J, Perry AS. The effect of Aroclor 1254 (PCB) on the physiology of reproduction in the female rat--I. Comparative Biochemistry and Physiology. C, Comparative Pharmacology and Toxicology. 1984;77:65-70.

112. Barsotti DA, Marlar RJ, Allen JR. Reproductive dysfunction in rhesus monkeys exposed to low levels of polychlorinated biphenyls (Aoroclor 1248). Food & Chemical Toxicology. 1976;14:99-103.

86

113. Aulerich RJ, Ringer RK. Current status of PCB toxicity to mink, and effect on their reproduction. Archives of Environmental Contamination & Toxicology. 1977;6:279-92.

114. Chu I, Villeneuve DC, Yagminas A, et al. Toxicity of 2,2',4,4',5,5'-hexachlorobiphenyl in rats: effects following 90-day oral exposure. Journal of Applied Toxicology. 1996;16:121-128.

115. Puhvel SM, Sakamoto M, Ertl DC, Reisner RM. Hairless mice as models for chloracne: a study of cutaneous changes induced by topical application of established chloracnegens. Toxicology and Applied Pharmacology. 1982;64:492-503.

116. Barsotti DA, Van Miller JP. Accumulation of a commercial polychlorinated biphenyl mixture (Aroclor 1016) in adult rhesus monkeys and their nursing infants. Toxicology. 1984;30:31-44.

117. Brown NM, Lamartiniere CA. Xenoestrogens alter mammary gland differentiation and cell proliferation in the rat. Environmental Health Perspectives. 1995;103:708-713.

118. Arnold DL, Mes J, Bryce F, et al. A pilot study on the effects of Aroclor 1254 ingestion by rhesus and cynomolgus monkeys as a model for human ingestion of PCBs. Food and Chemical Toxicology. 1990;28:847-857.

119. Vos JG, van Driel-Grootenhuis L. PCB-induced suppression of the humoral and cell-mediated immunity in guinea pigs. Science of the Total Environment. 1972;1:289-302.

120. Street JC, Sharma RP. Alteration of induced cellular and humoral immune responses by pesticides and chemicals of environmental concern: quantitative studies of immunosuppression by DDT, aroclor 1254, carbaryl, carbofuran, and methylparathion. Toxicology and Applied Pharmacology. 1975;32:587-602.

121. Silkworth JP, Loose LD. Cell-mediated immunity in mice fed either Aroclor 1016 or hexachlorobenzene. Toxicology and Applied Pharmacology. 1978;45:326-332.

122. Exon JH, Talcott PA, Koller LD. Effect of lead, polychlorinated biphenyls, and cyclophosphamide on rat natural killer cells, interleukin 2, and antibody synthesis. Fundamental & Applied Toxicology. 1985;5:158-164.

123. Imanishi J, Nomura H, Matsubara M, et al. Effect of polychlorinated biphenyl on viral infections in mice. Infection and Immunity. 1980;29:275-7.

124. Smialowicz RJ, Andrews JE, Riddle MM, Rogers RR, Luebke RW, Copeland CB. Evaluation of the immunotoxicity of low level PCB exposure in the rat. Toxicology. 1989;56:197-211.

125. Kodavanti PR, Derr-Yellin EC, Mundy WR, et al. Repeated exposure of adult rats to Aroclor 1254 causes brain region-specific changes in intracellular Ca2+ buffering and protein kinase C activity in the absence of changes in tyrosine hydroxylase. Toxicology and Applied Pharmacology. 1998;153:186-98.

126. Pantaleoni GC, Fanini D, Sponta AM, Palumbo G, Giorgi R, Adams PM. Effects of maternal exposure to polychlorobiphenyls (PCBs) on F1 generation behavior in the rat. Fundamental & Applied Toxicology. 1988;11:440-9.

127. Schantz SL, Moshtaghian J, Ness DK. Spatial learning deficits in adult rats exposed to ortho-substituted PCB congeners during gestation and lactation. Fundamental and Applied Toxicology. 1995;26:117-26.

128. Seegal RF, Brosch KO, Bush B. Regional alterations in serotonin metabolism induced by oral exposure of rats to polychlorinated biphenyls. Neurotoxicology. 1986;7:155-65.

87

129. Seegal RF, Bush B, Shain W. Lightly chlorinated ortho-substituted PCB congeners decrease dopamine in nonhuman primate brain and in tissue culture. Toxicology and Applied Pharmacology. 1990;106:136-44.

130. Hansen LG, Wilson DW, Byerly CS. Effects on growing swine and sheep of two polychlorinated biphenyls. American Journal of Veterinary Research. 1976;37:1021-1024.

131. Allen JR. Response of the nonhuman primate to polychlorinated biphenyl exposure. Federation Proceedings. 1975;34:1675-1679.

132. Kuratsune M, Yoshimura T, Matsuzaka J, Yamaguchi A. Epidemiologic study on Yusho, a Poisoning Caused by Ingestion of Rice Oil Contaminated with a Commercial Brand of Polychlorinated Biphenyls. Environmental Health Perspectives. 1972;1:119-128.

133. Roach JAG. Nonmetalic alteration of PCBs. In: S. WJ, ed. PCBs and the Environment. Boca Raton: CRC; 1986:207-213.

134. Kreiss K, Roberts C, Humphrey HE. Serial PBB levels, PCB levels, and clinical chemistries in Michigan's PBB cohort. Archives of Environmental Health. 1982;37:141-7.

135. Anderson HA, Falk C, Hanrahan L, et al. Profiles of Great Lakes critical pollutants: a sentinel analysis of human blood and urine. The Great Lakes Consortium. Environmental Health Perspectives. 1998;106:279-89.

136. Hanrahan LP, Falk C, Anderson HA, Draheim L, Kanarek MS, Olson J. Serum PCB and DDE levels of frequent Great Lakes sport fish consumers-a first look. The Great Lakes Consortium. Environmental Research. 1999;80:S26-S37.

137. Andersen HA. General population exposure to environmental concentrations of halogenated biphenyls. In: Kimbrough RD, Jensen AA, eds. Halogenated biphenyls, terphenyls, naphthalenes, dibenzodioxins and related products. Amsterdam: Elsevier North-Holland Biomedical Press; 1989:325-344.

138. Colombi A, Maroni M, Ferioli A, et al. Increase in urinary porphyrin excretion in workers exposed to polychlorinated biphenyls. Journal of Applied Toxicology. 1982;2:117-121.

139. Hill R, Bailey SL, Needham LL. Development and utilization of a procedure for measuring urinary porphyrins by high performance liquid chromatography. Journal of Chromatography. 1982;232:251-260.

140. Kunita N, Kashimoto T, Miyata H, Fukushima S, Hori S, Obana H. Causal agents of Yusho. American Journal of Industrial Medicine. 1984;5:45-58.

141. Kuratsune M. Yusho. In: Kimbrough RD, ed. Halogenated biphenyls, terphenyls, naphthalenes, dibenzodioxins and related products. Amsterdam: Elsevier; 1980.

142. Ko YC, Jao LT, Cheng CT, Hsu ST, Hsiao HC, Hu HT. [The blood level of PCB in the poisoned patients (author's transl)]. Taiwan i Hsueh Hui Tsa Chih - Journal of the Formosan Medical Association. 1981;80:774-9.

143. Rogan WJ, Gladen BC, Hung KL, et al. Congenital Poisoning By Polychlorinated-Biphenyls and Their Contaminants in Taiwan. Science. 1988;241:334-336.

144. Hsieh SF, Yen YY, Lan SJ, Hsieh CC, Lee CH, Ko YC. A cohort study on mortality and exposure to polychlorinated biphenyls. Archives of Environmental Health. 1996;51:417-424.

145. Kashimoto T, Murata H. Differences between Yusho and other kinds of poisoning involving only PCBs. In: S. WJ, ed. PCBs and the Environment. Boca Raton: CRC; 1987:1-26.

146. Yoshimura T. Yusho in Japan. Industrial Health. 2003;41:139-148.

88

147. Garthoff LH, Cerra FE, Marks EM. Blood-Chemistry Alterations in Rats After Single and Multiple Gavage Administration of Polychlorinated Biphenyl. Toxicology and Applied Pharmacology. 1981;60:33-44.

148. Fishbein L. Toxicity of Chlorinated Biphenyls. Annual Review of Pharmacology and Toxicology. 1974;14:139-156.

149. FDA. (Food and Drug Administration) Unavoidable contaminants in food for human consumption and food-packaging material. Tolerances for polychlorinated biphenyls (PCB's). Code of Federal Regulations 12:2; 21CFR109.30, April 1, 2008. Available at <http://www.accessdata.fda.gov/scripts/cdrh/cfdocs/cfcfr/CFRSearch.cfm?fr=109.30&SearchTerm=pcb> (Last accessed on Jan 17 2009); 2008.

150. Kimbrough RD, Squire RA, Linder RE, Strandberg JD, Montalli RJ, Burse VW. Induction of liver tumor in Sherman strain female rats by polychlorinated biphenyl aroclor 1260. Journal of the National Cancer Institute. 1975;55:1453-1459.

151. Schaeffer E, Greim H, Goessner W. Pathology of chronic polychlorinated biphenyl (PCB) feeding in rats. Toxicology and Applied Pharmacology. 1984;75:278-88.

152. Morgan RW, Ward JM, Hartman PE. Aroclor 1254-induced intestinal metaplasia and adenocarcinoma in the glandular stomach of F344 rats. Cancer Research. 1981;41:5052-5059.

153. EPA. (Environmental Protection Agency). Health effects of PCBs. Washington DC; 2005. In Engel, L.S., Q. Lan, and N. Rothman, Polychlorinated biphenyls and non-Hodgkin lymphoma. Cancer Epidemiology Biomarkers & Prevention, 2007. 16(3): p. 373-376.

154. IARC. Polychlorinated biphenyls and polybrominated biphenyls. IARC monographs on the evaluation of carcinogenic risks to humans. Lyon, France: World Health Organization, International Agency for Research on Cancer; 1987; 18:7.

155. NTP. (National Toxicology Program). Fifth annual report on carcinogens: summary 1989. NTP 89-239, USDHHS. ResearchTriangle Park, NC; 1989:239. In ATSDR. Toxicological Profile for Polychlorinated Biphenyls. Atlanta, Ga: Department of Health and Human Services, Public Health Service. Agency for Toxic Substances and Disease Registry; 2000.

156. ACGIH. American Conference of Governmental Industrial Hygienists. Threshold Limit Values for Chemical Substances and Physical Agents; 1996.

157. Ruder AM, Hein MJ, Nilsen N, et al. Mortality among workers exposed to polychlorinated biphenyls (PCBs) in an electrical capacitor manufacturing plant in Indiana: an update. Environmental Health Perspectives. 2006;114:18-23.

158. Sinks T, Steele G, Smith AB, Watkins K, Shults RA. Mortality among workers exposed to polychlorinated biphenyls. American Journal of Epidemiology. 1992;136:389-398.

159. Brown DP, Jones M. Mortality and industrial hygiene study of workers exposed to polychlorinated biphenyls. Archives of Environmental Health. 1981;36:120-9.

160. Prince MM, Ruder AM, Hein MJ, et al. Mortality and exposure response among 14,458 electrical capacitor manufacturing workers exposed to polychlorinated biphenyls (PCBs). Environmental Health Perspectives. 2006;114:1508-14.

161. Prince MM, Hein MJ, Ruder AM, Waters MA, Laber PA, Whelan EA. Update: cohort mortality study of workers highly exposed to polychlorinated biphenyls

89

(PCBs) during the manufacture of electrical capacitors, 1940-1998. Environmental Health: A Global Access Science Source. 2006;5:13.

162. Engel LS, Lan Q, Rothman N. Polychlorinated biphenyls and non-Hodgkin lymphoma. Cancer Epidemiology Biomarkers & Prevention. 2007;16:373-376.

163. Yassi A, Tate R, Fish D. Cancer mortality in workers employed at a transformer manufacturing plant. American Journal of Industrial Medicine. 1994;25:425-437.

164. Green S, Sauro FM, Friedman L. Lack of dominant lethality in rats treated with polychlorinated biphenyls (Aroclors 1242 and 1254). Food and Cosmetics Toxicology. 1975;13:507-510.

165. Green S, Carr JV, Palmer KA, Oswald EJ. Lack of cytogenetic effects in bone marrow and spermatagonial cells in rats treated with polychlorinated biphenyls (aroclors 1242 and 1254). Bulletin of Environmental Contamination and Toxicology. 1975;13:14-22.

166. Schoeny RS, Smith CC, Loper JC. Non-mutagenicity for Salmonella of the chlorinated hydrocarbons aroclor 1254, 1,2,4-trichlorobenzene, mirex and kepone. Mutation Research. 1979;68:125-132.

167. Hattula ML. Mutagenicity of PCBs and their pyrosynthetic derivatives in cell-mediated assay. Environmental Health Perspectives. 1985;60:255-257.

168. Robertson LW, Gupta RC. Metabolism of polychlorinated biphenyls (PCBs) generates electrophiles and reactive oxygen species that damage DNA. In: Williams G, Aruoma OI, eds. Molecular Drug Metabolism and Toxicology: OICA International (UK) Ltd; 2000.

169. Silberhorn EM, Glauert HP, Robertson LW. Carcinogenicity of polyhalogenated biphenyls: PCBs and PBBs. Critical Reviews in Toxicology. 1990;20:440-496.

170. Guo YL, Lambert GH, Hsu CC, Hsu MM. Yucheng: health effects of prenatal exposure to polychlorinated biphenyls and dibenzofurans. International Archives of Occupational & Environmental Health. 2004;77:153-8.

171. Chen YC, Guo YL, Hsu CC, Rogan WJ. Cognitive development of Yu-Cheng ("oil disease") children prenatally exposed to heat-degraded PCBs. JAMA. 1992;268:3213-3218.

172. Yoshimura T. [Epidemiological study on Yusho babies born to mothers who had consumed oil contaminated by PCB]. Fukuoka Igaku Zasshi. 1974;65:74-80.

173. Benya TJ, Leber AP. Halogneated Cyclic Hydrocarbons. In: Clayton GD, Clayton FE, eds. Patty's Industrial Hygiene and Toxicology. New York: John Wiley & Sons Inc.; 1994.

174. Jacobson JL, Jacobson SW. Intellectual impairment in children exposed to polychlorinated biphenyls in utero. New England Journal of Medicine. 1996;335:783-789.

175. Walkowiak J, Wiener JA, Fastabend A, Heinzow B, Krämer A, Schmidt E. Environmental exposure to polychlorinated biphenyls and quality of the home environment: effects on psychodevelopment in early childhood. the Lancet. 2001;358:1602-1607.

176. Ross G. The public health implications of polychlorinated biphenyls (PCBs) in the environment. Ecotoxicology & Environmental Safety. 2004;59:275-91.

177. Zhao F, Mayura K, Harper N, Safe SH, Phillips TD. Inhibition of 3,3',4,4',5-pentachlorobiphenyl-induced fetal cleft palate and immunotoxicity in C57BL/6 mice by 2,2',4,4',5,5'-hexachlorobiphenyl. Chemosphere. 1997;34:1605-1613.

90

178. Spencer F. An assessment of the reproductive toxic potential of Aroclor 1254 in female Sprague-Dawley rats. Bulletin of Environmental Contamination & Toxicology. 1982;28:290-297.

179. Overmann SR, Kostas J, Wilson LR, Shain W, Bush B. Neurobehavioral and somatic effects of perinatal PCB exposure in rats. Environmental Research. 1987;44:56-70.

180. Ockenden WA, Prest HF, Thomas GO, Sweetman A, Jones KC. Passive air sampling of PCBs: Field calculation of atmospheric sampling rates by triolein-containing semipermeable membrane devices. Environmental Science and Technology. 1998;32:1538–1543.

181. Haugen JE, Wania F, Lei YD. Polychlorinated biphenyls in the atmosphere of southern Norway. Environmental Science and Technology. 1999;33:2340-2345.

182. Offenberg JH, Baker J. Influence of Baltimore's urban atmosphere on organic contaminants over the northern Chesapeake Bay. Journal of the Air and Waste Management Association. 1999;49:959-965.

183. Tanabe S, Hidaka H, Tatsukawa R. PCBs and chlorinated hydrocarbon pesticides in Antarctic atmosphere and hydrosphere. Chemosphere. 1983;12:227-288.

184. Pearson RF, Hornbuckle KC, Eisenreich SJ, Swackhamer DL. PCBs in Lake Michigan water revisited. Environmental Science and Technology. 1996;30:1429-1436.

185. Van Metre PC, Wilson JT, Callender E, Fuller CC. Similar rates of decrease of persistent, hydrophobic and particle-reactive contaminants in riverine systems. Environmental Science and Technology. 1998;32:3312-3317.

186. Oliver BG, Charlton MN, Durham RW. Distribution, redistribution, and geochronology of polychlorinated biphenyl congeners and other chlorinated hydrocarbons in Lake Ontario sediments. Environmental Science and Technology. 1989;23:200-208.

187. Pereira WE, Hostettler FD, Cashman JR, Nishioka RS. Occurrence and distribution of organochlorine compounds in sediment and livers of striped bass (Morone saxatilis) from the San Francisco Bay-Delta Estuary. Marine Pollution Bulletin. 1994;28:434-441.

188. Vanier C, Sylvestre M, Planas D. Persistence and fate of PCBs in sediments of the Saint Lawrence River. Science of the Total Environment. 1996;192:229-244.

189. Bryant CJ, Hartle RW, Crandall MS. Polychlorinated biphenyl, polychlorinated dibenzo-p-dioxin, and polychlorinated dibenzofuran contamination in PCB disposal facilities. Chemosphere. 1989;18:569-576.

190. Risebrough RW, Rieche P, Herman SG, Peakall DB, Kirven MN. Polychlorinated biphenyls in the global ecosystem. Nature. 1968;220:1098-1102.

191. Horn EG, Hetling LJ, Tofflemire TJ. The problem of PCBs in the Hudson River system. Annals of the New York Academy of Sciences. 1979;320:591-609.

192. Veith GD, Kuehl DW, Leonard EN, Welch K, Pratt G. Polychlorinated biphenyls and other organic chemical residues in fish from major United States watersheds near the Great Lakes, 1978. Pesticides Monitoring Journal. 1981;15:1-8.

193. Dustman EH, Stickle LF. The occurance of pesticide residues in wild animals. Annals of the New York Academy of Sciences. 1969;160:162-172.

194. D'Itri FM. Contaminants in selected fishes from the upper Great Lakes. In: Schmidtke NW, ed. Toxic Contamination in Large Lakes. Chelsea, MI: Lewis Publishers; 1988.

91

195. Fürst P. Dioxins, polychlorinated biphenyls and other organohalogen compounds in human milk. Levels, correlations, trends and exposure through breastfeeding. Molecular Nutrition and Food Research. 2006;50:922-933.

196. Zietz BP, Hoopmann M, Funcke M, Huppmann R, Suchenwirth R, Gierden E. Long-term biomonitoring of polychlorinated biphenyls and organochlorine pesticides in human milk from mothers living in northern Germany. International Journal of Hygiene and Environmental Health. 2008;211:624-638.

197. Nawrot TS, Staessen JA, Den Hond EM, et al. Host and environmental determinants of polychlorinated aromatic hydrocarbons in serum of adolescents. Environmental Health Perspectives. 2002;110:583-589.

198. Sinkkonen S, Paasivirta J. Degradation half-life times of PCDDs, PCDFs and PCBs for environmental fate modeling. Chemosphere. 2000;40:943-949.

199. Wania F, Mackay D. Tracking the distribution of persistent organic pollutants. Environmental Science & Technology. 1996;30:A390-A396.

200. Wania F, Mackay D. Global Fractionation and Cold Condensation of Low Volatility Organochlorine Compounds in Polar-Regions. Ambio. 1993;22:10-18.

201. Klöpffer W. Photochemical degradation of pesticides and other chemicals in the environment: a critical assessment of the state of the art. Science of the Total Environment. 1992;123/124:145-159.

202. Atkinson R, Carter WPL. Kinetics and mechanisms of the gas-phase reactions of ozone with organic compounds under atmospheric conditions. Chemical Reviews. 1984;84:437-479.

203. Atkinson R. Estimation of OH radical reaction-rate constants and atmospheric lifetimes for Polychlorobiphenyls, dibenzo-dara-dioxins, and dibenzofurans. Environmental Science & Technology. 1987;21:305-307.

204. Leifer A. Atmospheric oxidation of chlorinated biphenyls. Environmental Transport and Transformations of Polychlorinated Biphenyls: Environmental Protection Agency; 1983.

205. Barr JR, Oida T, Kimata K, et al. Photolysis of environmentally important PCBs. Organohalogen Compounds. 1997;33:199-204.

206. Bunce NJ, Landers JP, Langshaw J, Nakai JS. An assessment of the importance of direct solar degradation of some simple chlorinated benzenes and biphenyls in the vapor phase. Environmental Science & Technology. 1989;23:213±218.

207. Gan DR, Berthouex PM. Disappearance and crop uptake of PCBs from sludge-amended farmland. Water Environment Research. 1994;66:54-69.

208. Abramowicz DA. Aerobic and anaerobic PCB biodegradation in the environment. Environmental Health Perspectives. 1995;103 Suppl 5:97-99.

209. Field JA, Sierra-Alvarez R. Microbial transformation and degradation of polychlorinated biphenyls. Environmental Pollution. 2008;155:1-12.

210. Thomas DJ, Tracey B, Marshall H, Norstrom RJ. Arctic terrestrial ecosystem contamination. Science of the Total Environment. 1992;122:135-164.

211. Robinson GK, Lenn MJ. The bioremediation of polychlorinated biphenyls (PCBs): problems and perspectives. Biotechnology & Genetic Engineering Reviews. 1994;12:139-188.

212. Van Oostdam JC, Dewailly E, Gilman A, et al. Circumpolar maternal blood contaminant survey, 1994-1997 organochlorine compounds. Science of the Total Environment. 2004;330:55-77.

213. Schecter A, Ryan JJ, Papke O. Decrease in levels and body burden of dioxins, dibenzofurans, PCBS, DDE, and HCB in blood and milk in a mother nursing twins over a thirty-eight month period. Chemosphere. 1998;37:1807-1816.

92

214. McCready D, Aronson KJ, Chu W, Fan W, Vesprini D, Narod SA. Breast tissue organochlorine levels and metabolic genotypes in relation to breast cancer risk Canada. Cancer Causes and Control. 2004;15:399-418.

215. Younglai EV, Foster WG, Hughes EG, Trim K, Jarrell JF. Levels of environmental contaminants in human follicular fluid, serum, and seminal plasma of couples undergoing in vitro fertilization. Archives of Environmental Contamination & Toxicology. 2002;43:121-126.

216. MfE. New Zealands National Implementation Plan under the Stockholm convention on persistent organic Pollutants. Available from <http://www.mfe.govt.nz/publications/hazardous/stockholm-convention-pops-dec06/stockholm-convention-pops-dec06.pdf>. (Last accessed on March 10 2008). Wellington, New Zealand: Minstery for the Environment; 2006.

217. Arnold DL, Bryce F, Stapley R, et al. Toxicological Consequences of Aroclor-1254 Ingestion By Female Rhesus (Macaca-Mulatta) Monkeys .1a. Prebreeding Phase - Clinical Health Findings. Food and Chemical Toxicology. 1993;31:799-810.

218. Arnold DL, Bryce F, Karpinski K, et al. Toxicological Consequences of Aroclor-1254 Ingestion By Female Rhesus (Macaca-Mulatta) Monkeys .1b. Prebreeding Phase - Clinical and Analytical Laboratory Findings. Food and Chemical Toxicology. 1993;31:811-824.

219. Tryphonas H, Luster MI, White KL, et al. Effects of PCB (Aroclor-1254) On Nonspecific Immune Parameters in Rhesus (Macaca-Mulatta) Monkeys. International Journal of Immunopharmacology. 1991;13:639-648.

220. Tryphonas H, Luster MI, Schiffman G, et al. Effect of Chronic Exposure of PCB (Aroclor-1254) On Specific and Nonspecific Immune Parameters in the Rhesus (Macaca-Mulatta) Monkey. Fundamental and Applied Toxicology. 1991;16:773-786.

221. Tryphonas H, Hayward S, Ogrady L, et al. Immunotoxicity Studies of PCB (Aroclor-1254) in the Adult Rhesus (Macaca-Mulatta) Monkey - Preliminary-Report. International Journal of Immunopharmacology. 1989;11:199-206.

222. Haake JM, Safe S, Mayura K, Phillips TD. Aroclor-1254 As an Antagonist of the Teratogenicity of 2,3,7,8-Tetrachlorodibenzo-Para-Dioxin. Toxicology Letters. 1987;38:299-306.

223. IRIS. (Integrated Risk Information Systems) U.S. Environmental Protection Agency, Office of Health and Environmental Assessment, Environmental Criteria and Assessment Office. Cincinnati, OH; 2001.

93

4.0 Dioxins

4.1 Physical Chemical Properties As there are 210 different dibenzo-p-dioxins and dibenzofurans, each with their own distinct properties, a summary of each is beyond the scope of this report; however, brief summaries of the most toxic dioxins are presented in Table 4.1.

4.1.1 Basic Properties Table 4.1 Fundamental physical chemical properties of the various dioxin homologues Homologue Vapour pressure

(mmHg at 25 oC) log Kow Solubility

(mg l−1 at 25 oC) Henry’s constant atm-m3/mol (at 25 oC)

TCDD 8.1 × 10−7 6.4 3.5 × 10−4 1.35 × 10−3 PeCDD 7.3 × 10−10 6.6 1.2 × 10−4 1.07 × 10−4 HxCDD 5.9 × 10−11 7.3 4.4 × 10−6 1.83 × 10−3 HpCDD 3.2 × 10−11 8.0 2.4 × 10−6 5.14 × 10−4 OCDD 8.3 × 10−13 8.2 7.4 × 10−8 2.76 × 10−4 TCDF 2.5 × 10−8 6.2 4.2 × 10−4 6.06 × 10−4 PeCDF 2.7 × 10−9 6.4 2.4 × 10−4 2.04 × 10−4 HxCDF 2.8 × 10−10 7.0 1.3 × 10−5 5.87 × 10−4 HpCDF 9.9 × 10−11 7.9 1.4 × 10−6 5.76 × 10−4 OCDF 3.8 × 10−12 8.8 1.4 × 10−6 4.04 × 10−5

4.1.2 Brief notes The term dioxins summarises a range of chlorinated dibenzenes known as polychlorinated dibenzo-p-dioxins (PCDDs) and dibenzofurans (PCDFs). These fundamental structures differ by their interconnecting groups that link the two aromatic rings (Figure 4.1). In the case of PCDDs, the benzene rings are joined by two oxygen bridges, and in the case of the PCDFs, the benzene rings are connected by a carbon bond and an oxygen bridge (a furan). Figure 4.1 The structural formule for (a) Dibenzo-para-dioxin and (b) dibenzofuran.

Both the PCDDs and the PCDFs are characterised by the number of attached chlorine atoms on the two aromatic rings that can range in number from 1 to 10. There are a total of 210 possible compounds (75 PCDDs and 135 PCDFs) and each is called a congener. Molecules with the same degree of chlorine saturation are defined as a homolog; for example, PCDDs with four chlorine molecules are defined as a homolog

94

and are termed tetrachlorinated dibenzodioxin (TCDD). Different substitutions of chlorine within a homolog are known as isomers. The PCDD and PCDF congeners that have chlorine atoms attached to positions 2, 3, 7 and 8 are of environmental and toxicological importance, and of these congeners 2,3,7,8-TCDD (TCDD) is regarded as the most toxic (see Figure 4.2). Thus, of the 210 known chlorinated dioxins and furans, only 17 are regarded as toxic. To enable calculation of the overall toxicity for these congeners for any given complex mixture of dioxins humans may be exposed to, a value known as Toxic Equivalent Factor (TEF), has been assigned to each of these 17 toxic congeners. These values are based on the relative potency compared to TCDD.

Figure 4.2 The structural formula of 2,3,7,8-tetra CDD. Though there have been several schemes to define these factors, the International Toxic Equivalents Factor was adopted1 and later modified and ratified by the WHO.2 These factors are commonly expressed as I-TEQ mg per mass of defined biological matrix (such as serum lipid). As TCDD is the most toxic congener, it has been assigned a value of 1 and the others, which are less toxic than but relative to TCDD (see Table 4.2 for a complete summary of these values),2 have been assigned numbers ranging from 0.0001 to 1. A small number of coplanar- and mono-ortho-PCBs have also been assigned equivalent factors but are of relative low toxicity, compared to the toxic dioxins. The toxicity of any mixture of PCDDs and PCDFs can be determined by multiplying the concentrations of each congener by its respective TEF and the resulting summation of these values is defined as a toxic equivalent (TEQ), with SI units identical to the concentrations of the individual congeners. This scheme has been recently modified by the WHO to include 12 coplanar congeners of PCBs, as these have ‘dioxin-like’ properties.2 Table 4.2 Toxic Equivalent Factors for the toxic dioxins and PCBs.

Congener WHO TEF Dioxins 2,3,7,8-Tetra-CDD 1 1,2,3,7,8-Penta-CDD 1 1,2,3,4,7,8-Hexa-CDD 0.1 1,2,3,6,7,8-Hexa-CDD 0.1 1,2,3,7,8,9-Hexa-CDD 0.1 1,2,3,4,6,7,8-Hepta-CDD 0.01 OCDD 0.0001 Dibenzofurans 2,3,7,8-Tetra-CDF 0.1 1,2,3,7,8-Penta-CDF 0.05 2,3,4,7,8-Penta-CDF 0.5 1,2,3,4,7,8-Hexa-CDF 0.1 1,2,3,6,7,8-Hexa-CDF 0.1 1,2,3,7,8,9-Hexa-CDF 0.1 2,3,4,6,7,8-Hexa-CDF 0.1

95

1,2,3,4,6,7,8-Hepta-CDF 0.01 1,2,3,4,7,8,9-Hepta-CDF 0.01 OCDF 0.0001 Coplanar PCBs 3,3’,4,4’-TCB (77) 0.0001 3,4,4’,5-TCB (81) 0.0001 3,3’,4,4’,5-PeCB (126) 0.1 3,3’,4,4’,5,5’-HxCB (169) 0.01 Mono-ortho-PCBs 2,3,3’,4,4’-PeCB (105) 0.0001 2,3,4,4’,5-PeCB (114) 0.0005 2,3’,4,4’,5-PeCB (118) 0.0001 20,3,4,4’,5-PeCB (123) 0.0001 2,3,3’,4,4’,5-HxCB (156) 0.0005 2,3,3’,4,4’,5’-HxCB (157) 0.0005 2,3’,4,4’,5,5’-HxCB (167) 0.00001 2,3,3’,4,4’,5,5’-HpCB (189) 0.0001

4.1.3 Uses No uses have been identified for dioxins, though small amounts are synthesized for use in research. It has been hypothesised that if chlorine, carbon, hydrogen and oxygen come into contact with heat, the byproducts can include trace levels of PCDDs and PCDFs.3 Indeed, research has shown that the ideal temperature for the formation of dioxins, based on chlorophenol as a precursor, is between 250oC and 550oC (Figure 4.3). This research, however, was based solely on the solid residues after the formation reaction and did not assay the gas phase (the temperature profile could therefore theoretically differ in the gas phase).

Figure 4.3 The (relative) formation of dioxins; each value is normalized to the 400oC case. Reproduced with permission.4 Essentially, dioxins are a product of industrialisation though they can also form as a result of geological processes and natural combustion.5-7 Dioxins can also form naturally in sewage sludge, forest soils and sediments, and in compost under normal environmental conditions.8, 9 There is also evidence of dioxin formation in historical times.10 However, since the 1930s, there has been a substantial increase in their formation as a result of large-scale production and use of chlorinated chemicals. They have become unwanted contaminants, resulting from industrial processes including

96

incineration, the bleaching of paper and pulp, and from the manufacture and application of chlorinated herbicides (such as pentachlorophenol and 2,4,5-T), pesticides and fungicides.11 Nevertheless, since the mid-1960s, following awareness of the toxicity of dioxins and the increasing levels in the environment and subsequent food chains, there has been a concerted effort to minimise their release. An inventory of dioxin emissions from numerous known sources in the United States for the years 1987 to 1995 have shown an almost 80% decline.12 Accumulations in the environment, food chain and body burdens in humans have all shown corresponding declines.

4.2 Routes of Exposure Workplace exposure to dioxins can result in dermal and inhalation uptake. Any ingestion will be as a result of contaminated food. For the general population, the predominant route of exposure is from the ingestion of food, with minimal input from the inhalation of air or skin contact with contaminated soil.

4.2.1 Ingestion A major contributor to the human body burden of dioxin is the ingestion of food contaminated with dioxins, particularly dairy and poultry products (body burden is expanded in section 4.3.4). It has been estimated that in excess of 90% of the dioxin humans are exposed to is food sourced from animal fat.12 The various dioxin congeners have always been present in food though this has risen sharply since the 1930’s, with the advent of large-scale industrial production and use of chlorinated products. One recent study, for example, assayed dioxins in lipid from beef, pork and chicken from the turn of the 20th century through to the mid 1980s.13 The TEQ pg per gram of lipid in assayed food samples has risen from ~0.34 in 1908 to peak at ~3.73 during the 1960s, and then fell back to early 20th century values of 0.5 – 0.55 pg/g lipid in 1983. As a measure of total dietary intake of PCDDs and PCDFs, expressed as I-TEQ pg per kilogram body weight per day, input from dietary sources has decreased over a 20 year period, from 4.5 to approximately 05 pg TEQ/kg/day for the years spanning 1976 and 1997. A summary of these reports are presented in Figure 4.4. The average New Zealand dietary intake by an adult during this interval of time (including dioxin-like PCBs) has been determined to be 0.33 pg TEQ/kg/day, one of the lowest values reported in food, as compared with overseas studies.14

97

Figure 4.4 Reported estimated daily exposure to PCDDs and PCDFs, expressed in TEQ pg/kg body weight, from the United States,12 Netherlands,15 the United Kingdom16 and Germany.17 Reproduced with permission.18

4.2.2 Skin There is very little information on dermal contact following non-specific contact amongst the general population. TCDD bound soil has a significantly lower bioavailability (more than an one order of magnitude) compared to skin applications of dioxins alone.19 One report has estimated that dermal contact with soil may contribute less that 0.2 % to the total dioxin body burden.20 In contrast, despite poorer absorption of particulate bound dioxins, when compared to the pulmonary or intestinal route (see section 4.3.1), dermal exposure may make a contribution to the burden following workplace exposure to high levels of chlorinated aromatic compounds that may not have been bound to organic soil or other particulate matter. One recent case report has described a ceramic hobbyist working with clays containing elevated levels of dioxin congeners for over 30 years; she had substantially elevated serum levels, with congener profiles unique only to the clay, and route of entry was most likely chronic dermal exposure.21

4.2.3 Eye To the knowledge of the author, there have been no reports of dioxon exposure to the eye, or of the eye being a route of exposure leading to an increasing burden of PCDDs and PCDFs.

4.2.4 Inhalation Inhalation of PCDDs and PCDFs from sources of ambient air amongst the general population remains very low. The percentages of total dioxin derived from this route have been reported to be 3%,22 and 1.6%.20 Concentrations from cigarette smoke are minimal.119 Inhalation of dioxins, however, may not necessarily be a recent phenomena. PCDDs and PCDFs assayed from plant material dating back to the turn of the 20th century have shown distinct changes in chlorine saturation profiles of the dioxin aromatic rings, that corresponded with the phasing out of household wood and

98

coal fires; another profile emerged following reduced emissions of dixoins from other sources such as industrial discharges.23 However, over the complete time interval, the TEQ profile did still showed an overall decline. Another study from the same institute has demonstrated TCDD in all herbage samples back to 1861-65,24 suggesting this toxic congener has been present in ambient air, predating the advent of mass industrialisation in the 1930s.

4.3 Toxicokinetics

4.3.1 Absorption Due to their lipophilic nature, the various congeners are readily absorbed across the gastrointestinal tract, skin and lungs. The bioavailability of the congeners will be dependent on the degree of chlorine saturation and the composition of the delivery vehicle. For example, the bioavailability of TCDD in rats approaches 70-87% following dissolution in plant oil (Rose et al., 1976) is reduced to 50-60% in dietary ingestions,25 and is lower still when presented in a soil matrix.26 In one human study, the bioavailability of TCDD was determined to be >86%.27 Absorption into systemic circulation following ingestion occurs predominantly via the lymphatic route where transport occurs by both chylomicrons28 and lipoproteins.29 A study of dermal application to young male rats demonstrated that approximately 41 percent of the administered dose was absorbed into the systemic circulation following 120 hours exposure.30 Such absorption is age- and dose-related with younger rats displaying higher dermal absorption compared with older animals.31 Research has shown that administration of contaminated soil by intratracheal installation to female rats results in 100% bioavailability of TCDD.32 Indeed, the smaller particles of soil (those having a small enough aerodynamic diameter to achieve delivery to the alveoli where absorption occurs) had a 33-fold enrichment of the TCDD congener when compared with the total unfractionated soil. Similar bioavailabilites have been reported in another study with hamsters where the absorbed percentages of the administered doses for oral, pulmonary and dermal delivery were 95%, 88% and 40% respectively.33

4.3.2 Distribution Following absorption into the body, research on animals suggests dioxins are predominantly distributed to the liver, and adipose tissue,33, 34 though the ratio of distribution may vary depending on the congener.35-37 Overall, profiles of distribution, however, remain consistent throughout various animal models. To the knowledge of the author, the only data on human tissue distribution arose from post-mortem analysis of a woman living in Zona A at Seveso (Table 4.3) who subsequently died of pancreatic cancer,38 though this was not related to the exposure. She experienced no signs and symptoms following her contact with this agent, though two of her children suffered chloracne.

99

Table 4.3 Post-mortem TCDD concentrations (ng/g) in various tissues from a 55 year old woman who died of pancreatic cancer 7 months after exposure at Seveso

Blood Brain Lung Liver Pancreas Kidney Fat 0.006 0.06 0.06 0.15 1.04 0.04 1.84

Following tissue distribution, research has indicated that, for each TCDD congener, dioxin concentration in adipose closely parallel serum lipid values over a wide range of concentrations in mammals including humans.39, 40

4.3.3 Metabolism Due to their high degree of lipophilicity (see Table 4.1), both PCDDs and PCDFs will readily partition into host adipose tissue where they can persist for years41 and where they will remain predominantly unmetabolised. Evidence in hamsters has shown that following administration of TCDD, in excess of 97% of the resulting body burden was the unmetabolised parent chemical.42 The toxicity of TCDD is associated with the parent compound alone.43, 44 Thus metabolism is a pathway of detoxification in the host. Metabolism is slow, but once it has occurred elimination is rapid. Metabolism of the various congeners occurs in the liver. Phase I metabolism, catalysed by CYP 1A1 and CYP 1A2, is essentially an oxidative process, which is followed by Phase II conjugation of the oxidative products leading to water-soluble products that can readily be excreted in the bile or urine.45 The most common conjugated products are glucuronides and sulfates.

4.3.4 Elimination As discussed above, unmetabolised TCDD (and other toxic congeners) will remain in adipose and hepatic tissue for many years. Elimination half-lives are determined by rates of metabolism to low toxicity water-soluble products. The most potent of the dioxin-like compounds, 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD), has an estimated half-life in humans of 7.0- 11.3 years.46, 47 Half-lives can vary considerably, depending on the measured congener, concentration within the body and variabilities of metabolism amongst different human populations; a summary of calculated half-lives for the 17 toxic congeners from two studies are presented in Table 4.4. Despite such variation between the two investigations, it is clear the various dioxin congeners can persist in the body for many years. Elimination is also dependent on the age of the subject48 and their total fat content46; as dioxins distribute predominantly in adipose tissue, as the percentage of fat increases over the lifetime of a subject, the half-life of these compounds will also increase.49, 50 In contrast to humans, however, half-lives of the various dioxin congeners in other mammalian species are rapid. Absorbed TCDD following oral gavage was eliminated with a half-life of 20 days in the mouse.51 Guinea pigs, the most sensitive of the mammalian species, demonstrated a TCDD elimination half-life of 93 days.36

100

Table 4.4 Dose weighted estimated first-order elimination half-lives (years) for the 17 toxic dioxin congeners, as reported by Flesch-Janys and colleagues (1996)52 and Van der Molen and colleagues (2000)53

Congener Flesch-Janys et al (1996)

Van der Molen et al. (2000)

Dioxins 2,3,7,8-Tetra-CDD 7.2 7.6 1,2,3,7,8-Penta-CDD 15.7 11.1 1,2,3,4,7,8-Hexa-CDD 8.4 11.3 1,2,3,6,7,8-Hexa-CDD 13.1 11.5 1,2,3,7,8,9-Hexa-CDD 4.9 6.6 1,2,3,4,6,7,8-Hepta-CDD 3.7 8.5 OCDD 6.7 5.5 Dibenzofurans 2,3,7,8-Tetra-CDF 7.2 1.4 1,2,3,7,8-Penta-CDF 15.7 2.8 2,3,4,7,8-Penta-CDF 19.6 9.7 1,2,3,4,7,8-Hexa-CDF 6.2 7.5 1,2,3,6,7,8-Hexa-CDF 6 23.2 1,2,3,7,8,9-Hexa-CDF 6 NA 2,3,4,6,7,8-Hexa-CDF 5.8 3.5 1,2,3,4,6,7,8-Hepta-CDF 3 4.8 1,2,3,4,7,8,9-Hepta-CDF 3.2 10 OCDF 6.7 0.7

Body Burden As the rates of elimination of the various dioxin congeners in humans are measured in years, and the population is continually exposed to low levels of dioxins, people will have a pool of dioxins predominantly complexed in their hepatic and adipose tissue. This mass within the body is defined as the body burden, which may represent the accumulation of many years exposure,50 predominantly from dietary intake and to a lesser extent from (non-worker) industrial exposure or as a result of exposure to airborne pollution from wood or coal fires. With this in mind, a constant diet of ~1.3 pg/kg/day TCDD, for example, for the duration of an adult’s life will finally achieve steady state levels approximating 20 pg/g lipid following 30 to 37.5 years of exposure (representing 4 to 5 TCDD half-lives).54 The body burden for the 17 dioxin congeners is most commonly defined as the mean I-TEQ pg per gram of blood lipid (there is a parallel relationship between blood lipid and adipose concentration - see section 4.3.2). Although this burden may vary between different populations, it has been steadily declining since the mid-1960s when worldwide release of dioxins had peaked (see sections 4.1.3, 4.2.1, 4.9.2). Dioxin body burdens in Germany, for example, have declined in 10 years from ~45 in 1987 to ~ 12 I-TEQ pg/g lipid in 1996.55 A comprehensive report of non-occupational exposure to dioxins in the New Zealand population for 1996-1997 published a mean burden of 12.8 TEQ pg/g lipid,56 which is similar to values reported in Germany over the same time interval. Levels in breast milk in New Zealand have also shown a considerable decline, haven fallen by two-thirds over a recent 10-year period.115 The New Zealand Ministry of Health is continuing to monitor levels. Worldwide decreases in body burden can also be seen in decreasing plasma lipid concentrations of TCDD,

101

which has steadily decreased by a factor of ten over the last thirty years (see Figure 4.5).

Figure 4.5 Mean lipid adjusted TCDD values from populations (ppt) sampled in Europe and North America from 1970 to 2000. Reproduced with permission.54 In most instances, any addition to the body burden occurs through the ingestion of food but may also result from exposure to air pollution sourced from coal or wood fires, though this will be minimal. Where food has been accidentally contaminated, there can be a substantial increase, but such additions to the accumulated burden would be well below doses necessary to cause chloracne, the predominantly recognised symptom following dioxin poisoning. In 1999, poultry and eggs in Belgium were contaminated with various dioxin and PCB congeners.57 Assuming a severe contamination of 1,000 pg TEQ /per g of fat in chicken or eggs, it would require 30-40 meals to double the body burden of dioxins.57 Such a value would still be two orders of magnitude less than the burden of Seveso residents in Zone A, some of whom experienced chloracne, and would be similar to average burdens in the 1980s55 or those who regularly ate contaminated sea food.58

4.4 Mechanism of Action within the Body The mechanism of action of dioxin is poorly understood. Most of the understanding is based on the induction of a cytochrome P-450 intracellular receptor, termed the aromatic hydrocarbon receptor (AhR). A wide range of halogenated aromatic compounds, including PCDDs, PCDFs and co-planar PCBs, are stereo-specific for this receptor.59 The most active compound in this group of chemicals is TCDD, due to its chlorine substitution on the four lateral positions (positions 2,3,7,8 – see Figure 4.1). Upon binding, the dioxin molecule induces (or suppresses) the transcription of a number of genes, including CYP 1A1.60 Molecular mechanisms occurring downstream of AhR, as a result of TCDD receptor binding, have included changes in cellular signal proteins, mobilisation of calcium, tumour suppressor proteins, growth factors, oncogenes and cell cycle proteins.61, 62 Thus, improper alterations of gene expression may represent the initial steps in a series of biochemical, cellular and tissue changes that could result in host toxicity.

102

Recent research conducted on random subjects in Seveso, for example, have demonstrated that 20 years following the accident, a persistent down-regulation of this receptor was observed.63 The impacts on their health have yet to be clarified. In contrast, however, body burdens in human populations with values in excess of 250 ng TEQ/kg bw (25 times greater than the normal population) have shown no evidence of increased enzyme induction consequent to their elevated dioxin levels; elevated enzyme activity only became apparent at or above burdens of 750 ng TEQ/kg bw.64


4.5.1 Signs and Symptoms Most of the investigations involving dioxins have focused on exposures to TCDD as a result of its contamination of herbicides (both in their manufacture and distribution), or as a direct consequence of its accidental release from chemical production units. Apart from chloracne, most symptoms related to exposure are subtle and require sophisticated statistical studies, typically decades after the event, to identify them. As such the following section will summarise together the acute and chronic non-cancer investigations (carcinogenicity is addressed in section 4.6).

4.5.1.1 Human

Chloracne and Hepatic Effects

The only consistent manifestation of toxicity following dioxin (predominantly TCDD) exposure is chloracne.65, 204 For TCDD exposure, it is regarded as the “hallmark” of intoxication. Chloracne is characterised by inclusion cysts, comedones and pustulates with prominent scarring of the skin,65 more often on the face, but this can occur on other parts of the body as well. Such symptoms may be preceded by erythematous and oedematous skin. One early report suggests the time interval between exposure and the appearance of symptoms can range from 1-2 weeks to a few months.66 Following exposure to dioxin, as a result of workplace or accidental environmental contamination, this acne-like condition has occurred both with or without other effects.67 Prominent amongst the other symptoms are evidences of mild but transient hepatic injury. Some Seveso residents who suffered chloracne, for example, experienced temporary hepatomegaly for several months following exposure, but without evidence of elevated hepatic enzymes.68 Enlargement was not reported following other exposures.69-72 Transient elevations of transaminases and urinary D-glucaric acid have also been reported.73, 74

Indices of Endocrine Function and Biochemistry

Significant associations with several lipid-related health indices amongst US army veterans serving in Vietnam have been reported75; however, such associations were not reported in other investigations. There has been no evidence of statistical changes to thyroid function. Investigations of herbicide production workers did show a positive association between TCDD and the reproductive hormones luteinizing hormone and follicle stimulating hormone and inversely related to total testosterone.76 This has not been confirmed in studies of other intoxicated populations. Elevated diabetes-related mortalities amongst adult females have been reported in Seveso.77, 78

103

Cardiovascular and Pulmonary Effects

German chemical workers exposed to contaminated TCDD had higher mortalities from all circulatory diseases.79 However, this is in contrast to other investigations where there were no significant increases reported.80-82 Amongst 1261 Ranch Hand personnel, circulatory diseases were not significant47, 83; similar results have been reported on Australian veterans.84 Amongst Seveso residents, a reported three-fold increase in mortality from chronic ischaemic heart disease amongst males in Zone A has been reported.77, 78 Temporary respiratory tract irritation and tracheobronchitis has been reported amongst herbicide production workers following exposure to TCDD contaminant. However, conflicting evidence or that of little significance has been reported in other studies.85-87 Male Seveso residents in Zone B did show a statistically significant increase in obstructive pulmonary disease.77, 78

4.6 Carcinogenicity

4.6.1 Human Studies The International Agency for Research on Cancer (IARC) reclassified TCDD (the most potent congener) as a group 1 carcinogen in 1997.121 They based this on limited evidence in humans, sufficient evidence in animals, and extensive mechanistic information (indicating TCDD acts by binding to the AhR receptor). They judged it might cause an increase in all cancers combined, rather than at any specific site. Evidence in long-latency human epidemiological studies was based on four papers that summarised studies on four highly exposed industrial cohorts and the Seveso cohort.122-125 These studies demonstrated that though the relative risk was not high (they defined the overall relative risk at 1.4), increasing exposure (determined in part by estimated blood levels at the time of exposure) led to increased risks of cancer, and TCDD at high doses may act as a multi-site carcinogen. This committee did note, however, that the general population is exposed to levels far lower than those experienced by workers assessed in these studies; the general population have estimated TCDD levels at or below 5 ppt (see figure 4.5) whereas the studied populations (that suggested a relative increased risk of cancers) had estimated serum levels of 418,122 402,123 286124 and 400 ppt125 at the time of exposure, as reported by IARC.121 The United Kingdom Committee of Cancer (COC) concluded, that on the basis of recent epidemiological data, that TCDD is a probable weak human carcinogen and that a threshold approach to risk assessment was likely to be appropriate.126 They also noted that any background levels of exposure in the general population were extremely small and any health effects could not be detectable by current epidemiological methods. The Institute of Medicine determined there is sufficient evidence of an association between exposure to the chemicals of interest and soft-tissue sarcoma (including heart) non-Hodgkin’s lymphoma, chronic lymphocytic leukemia (including hairy-cell leukemia and other chronic B-cell leukemias), and Hodgkin’s disease.204 The United States EPA has defined TCDD as a human carcinogen and the other congeners as likely human carcinogens.12 They concluded there was suggestive evidence from epidemiological studies, unequivocal evidence from animal studies and inferences drawn from mechanistic data. As such they used a non-threshold model for

104

extrapolating a risk-specific dose for cancer risk, based on human and animal data (see sections 4.6 and 4.10.4). In contrast to the conservative approach undertaken by the EPA, the Joint FAO/WHO Expert Committee on Food Additives127 summarised that “TCDD was not an initiator of carcinogenesis” and the European Commission Scientific Committee on Foods128 stated that based on available information, “TCDD was not a direct-acting genotoxic agent”. Both committees concluded that cancer was appropriately addressed as a threshold phenomenon. Furthermore, the Agency for Toxic Substances and Disease Registry129 concluded there was insufficient data available to conduct low-dose extrapolation to define a dose range for cancer risk.130

4.6.2 Animal Studies Risk assessments for human exposure to TCDD and other dioxins is predominantly based on an animal study that consisted of a lifetime feeding study of female (n=50) and male (n = 50) rats.113 Sprague-Dawley rats were maintained on diets supplying 0, 10, and 100 ng/kg/day for two years to evaluate chronic toxicity and the potential for carcinogenicity. There was a significant incidence of hepatocellular carcinomas in rats fed 100 ng/kg/day whereas those fed at the lower dose of 10 ng/kg/day had increased incidents of hyperplastic nodules. Other carcinomas observed in the 100 ng/kg/day group included squamous cell carcinoma of the tongue and nasal turbinates/hard palate, and keratinizing squamous cell carcinoma of the lung. Increases in preneoplastic or neoplastic lesions were not found in the 1 ng/kg/day dose group. Later re-evaluations of liver slides from this study concluded incidents of hepatic tumours were consistently detected at 10 ng/kg/day.131-133 In addition, Squire and colleagues (1980) identified squamous cell carcinoma in 2 rats of the 1 ng/kg/day dose group suggesting 1 ng/kg/day may not represent a NOAEL. Using data from Kociba’s study, the US EPA fitted a non-threshold model to extrapolate a risk-specific dose for cancer risk (see section 4.10.4).

4.7 Mutagenicity Most research on mutagenicity has been conducted on the most toxic dioxin TCDD. Evidence has shown TCDD is not mutagenic in the Salmonella/Ames test with or without an activating system134, 135 nor does it cause DNA adducts in rodent tissues.136, 137 Rather than directly causing cancer by inducing mutation, it has been proposed that TCDD may damage DNA indirectly by inducing the CYP1A1- and CYP1A2-dependent drug metabolizing enzymes, possibly leading to increased formation of DNA-damaging metabolites (e.g. 138).


4.8.1 Human Studies There have been a number of studies that have assessed the possible association between TCDD and teratogenic effects in humans. There was no significant increase in spontaneous abortions amongst paternal military service personnel in Vietnam,139 nor amongst those involved in Operation Ranch Hand (nor for stillbirths).140 Similar results were reported with wives’ of 2,4,5-T applicators in New Zealand.141 Between April 1977 to December 1984 more females than males were been born to residents in Zone A in Seveso following high level TCDD exposure,142 but from 1985 to 1994 the

105

sex ratio was normal. Similar changes in sex ratio have been reported in Ufa, Russia following TCDD exposure in contaminated 2,4,5-T.143 Another study assessing workers manufacturing 2,4, 5-T, and who demonstrated eleveated levels of TCDD, did not show any change in the birth sex ratio.144 There was a rise in total birth defects following TCDD exposure in Seveso, though the incidence of any particular defect was not significant.102 The authors themselves suggested the defects may not be related to the TCDD exposure, as birth defects had been previously under-reported in Italy, and defects in Seveso were within the range reported by other Western countries. There were no significant increases over the next six successive years following exposure.104 However, the power of this study to determine significance was limited due to the small number of observed defects. Over the years 1973 to 1976, there was no significant increase in birth defects in residents in Northland (New Zealand), following 2,4,5-T exposure though there was an increased incidence of clubfoot.145 No significance increases in birth defects were found in mothers exposed to 2,4,5-T aerial spraying in a cohort study in Missouri.146 Two major studies examined the risk of Vietnam veterans having children with defects determined the risk of birth defects was not significantly increased though there was a slightly higher proportion of children with some defects.139, 147 However, there was no documentation of TCDD exposure. In a study where veterans did report self-exposure, there was no significant increase in birth defects in their children.148

4.8.2 Animal Studies Following oral exposure in rodents and monkeys, at doses ranging from 0.00012 to 25 µg/kg/day of TCDD, effects have been observed to cause a range of malformations to offspring animals including structural malformations-cleft palate and kidney anomalies (e.g. 149), functional alterations-damage to the immune system (e.g. 150) and impaired development of the reproductive system (e.g. 151), decreased growth and fetal/newborn mortality (e.g. 152). Of note was the LOAEL following 1.2 x 10-4 µg/kg/day chronic exposure, resulting in offspring neurobehavioral effects.153 This study was used by the ATSDR to derive a chronic oral MRL of 1×10-6 µg/kg/day.



Water

Selected freshwater streams and lakes in New Zealand were shown to contain between 1.2 and 5.4 TEQ pg/l; TCDD levels were below the detectable limits of the assay, and when 50% of the limits of detection were used, it accounted for 6-29% of the total TEQs. The highest value was determined downstream from a sawmill.154 Comparable studies in some other freshwater investigations have shown similar concentrations between 1.3 to 8.9 TEQ pg/l in the Xijiang River in China155 and 0.09 to 2.91 TEQ pg/l in the Houston Ship channel in Texas,156 though other report describe concentrations in both fresh and sea water at levels at 33,157 41158 and 73 TEQ pg/l.158

106

Air

Recent monitoring of ambient air across New Zealand reported mean concentrations that ranged from 1.4 (Nelson Lakes) to 317 fg TEQ/m3 (industrial region in Auckland South).120 Mean concentrations in selected urban areas were elevated, compared to rural areas, and ranged from 28 to 77 fg TEQ/m3; Christchurch had the highest levels at 77 fg TEQ/m3. Values were lower than comparable regions in other countries. In Sweden, for example, values in rural regions were 3.8 to 55 fg TEQ/m3,159 whereas those in extreme remote coastal regions of Sweden were comparable to New Zealand levels at 2.6 to 4.4 fg TEQ/m3.160 Levels in urban areas of the industrialised countries were also in excess of New Zealand counterpart regions; urban regions of Flanders (Belgium) had values ranging from 17 to 194 fg TEQ/m3,161 whereas urban areas in Japan have recorded values in the summer months that range from 403 to 1310 fg TEQ/m3.162

Soil

Soil concentrations at Paritutu, regarded as the most heavily contaminated region in New Zealand for dioxins, particularly TCDD, were assessed at 2.6-79 ng TEQ per kg soil163; these values were dominated by TCDD. Overall dioxin concentrations across New Zealand ranged from 0.17 - 1.99 ng I-TEQ/kg dry weight in forest and grassland soils, 0.17 -0.90 ng I-TEQ/kg in agricultural soils and 0.26 - 6.67 ng I-TEQ/kg in urban soils.164 Values in New Zealand are lower than studies reported from other countries. Congener concentrations in German165 and Swiss forests,166 for example were 5.4-112 and 0.5-26.1 ng TEQ/kg soil respectively whereas typical agricultural soils in Sweden167 and the United States168 were determined to be 9-49 and 0.08-22.9 ng TEQ/kg soil respectively. Wide variations in soil surrounding industrial sites and urban areas have also been reported; these include industrial and urban areas in Australia (1.6-14.4 ng TEQ/kg) and industrial regions in Brazil.(15-654 ng TEQ/kg).169, 170

4.9.2 Environmental Fate Environmental emissions of dioxins, such as from incineration and combustion sources, have led to their widespread distribution in the environment. As such, they can be found at low levels in rural soils and in sediments of unpolluted water bodies.171 Following release and transport in the atmosphere, dioxins will readily partition onto particulate matter and, to a lesser extent, to the vapour phase.172 Airborne dioxin conegeners that are partitioned in the vapour phase will undergo photo-degradation; in contrast, the more chlorinated congeners will readily bind to particulate matter, and are removed from the atmosphere predominantly by precipitation.173 Upon contact with water, dioxins may volatilise back to the atmosphere, bind to particulates and sediments or become bioaccumulated by the aquatic biota.174, 175 As dioxins have very low Henry’s law constants (see Table 4.1), volatilization from water is very slow and negligible. Due to their poor water solubility (see Table 4.1) dioxins will readily adsorb to sediments, dissolved organic carbon and to soil.176, 177 Upon entry into groundwater and other waterways, dioxins, due to their low water solubility, will readily enrich sediments.178 Historically, dioxin congeners will remain buried in sediments and are not considered to be bioavailable to the surrounding ecosystem. However, they can be mobilized during floods or due to other activities

107

that may disturb the sediments, leading to their re-entry into the surrounding environment.179 Binding to soil and any subsequent mobility through soil will depend on the degree of dioxin chlorine saturation. Homologues with increasing saturation of chlorine atoms, such as TCDD, possess stronger hydrophobicity and therefore have poor to almost non-existent mobility in soil. Consequently, these homologues will persist in soil for many decades.180 Nevertheless, some movement and leaching from soil can occur, though this is probably dependent on the presence of dissolved organic matter.181 It is thought the mobility of the congeners is achieved by adsorption to organic colloidal particles enabling them to reach-ground water and other water bodies.182-184 One of the concerning features of TCDD and other dioxin congeners is their ability to persist within the environment, bioaccumulate and biomagnify along the food chain.185 Consequently, foodstuffs of animal origin typically contain higher levels of dioxins than those sourced from a vegetable material.128 Ball clay sourced from a Mississippi geological formation, for example, contained substantially elevated levels of congeners including TCDD, that were most likely laid down in ancient sediment beds approximately 40 to 45 million years before the present.186 Processed clay from this source has been used as an anticaking agent for soybean meal destined for chicken and catfish consumption. Both feed187 and animals6 were later shown to have high levels of a distinct range of congeners, which included TCDD, unique to this clay. Such animals were later used for human consumption, which, over time, could lead to an increase in the host’s body burden, greater than other members of the population not consuming this food.

4.9.3 Ecological Degradation The rates of ecological degradation of dioxins will depend on the degrees of chlorine saturation and the environment. Based on limited data, rates of elimination have been determined for air water and soil188; predicted values were reported to be 0.02-1.1 years, 0.4-22 years and 17-273 years respectively. Homologues with increasing saturation of chlorine atoms, for example, possess the longer half-lives in soil.177 Dioxins in the vapour phase will undergo photolysis, predominantly by reaction with OH radicals173, 189; higher chlorinated dioxins bound to airborne particulate matter are poorly photo-degraded.190 In aqueous solution, photolysis is also the dominant route of degradation,191 though it is slower. Rates will vary according to the time of season.192 Degradation by microorganisms in water is low.193 In contrast, degradation in soil occurs predominantly by anaerobic and aerobic bacteria194 as well as fungi such as white rot fungus.195 Photolysis may occur, but is only effectively at soil surfaces.196


4.10.1 Doses Necessary to Cause Toxicity There have been a wide range of animal studies conducted to assess the toxicity of TCDD. As this report examines the risk of human toxicity following exposure to the general population, the summary of animal toxicity will be kept to experiments that were used to determine what is defined as an tolerable daily intake of dioxins for a life-time exposure that will have no health consequences.

108

The WHO88 has defined 1-4 pg TEQ body weight per day as an tolerable daily intake of dioxins. This value has been based on developmental effects in a range of animal species. They chose several experiments that demonstrated doses-levels that caused the most sensitive toxic and biochemical endpoints. A summary of these experiments is presented in table 4.5, summarising the lowest daily doses (LOAEL) that have resulted in observable adverse effects in the studied species. Doses administered to the animals necessary to achieve this increment of body burden were 64 ng/kg acute (rat),89 100 ng/kg acute (rat),90 200 ng/kg acute (rat),91 0.160 ng/kg/day chronic (monkey)92 and 0.160 ng/kg/day chronic (monkey).93 The table summarises the observed developmental and reproductive effects in rats and monkeys in association with the necessary increment to background body burdens in the experimental animals. These body burdens can readily be transformed into estimated daily human intakes that, on a chronic basis, would be expected to lead to similar body burdens in humans (see table 4.6, section 4.10.4 for further details) and that may result in some adverse effect upon the host. Table 4.5 Summary of LOAEL responses, animal body burdens, and the calculated estimated daily intakes (EDI) of TCDD in humans88

Reference Response (LOAEL) Body Burden (ng/kg bw

Extrapolated Human EDI (pg/kg bw/day)

91 Decreased sperm count in offspring (rat)

28 14

90, 94 Immune suppression in offspring (rat)

50 25

89 Increased genital malformations in offspring (rat)

73 37

92 Neurobehavioural (object learning) effects in offspring (monkey)

42 21

93 Endometriosis (monkey)

69 35

In 2002, the New Zealand Organochlorines Technical Advisory Committee adopted an Interim Maximum Monthly Intake (IMMI) of 30 pg TEQ per kg body weight per monthly intake.203 This is an interim value because further research on dioxins is underway and because the margin between current exposure even in NZ (which are low by international standards) and intakes that cause toxic effects in animals are undesirably small (see table 4.6). It is exposed as a monthly figure because dioxin stays in the body for a very long time. The lower end of the WHO TDI range (1 pg TEQ kg bw/ day)88 was used in deriving this value.

109

4.10.2 Summary of Case Reports Following Toxicological Exposure

Seveso

On the afternoon of July 10, 1976, an explosion occurred in the ICMESA (a subsidiary of Hoffman-La Roche) plant in Meda, about 25 km north of Milan (the Lombardia region of Italy) sending a plume of chemicals that was deposited south-southeast of the plant over an area of about 2.8 km2 (700 acres) that included parts of the towns of Seveso, Meda, Cesano, Moderno, and Desi. This cloud consisted of 2,4,5-trichlorophenate (the manufactured pesticide), sodium hydroxide, ethylene glycol, and approximately 30 kg of TCDD,95 which is a by-product of 2,4,5-trichlorophenol synthesis. Based on TCDD grass and soil levels and the subsequent death of domestic animals and plants, the contaminated area was divided into three zones (A, B and R). The surrounding area that was not contaminated was designated as a control zone (Zone non-ABR). All residents in Zone A were evacuated, but 511 (of 736) were allowed to re-enter areas of this zone after the least contaminated regions were decontaminated. Soil levels of TCDD in the exposed regions of Seveso were assessed in the three zones and were calculated to range from <0.75 to ~20 x 103 µg/m2 in Zone A, <50 µg/m2 in Zone B and <5 µg/m2 in Zone R.96 Subsequent investigations have shown these values were most likely an underestimation.97 Lesions and symptoms were the first symptoms to develop and these were later attributed to sodium hydroxide burns.98 By September, many cases of chloracne ranging from mild (type 1) to severe (type 4) were diagnosed; those identified with type 4 were residents in the heavily contaminated regions of zone A. Children from zone A who presented with chloracne had mean elevated TCDD blood lipid levels of 16.6 ng/mL99; estimated body burdens in these children, based on this investigation, have been estimated to be approximately 2,500 ng/kg, which is similar to the LD50 for guinea pigs100 (see section 4.10.3.1). Following the accident, health outcomes were investigated, including tests for liver function and lipid metabolism,73, 74 chloracne,101 miscarriages,102 cytogenetic abnormalities,103, 104 and neurological105 and immunological abnormalities.106, 107 A government appointed committee concluded that the only definable health effect was chloracne, and predominantly in children.108 Long term cohort studies following this event have shown some statistical increases in morbidity and mortality. In the first years after the incident there were elevated mortalities associated with cardiovascular disease and a suggested increase in chronic lung diseases in males and diabetes in females.77, 78 There have been however, consistent increases in mortality due to non-Hodgkin's lymphoma and myeloid leukemia109, 110 suggesting high levels of TCDD may pose a carcinogenic risk to humans.


4.10.3.1 LD50 values Reported acute oral LD50 values for TCDD vary considerably, depending on the studied species. For example, guinea pigs are the most sensitive at 0.6 to 2.1 µg/kg; 22 - 45 µg/kg in rats and 115 µg/kg in rabbits.100 In hamsters, acute oral values are 5500 µg/kg.111 Values in mice range from 100 to an excess of 3000 µg/kg.112

110

4.10.4 Guidance Values Tolerable daily intake (non-cancer related) The WHO has defined 1-4 pg TEQ per kg body weight per day as the tolerable daily intake (TDI) for lifetime exposure of dioxins that would have no health consequences.88 Such a value has been derived from animal studies assessing the LOAEL effect that caused the most sensitive toxic and biochemical endpoints in a range of animal species (see section 4.10.1 for more details). From these experiments, doses that were considered adverse occurred at body burdens in the range of 10 – 73 ng/kg. The WHO committee considered this range of studies (and their projected burdens) critical for the assessment of low dose effects of PCDDs and PCDFs. These burdens, derived from studies on monkeys and rats, can (according to the WHO committee) be readily transformed into estimated daily intakes. Chronic ingestions at these levels would be expected to lead to adverse effects. The calculation for a daily intake is estimated with the following equation:

Intake (ng/kg/day) = Body Burden (ng/kg) x (ln(2)/half-life)/f Where f is the fraction of dose absorbed and is assumed to be 50% for absorption from food for humans, and an estimated half-life for TCDD of 7.5 years is assumed. It also assumes elimination is linear (i.e. first-order kinetics), and the dioxins are distributed into one compartment alone (adipose tissue). The animal body burdens and related human estimated daily intakes (EDI) are presented in Table 4.5. The TDI is based on these EDI values. The WHO committee concluded that the uncertainty factor associated with species variability is covered in the range of these body burdens and only the uncertainty factor between animal and human need be considered; hence the EDI values were divided by 10 (human to animal variability) to yield the TDI which ranges from 1-4 pg TEQ per kg body weight per day. In addition to representing a tolerable daily intake for life-time exposure, the WHO committee concluded that an occasional short-term increase above this value would have no health consequences provided the average intake over a long time interval was not exceeded. They did stress that the upper value of 4 pg TEQ per kg body weight per day not be considered a maximal value, but rather the ultimate goal is to reduce human intake levels to below 1 TEQ per kg body weight per day. A summary of other recent evaluations, including the WHO TDI, is presented in table 4.6. Not surprising, in all assessments, tolerable daily intakes are within an average value set by the WHO.

111

Table 4.6 Summaries of international assessments of tolerable daily intakes of dioxins

Organisation Key end-point TDI World Health Organisation

Developmental effects 1-4 pg TEQ / kg bw per day

European Commission, Scientific Committee on Food

Developmental effects 14 pg TEQ / kg bw per week

Joint FAO/WHO Expert Committee on Food Additives

Developmental effects 70 pg TEQ / kg bw per month

Committee of Toxicity Developmental effects 2 pg TEQ / kg bw per day NZ Ministry of Health Developmental effects 30 pg TEQ / kg bw per month

(Interim Maximum Monthly Intake)

Other Guideline Values

The EU Scientific Committee on Food (SCF) set a provisional tolerable weekly intake (pTWI) for dioxins at 14 pg TEQ/kg body weight (bw)/week, analogous to 2 pg TEQ/kg bw/day.118 The maximum contaminant levels in water has been set at 0.00003 parts per billion (ppb) because EPA believes, given present technology and resources, this is the lowest level to which water systems can reasonably be required to remove this contaminant should it occur in drinking water.

Acceptable Risk Level (cancer-related)

Acceptable dietary intakes following exposure to TCDD (or other dioxin congeners) exposure, in association with risks of non-carcinogenic diseases, have been established from LOAEL animal studies. In contrast, investigations to establish an acceptable risk level for carcinogenic substances have been more complicated. An acceptable risk level is often used to define the acceptable risk associated with exposure to non-threshold contaminants. Rather than factoring in a margin of safety below a NOAEL or LOAEL, carcinogenic risk to human health has been defined by mathematically extrapolating a value from an appropriate dose response curve; the EPA have undertaken such an approach to determine a suitable ‘risk specific dose’ (RsD) for a one in a million lifetime cancer risk. The EPA regards such an approach is necessary as there is yet no known threshold for TCDD-related cancers. Linear modelling of female rat liver tumour data, over a range of high but decreasing doses of TCDD,113 has provided an estimate for an oral RsD of ~ 0.01 pg TCDD kg per day necessary for a one in a million lifetime cancer risk.114 This RsD corresponds to a unit risk estimate of 1 x 10-4 pg per kg per day. The most recent investigations by the EPA, based on the upper confidence interval of the mathematically modelled slope, have proposed an upper bound unit cancer risk of 1.4 x 10-3 pg per kg per day, is necessary for a one in a million lifetime cancer risk.12 From this conservative unit cancer risk value (1.4 x 10-3 pg/kg/day), the risk level for the general population over a lifetime of 70 years can be determined from their respective average daily dose, using the following formula:

112

RL = LADD x ql

Where RL represents the risk level over a life-time of 70 years, LADD defines the lifetime average daily dose (pg/ kg/ day) and ql is the unit risk cancer estimate (1.4 x 10-3 pg/kg/day). Thus the present New Zealand intake, averaged over the last 30 years,115 is ~1.4 pg TEQ per kg per day. This represents an approximately one additional cancer per 1000 people over a lifetime of 70 years. Such a risk is 100 times higher than the value of 1:100,000, the value that has been adopted in a number of government publications in New Zealand, such as the Drinking-water Standards New Zealand 2005 and the joint Ministry for the Environment and Ministry of Health’s “Health and environmental guidelines for selected timber treatment chemicals (1997).115 However, given the controversy associated with the EPA’s risk assessment of TCDD and related dioxins, they published a draft reassessment in 2003 which was provided for review by the National Academy of Sciences.116. A comprehensive review of the EPA’s draft assessment included criticisms that the agency had understated the uncertainties associated with their risk assessment and they may have overstated the cancer risk.117

4.11 A Summary of Biological Assays that are Available to Test for Evidence of These Chemicals in Biological Samples (e.g. blood, urine) TCDD and other dioxin congeners in biological samples are predominantly determined by gas chromatography mass spectrometry (high resolution).197 They can be assayed in serum,198 plasma,199 breast milk200 and adipose tissue201; levels of detection in these biological matrices have been determined, in some instances, to levels as low as sub parts per trillion.40, 198, 202 In recent years, blood lipid assessment has replaced adipose level determination as adipose samples required surgery for measurement. When based on a lipid-weight basis, there is a parallel correlation between adipose lipid and serum lipid content over a wide range of concentrations.40 Dioxins have also been determined from other media such as foodstuffs of animal origin (meat, eggs, fish and milk), ambient air, soil, sediments and water.129 Nevertheless, despite advances in technology, methodology and levels of detection, dioxin determinations in biological samples still require a series of complicated steps in order to extract, validate and finally assay the various congeners.197 Therefore, analysis is both slow and expensive. For dioxin blood lipid determination, approximately 30-50 mL of serum (60-80 mL blood) must be collected from each subject, plus an additional 5mL for lipid testing. Laboratory testing must comply with EPA method 1613 for dioxin and furan analysis by High Resolution Gas Chromatography Mass Spectrometry using isotopic dilution. Such a lab must also have ISO/IEC 17025 registration. This ensures the laboratory has been tested on managerial and technical competency on achieving accuracy, reproducibility and traceability of their test measurements and calibrations. Analysis must be undertaken for all 17 toxic dioxins and furans as well as the 12 PCB dioxin-like congeners. Final results must be reported as pg TEQ / g blood lipid. In New Zealand, the only laboratory that meets these criteria, and routinely assays dioxins is AsureQuality Limited in Lower Hutt.

113

4.12 Summary The term dioxins describes 210 different dibenzo-p-dioxins and dibenzofurans of which only 17 are regarded as toxic; TCDD is regarded as the most toxic. The overall toxicity for these congeners, relative to TCDD, are summated and expresssed as a toxic equivalent (TEQ). Dioxins are overwhelmingly a product of industrialisation though they can form naturally. Most forms of exposure are a consequence of dietary ingestion though elevated exposure can occur following dermal or respiratory uptake, due to accidental industrial release. Daily exposure to dioxins, predominantly by dietary intake, has increased since the advent of mass industrialisation until the 1970s, from which time it has been steadily declining. Dermal contact and inhalation can both contribute to the burden, though contributions are only significant following workplace exposure or industrial accidents. The bioavailability of dioxin congeners is high following oral absorption with systemic circulation occurring via the lymphatic route; this results in distribution, predominantly in the adipose and hepatic tissue. Serum lipid concentrations closely parallel values in adipose tissue. Dioxins persist for many years in the adipose tissue; the elimination half-life of TCDD, for example, is approximately 7.4 years, though this will vary according to the congener. In animals, however, rates of elimination are measured in weeks. The pool of dioxins that can be calculated from serum dioxin levels is defined as the body burden, and is termed as the mean TEQ ng per kg body weight. This reflects the lifetime accumulation (to achieve steady-state levels) of dioxins and is an indicator of toxicity. Over the last 30 years there has been a worldwide steady decrease (by a factor of ten) in body burden. The mechanism of action of dioxin is poorly understood, though it is understood it binds to an intracellular AhR receptor. This induces a number of genes, including CYP 1A1. Molecular mechanisms occurring downstream of AhR binding are thought to initiate a series of biochemical steps necessary for host toxicity, rather than direct mutagenicity (this has not been conclusively demonstrated in eukaryotic or prokaryotic organisms). The most consistent symptoms following exposure are chloracne and to a lesser extent mild but transient hepatic injury. Other symptoms are subtle and require sophisticated statistical studies, typically decades after the exposure, to identify them. More often, conclusions drawn from different cohort studies are in conflict. There is some evidence of subtle changes in endocrine and biochemistry function. Some studies also suggest there is an increased risk of cardiovascular and pulmonary diseases and diabetes. IARC concluded there is a relative risk, though not high, of increased rates of risks of cancer and TCDD at high doses may act as a multi-site carcinogen. Such a risk is in certain workers that have had high exposure, and not the general population. There is still disagreement with other international bodies such as the US EPA. They extrapolated risks of increased cancers from a range of studies; their draft document, that is still under review, considered that even amongst the general population, who currently have relatively low body burden of dioxins, the cancer risk is regarded as unacceptably high. In contrast, JEFCA, ECSCF and ATSDR have concluded risks to human health are presently adequately addressed by the threshold phenomenon, which suggests current dietary exposure is well within tolerable daily intakes. At substantially higher doses of TCDD, animal studies have reported a range of different cancers. Evidence of teratogenicity in humans has been both subtle and inconclusive, though in animal models malformations have been demonstrated though at substantially high doses of TCDD.

114

The various congeners of dioxin have been determined in environmental samples sourced from water soil and air. Concentrations in New Zealand are relatively lower than overseas investigations. Dioxins released into the atmosphere either bind to particulate matter and remain resistant to degradation or remain in the vapour phase, where photo-degradation can occurs. Water-borne dioxins will either be bioaccumulated by the aquatic biota or adsorb to sediments where they remain buried, unless disturbed by floods or similar environmental action. In soil they are poorly mobile and can remain for many decades or centuries, unless adsorption to organic colloidal particles occurs, thereby leading to ground water leaching. Predicted half-lives in air, water and soil are reported to be 0.02-1.1 years, 0.4-22 years and 17-273 years respectively. Analysis of TCDD and other congeners are predominantly carried out by gas chromatography mass spectrometry (high resolution). Analysis of serum lipid content clearly defines adipose tissue concentration, from which the human body burden can be determined.

4.13 References 1. Kutz FW, Barnes DG, Bottimore DP, Greim H, Bretthauer EW. The

International Toxicity Equivalency Factor (I-TEF) method of risk assessment for complex-mixtures of dioxins and related-compounds. Chemosphere. 1990;20:751-757.

2. Van den Berg M, Birnbaum L, Bosveld ATC, et al. Toxic equivalency factors (TEFs) for PCBs, PCDDs, PCDFs for humans and wildlife. Environmental Health Perspectives. 1998;106:775-792.

3. Dow Chemicals. The Trace Chemistries of Fire—A Source of and Routes for the Entry of Chlorinated Dioxins into the Environment. The Chlorinated Dioxin Task Force. Michigan Division,: Dow Chemicals, USA; 1978. In McKay G. Dioxin characterisation, formation and minimisation during municipal solid waste (MSW) incineration: review. Chemical Engineering Journal. 2002;86:343-368

4. McKay G. Dioxin characterisation, formation and minimisation during municipal solid waste (MSW) incineration: review. Chemical Engineering Journal. 2002;86:343-368.

5. Gribble GW. The natural production of chlorinated compounds. Environmental Science & Technology. 1994;28:A310-A319.

6. Ferrario J, Byrne C. The concentration and distribution of 2,3,7,8-dibenzo-p-dioxins/-furans in chickens. Chemosphere. 2000;40:221-224.

7. Holmstrand H, Gadomski D, Mandalakis M, et al. Origin of PCDDs in ball clay assessed with compound-specific chlorine isotope analysis and radiocarbon dating. Environmental Science & Technology. 2006;40:3730-3735.

8. Klimm C, Schramm KW, Henkelmann B, Martens D, Kettrup A. Formation of octa- and heptachlorodibenzo-p-dioxins during semi anaerobic digestion of sewage sludge. Chemosphere. 1998;37:2003-2011.

9. Hoekstra EJ, De Weerd H, De Leer EWB, Brinkman UAT. Natural formation of chlorinated phenols, dibenzo-p-dioxins, and dibenzofurans in soil of a Douglas fir forest. Environmental Science & Technology. 1999;33:2543-2549.

10. Ligon WV, Dorn SB, May RJ, Allison MJ. Chlorodibenzofuran and chlorodibenzo-para-dioxin levels in Chilean mummies dated to about 2800 years before the present. Environmental Science & Technology. 1989;23:1286-1290.

11. Gilpin K, Wagel DJ, Soch JG. Production, distribution, and fate of polychlorinated dibenzo-p-dioxins, dibenzofurans, and related organohalogens

115

in the environment. In: Schecter A, Gasiewicz TA, eds. Dioxins and Health. New York, NY: Wiley–Interscience; 2003:55-88.

12. EPA. (Environmental Protection Agency) Exposure and Human Health Reassessment of 2,3,7,8-Tetrachlorodibenzo-p-Dioxin (TCDD) and Related Compounds National Academy Sciences (NAS) Review Draft National Center for Environmental Assessment, Office of Research and Development, U.S. EPA.Review Draft. Available at <http://www.epa.gov/ncea/pdfs/dioxin/nas-review>(Last accessed on Oct 11 2008); 2003.

13. Winters D, Anderson S, Lorber M, Ferrario J, Byrne C. Trends in dioxin and PCB concentrations in meat samples from several decades of the 20th century. Organohalogen Compounds. 1998;38:75-77.

14. Buckland SJ, Scobie S, Heslop V. Organochlorines in New Zealand: Concentrations of PCDDs, PCDFs and PCBs in retail foods and an assessment of dietary intake for New Zealanders. Available at <http://www.mfe.govt.nz/publications/hazardous/diet-assessment-nov98.pdf> (Last accessed from 13 Oct 2008). Wellington: Ministry for the Environment; 1998.

15. Liem AK, Fürst P, Rappe C. Exposure of populations to dioxins and related compounds. Food Additives and Contaminants. 2000;17:241-259.

16. FSA. Food Standards Agency UK, Dioxins and PCBs in the UK diet: 1997 Total diet study samples. Information Sheet Number 4/00.; 2000.

17. Fürst P, Wilmers K. Decline of human PCDD/F intake via food between 1989 and 1996. Organohalogen Compounds. 1997;33:116-121.

18. Hays SM, Aylward LL. Dioxin risks in perspective: past, present, and future. Regulatory Toxicology and Pharmacology. 2003;37:202-217.

19. Roy TA, Hammerstrom K, Schaum J. Percutaneous absorption of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) from soil. Journal of Toxicology & Environmental Health Part A. 2008;71:1509-1515.

20. Theelen RMC. Modeling of human exposure to TCDD and I-TEQ in the Netherlands: background and occupational. In: Gallo MA, Scheuplein RJ, van der Heijden KA, eds. Banbury Report 35, Biological basis for risk assessment of dioxins and related compounds. Woodbury, NY: Cold Spring Harbor Laboratory Press; 1991:227-290.

21. Franzblau A, Hedgeman E, Chen Q, et al. Case report: human exposure to dioxins from clay. Environmental Health Perspectives. 2008;116:238-242.

22. Birmingham B, Gilman A, Grant D, et al. PCDD/PCDF Multi-media exposure analysis for the Canadian population - detailed exposure estimation. Chemosphere. 1989;19:637-642.

23. Hassanin A, Lee RG, Steinnes E, Jones KC. Reductions and changing patterns of ambient PCDD/Fs in the UK: Evidence and implications. Chemosphere. 2006;65:530-539.

24. Kjeller LO, Jones KC, Johnston AE, Rappe C. Evidence for a decline in atmospheric emissions of PCDD/Fs in the UK. Environmental Science & Technology. 1996;30:1398-1403.

25. EPA. (Environmental Protection Agency). Health Assessment Document for Polychlorinated Dibenzo-p-dioxin. EPA, Cincinnati, OH. PB86-122546; 1985.

26. Lucier GW, Rumbaugh RC, McCoy Z, Hass R, Harvan D, Albro P. Ingestion of soil contaminated with 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) alters hepatic enzyme activities in rats. Fundamental & Applied Toxicology. 1986;6:364-371.

116

27. Poiger H, Schlatter C. Pharmacokinetics of 2,3,7,8-TCDD in man. Chemosphere. 1986;15:1489-1494.

28. Lakshmanan MR, Campbell BS, Chirtel SJ, Ekarohita N, Ezekiel M. Studies on the mechanism of absorption and distribution of 2,3,7,8-tetrachlorodibenzo-p-dioxin in the rat. Journal of Pharmacology & Experimental Therapeutics. 1986;239:673-677.

29. Henderson LO, Patterson DG, Jr. Distribution of 2,3,7,8-tetrachlorodibenzo-p-dioxin in human whole blood and its association with, and extractability from, lipoproteins. Bulletin of Environmental Contamination & Toxicology. 1988;40:604-611.

30. Banks YB, Birnbaum LS. Absorption of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) after low dose dermal exposure. Toxicology & Applied Pharmacology. 1991;107:302-310.

31. Banks YB, Brewster DW, Birnbaum LS. Age-related changes in dermal absorption of 2,3,7, 8-tetrachlorodibenzo-p-dioxin and 2,3,4,7,8-pentachlorodibenzofuran. Fundamental & Applied Toxicology. 1990;15:163-173.

32. Nessel CS, Amoruso MA, Umbreit TH, Meeker RJ, Gallo MA. Pulmonary bioavailability and fine particle enrichment of 2,3,7,8-tetrachlorodibenzo-p-dioxin in respirable soil particles. Fundamental & Applied Toxicology. 1992;19:279-285.

33. Diliberto JJ, Jackson JA, Birnbaum LS. Comparison of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) disposition following pulmonary, oral, dermal and parenteral exposures to rats. Toxicology and Applied Pharmacology. 1996;138:158-168.

34. Rose JQ, Ramsey JC, Wentzler TH, Hummel RA, Gehring PJ. The fate of 2,3,7,8-tetrachlorodibenzo-p-dioxin following single and repeated oral doses to the rat. Toxicology & Applied Pharmacology. 1976;36:209-226.

35. Birnbaum LS. Distribution and excretion of 2,3,7,8-tetrachlorodibenzo-p-dioxin in congenic strains of mice which differ at the Ah locus. Drug Metabolism & Disposition. 1986;14:34-40.

36. Olson JR. Metabolism and disposition of 2,3,7,8-tetrachlorodibenzo-p-dioxin in guinea pigs. Toxicology & Applied Pharmacology. 1986;85:263-273.

37. Neal RA, Beatty PW, Gasiewicz TA. Studies of the mechanisms of toxicity of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD). Annals of the New York Academy of Sciences. 1979;320:204-213.

38. Montagna M, Fornari A, Facchetti S. Analysis of autopsy samples for the detection of 2,3,7,8-tetrachlorodibenzo-p-dioxin by high resolution GC-MS. In: Oliver JS, ed. Forensic Toxicology. London: Croom Helm; 1979:78-85.

39. Geyer HJ, Scheuntert I, Rapp K, et al. Correlation between acute toxicity of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) and total body fat content in mammals. Toxicology. 1990;65:97-107.

40. Patterson DG, Jr., Needham LL, Pirkle JL, et al. Correlation between serum and adipose tissue levels of 2,3,7,8-tetrachlorodibenzo-p-dioxin in 50 persons from Missouri. Archives of Environmental Contamination & Toxicology. 1988;17:139-143.

41. Schecter A, Quynh HT, Papke O, Tung KC, Constable JD. Agent Orange, dioxins, and other chemicals of concern in Vietnam: update 2006. Journal of Occupational & Environmental Medicine. 2006;48:408-413.

117

42. Olson JR, Gasiewicz TA, Neal RA. Tissue distribution, excretion, and metabolism of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) in the Golden Syrian hamster. Toxicology & Applied Pharmacology. 1980;56:78-85.

43. Weber H, Poiger H, Schlatter C. Acute oral toxicity of TCDD-metabolites in male guinea pigs. Toxicology Letters. 1982;14:117-122.

44. Mason G, Safe S. Synthesis, biologic and toxic effects of the major 2,3,7,8-tetrachlorodibenzo-p-dioxin metabolites in the rat. Toxicology. 1986;41:153-159.

45. Hu K, Bunce NJ. Metabolism of polychlorinated dibenzo-p-dioxins and related dioxin-like compounds. Journal of Toxicology & Environmental Health Part B: Critical Reviews. 1999;2:183-210.

46. Michalek JE, Tripathi RC. Pharmacokinetics of TCDD in veterans of Operation Ranch Hand: 15-year follow-up. Journal of Toxicology & Environmental Health Part A. 1999;57:369-378.

47. Wolfe WH, Michalek JE, Miner JC, et al. Determinants of TCDD half-life in veterans of operation ranch hand. Journal of Toxicology & Environmental Health. 1994;41:481-488.

48. Kerger BD, Leung HW, Scott P, et al. Age- and concentration-dependent elimination half-life of 2,3,7,8-tetrachlorodibenzo-p-dioxin in Seveso children. Environmental Health Perspectives. 2006;114:1596-1602.

49. Van der Molen GW, Kooijman BA, Slob W. A generic toxicokinetic model for persistent lipophilic compounds in humans: an application to TCDD. Fundamental & Applied Toxicology. 1996;31:83-94.

50. Pinsky PF, Lorber MN. A model to evaluate past exposure to 2,3,7,8-TCDD. Journal of Exposure Analysis and Environmental Epidemiology. 1998;8:187-206.

51. Koshakji RP, Harbison RD, Bush MT. Studies on the metabolic fate of [14C]2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) in the mouse. Toxicology & Applied Pharmacology. 1984;73:69-77.

52. Flesch-Janys D, Becher H, Gurn P, et al. Elimination of polychlorinated dibenzo-p-dioxins and dibenzofurans in occupationally exposed persons. Journal of Toxicology & Environmental Health. 1996;47:363-378.

53. Van der Molen GW, Kooijman BA, Wittsiepe J, Schrey P, Flesch-Janys D, Slob W. Estimation of dioxin and furan elimination rates with a pharmacokinetic model. Journal of Exposure Analysis and Environmental Epidemiology. 2000;10:579-585.

54. Aylward LL, Hays SM. Temporal trends in human TCDD body burden: decreases over three decades and implications for exposure levels. Journal of Exposure Analysis and Environmental Epidemiology. 2002;12:319-328.

55. Päpke O. PCDD/PCDF: human background data for Germany, a 10-year experience. Environmental Health Perspectives. 1998;106 Suppl 2:723-731.

56. Bates MN, Buckland SJ, Garrett N, et al. Persistent organochlorines in the serum of the non-occupationally exposed New Zealand population. Chemosphere. 2004;54:1431-1443.

57. Bernard A, Hermans C, Broeckaert F, De Poorter G, De Cock A, Houins G. Food contamination by PCBs and dioxins. Nature. 1999;401:231-232.

58. Asplund L, Svensson BG, Nilsson A, et al. Polychlorinated biphenyls, 1,1,1-trichloro-2,2-bis(p-chlorophenyl)ethane (p,p'-DDT) and 1,1-dichloro-2,2-bis(p-chlorophenyl)-ethylene (p,p'-DDE) in human plasma related to fish consumption. Archives of Environmental Health. 1994;49:477-486.

118

59. Safe SH. Comparative toxicology and mechanism of action of polychlorinated dibenzo-p-dioxins and dibenzofurans. Annual Review of Pharmacology & Toxicology. 1986;26:371-399.

60. Whitlock JP, Jr. Induction of cytochrome P4501A1. Annual Review of Pharmacology & Toxicology. 1999;39:103-125.

61. Carlson DB, Perdew GH. A dynamic role for the Ah receptor in cell signaling? Insights from a diverse group of Ah receptor interacting proteins. Journal of Biochemical & Molecular Toxicology. 2002;16:317-325.

62. Enan E, El-Sabeawy F, Scott M, Overstreet J, Lasley B. Alterations in the growth factor signal transduction pathways and modulators of the cell cycle in endocervical cells from macaques exposed to TCDD. Toxicology & Applied Pharmacology. 1998;151:283-293.

63. Landi MT, Bertazzi PA, Baccarelli A, et al. TCDD-mediated alterations in the AhR-dependent pathway in Seveso, Italy, 20 years after the accident. Carcinogenesis. 2003;24:673-680.

64. Connor KT, Aylward LL. Human response to dioxin: aryl hydrocarbon receptor (AhR) molecular structure, function, and dose-response data for enzyme induction indicate an impaired human AhR. Journal of Toxicology & Environmental Health Part B: Critical Reviews. 2006;9:147-71.

65. Panteleyev AA, Bickers DR. Dioxin-induced chloracne--reconstructing the cellular and molecular mechanisms of a classic environmental disease. Experimental Dermatology. 2006;15:705-730.

66. May G. Chloracne from the accidental production of tetrachlorodibenzodioxin. British Journal of Industrial Medicine. 1973;30:276-283.

67. Sweeney MH, Mocarelli P. Human health effects after exposure to 2,3,7,8-TCDD. Food Additives & Contaminants. 2000;17:303-316.

68. Reggiani G. Acute human exposure to TCDD in Seveso, Italy. Journal of Toxicology & Environmental Health. 1980;6:27-43.

69. Moses M, Lilis R, Crow KD, et al. Health status of workers with past exposure to 2,3,7,8-tetrachlorodibenzo-p-dioxin in the manufacture of 2,4,5-trichlorophenoxyacetic acid: comparison of findings with and without chloracne. American Journal of Industrial Medicine. 1984;5:161-182.

70. RH, Grubbs WD, Lustik MB, et al. Air Force Health Study, an epidemiologic investigation of health effects in Air Force personnel following exposure to herbicides. Serum dioxin analysis of 1987 examination results. National Technical Information Service # AD A-237-516 through AD A-237-524; 1991. Summarised in Wolfe WH, Michalek JE, Miner JC, et al. The Air-Force health study - an epidemiologic investigation of health-effects in Air-Force personnel following exposure to herbicides, serum dioxin analysis of 1987 examination results. Chemosphere. 1992;25:213-216.

71. Hoffman RE, Stehr-Green PA, Webb KB, et al. Health effects of long-term exposure to 2,3,7,8-tetrachlorodibenzo-p-dioxin. JAMA. 1986;255:2031-2038.

72. Calvert GM, Hornung RW, Sweeney MH, Fingerhut MA, Halperin WE. Hepatic and gastrointestinal effects in an occupational cohort exposed to 2,3,7,8-tetrachlorodibenzo-para-dioxin. JAMA. 1992;267:2209-2214.

73. Mocarelli P, Marocchi A, Brambilla P, Gerthoux P, Young DS, Mantel N. Clinical laboratory manifestations of exposure to dioxin in children. A six-year study of the effects of an environmental disaster near Seveso, Italy. JAMA. 1986;256:2687-2695.

119

74. Ideo G, Bellati G, Bellobuono A, Bissanti L. Urinary D-glucaric acid excretion in the Seveso area, polluted by tetrachloro-dibenzo-p-dioxin (TCDD): five years of experience. Environmental Health Perspectives. 1985;60:151-157.

75. Wolfe WH, Michalek JE, Miner JC, et al. The Air-Force health study - an epidemiologic investigation of health-effects in Air-Force personnel following exposure to herbicides, serum dioxin analysis of 1987 examination results. Chemosphere. 1992;25:213-216.

76. Egeland GM, Sweeney MH, Fingerhut MA, Wille KK, Schnorr TM, Halperin WE. Total serum testosterone and gonadotropins in workers exposed to dioxin. American Journal of Epidemiology. 1994;139:272-281.

77. Pesatori AC, Zocchetti C, Guercilena S, Consonni D, Turrini D, Bertazzi PA. Dioxin exposure and non-malignant health effects: a mortality study. Occupational & Environmental Medicine. 1998;55:126-131.

78. Consonni D, Pesatori AC, Zocchetti C, et al. Mortality in a population exposed to dioxin after the Seveso, Italy, accident in 1976: 25 years of follow-up. American Journal of Epidemiology. 2008;167:847-858.

79. Flesch-Janys D, Berger J, Gurn P, et al. Exposure to polychlorinated dioxins and furans (PCDD/F) and mortality in a cohort of workers from a herbicide-producing plant in Hamburg, Federal-Republic-of-Germany. American Journal of Epidemiology. 1995;142:1165-1175.

80. Bueno de Mesquita HB, Doornbos G, Van der Kuip DA, Kogevinas M, Winkelmann R. Occupational exposure to phenoxy herbicides and chlorophenols and cancer mortality in The Netherlands. American Journal of Industrial Medicine. 1993;23:289-300.

81. Collins JJ, Strauss ME, Levinskas GJ, Conner PR. The mortality experience of workers exposed to 2,3,7,8-tetrachlorodibenzo-p-dioxin in a trichlorophenol process accident. Epidemiology. 1993;4:7-13.

82. Coggon D, Pannett B, Winter P. Mortality and incidence of cancer at four factories making phenoxy herbicides. British Journal of Industrial Medicine. 1991;48:173-178.

83. Michalek JE, Wolfe WH, Miner JC. Health-status of Air-Force veterans occupationally exposed to herbicides in Vietnam .2. Mortality. Jama-Journal of the American Medical Association. 1990;264:1832-1836.

84. Fett MJ, Adena MA, Cobbin DM, Dunn M. Mortality among Australian conscripts of the Vietnam conflict era .1. Death from all causes. American Journal of Epidemiology. 1987;125:869-877.

85. Calvert GM, Sweeney MH, Morris JA, Fingerhut MA, Hornung RW, Halperin WE. Evaluation of chronic bronchitis, chronic obstructive pulmonary disease, and ventilatory function among workers exposed to 2,3,7,8-tetrachlorodibenzo-p-dioxin. American Review of Respiratory Disease. 1991;144:1302-1306.

86. Suskind RR, Hertzberg VS. Human health effects of 2,4,5-T and its toxic contaminants. JAMA. 1984;251:2372-2380.

87. Grubbs WD, Wolfe WH, Michalek JE, et al. Air Force Health Study, An epidemiologic investigation of health effects in Air Force personnel following exposure to herbicides. Report number AL-TR-920 107; 1995. In Sweeney MH, Mocarelli P. Human health effects after exposure to 2,3,7,8-TCDD. Food Additives & Contaminants. 2000;17:303-316

88. WHO. Assessment of the health risks of dioxins: re-evaluation of the Tolerable Daily Intake (TDI). Executive Summary. Available from <http://www.who.int/ipcs/publications/en/exe-sum-final.pdf>. (Last accessed on Oct 23 2008). Geneva, Switzerland: World Health Organization; 1998.

120

89. Gray LE, Wolf C, Mann P, Ostby JS. In utero exposure to low doses of 2,3,7,8-tetrachlorodibenzo-p-dioxin alters reproductive development of female Long Evans hooded rat offspring. Toxicology & Applied Pharmacology. 1997;146:237-244.

90. Gehrs BC, Riddle MM, Williams WC, Smialowicz RJ. Alterations in the developing immune system of the F344 rat after perinatal exposure to 2,3,7,8-tetrachlorodibenzo-p-dioxin: II. Effects on the pup and the adult. Toxicology. 1997;122:229-240.

91. Gray LE, Ostby JS, Kelce WR. A dose-response analysis of the reproductive effects of a single gestational dose of 2,3,7,8-tetrachlorodibenzo-p-dioxin in male Long Evans Hooded rat offspring. Toxicology & Applied Pharmacology. 1997;146:11-20.

92. Schantz SL, Bowman RE. Learning in monkeys exposed perinatally to 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD). Neurotoxicology & Teratology. 1989;11:13-19.

93. Rier SE, Martin DC, Bowman RE, Dmowski WP, Becker JL. Endometriosis in Rhesus-monkeys (Macaca-mulatta) following chronic exposure to 2,3,7,8-tetrachlorodibenzo-p-dioxin. Fundamental and Applied Toxicology. 1993;21:433-441.

94. Gehrs BC, Smialowicz RJ. Persistent suppression of delayed-type hypersensitivity in adult F344 rats after perinatal exposure to 2,3,7,8-tetrachlorodibenzo-p-dioxin. Toxicology. 1999;134:79-88.

95. Mocarelli P. Seveso: a teaching story. Chemosphere. 2001;43:391-402. 96. di Domenico A, Silano V, Viviano G, Zapponi G. Accidental release of 2,3,7,8-

tetrachlorodibenzo-p-dioxin (TCDD) at Seveso, Italy. II. TCDD distribution in the soil surface layer. Ecotoxicology & Environmental Safety. 1980;4:298-320.

97. Cerlesi S, Di Domenico A, Ratti S. Recovery yields of early analytical procedures to detect 2,3,7,8-tetrachloro-dibenzo-p-dioxin (TCDD) in soil samples at Seveso, Italy. Chemosphere. 1989;18:989-1003.

98. Gianotti F. Chloracne au tetrachloro-2,3,7,8-dibenzo-p-dioxine chez les enfants. Annales de Dermatologie et de Vénéréologie. 1977;104:825-829.

99. Needham LL, Gerthoux PM, Patterson DG, Jr., et al. Serum dioxin levels in Seveso, Italy, population in 1976. Teratogenesis, Carcinogenesis, & Mutagenesis. 1997;17:225-240.

100. Schwetz BA, Norris JM, Sparschu GL, et al. Toxicology of chlorinated dibenzo-p-dioxins. Environmental Health Perspectives. 1973;5:87-99.

101. Caramaschi F, del Corno G, Favaretti C, Giambelluca SE, Montesarchio E, Fara GM. Chloracne following environmental contamination by TCDD in Seveso, Italy. International Journal of Epidemiology. 1981;10:135-143.

102. Bisanti L, Bonetti F, Caramaschi F, et al. Experiences from the accident of Seveso. Acta Morphologica Academiae Scientiarum Hungaricae. 1980;28:139-157.

103. Rehder H, Sanchioni L, Cefis F, Gropp A. Pathologisch-embryologische Untersuchungen an Abortusfallen im Zusammenhang mit dem Seveso-Ungluck. Schweizerische Medizinische Wochenschrift. Journal Suisse de Medecine. 1978;108:1617-1625.

104. Mastroiacovo P, Spagnolo A, Marni E, et al. Birth defects in the Seveso area after TCDD contamination. JAMA. 1988;259:1668-1672.

105. Sirchia GG. Exposure to TCDD: immunologic effects. In: Dardanoni L, Miller RW, eds. Plans for clinical and epidemiological follow up after area-wide chemical contamination (Proceedings of an International Workshop,

121

Washington, DC, March 1982). Washington, DC: National Academy Press; 1982:234-266. In Consonni D, Pesatori AC, Zocchetti C, et al. Mortality in a population exposed to dioxin after the Seveso, Italy, accident in 1976: 25 years of follow-up. American Journal of Epidemiology. 2008;167:847-858.

106. Filippini G, Bordo B, Crenna P, Massetto N, Musicco M, Boeri R. Relationship between clinical and electrophysiological findings and indicators of heavy exposure to 2,3,7,8-tetrachlorodibenzo-p-dioxin. Scandinavian Journal of Work, Environment & Health. 1981;7:257-262.

107. Barbieri S, Pirovano C, Scarlato G, Tarchini P, Zappa A, Maranzana M. Long-term effects of 2,3,7,8-tetrachlorodibenzo-p-dioxin on the peripheral nervous system. Clinical and neurophysiological controlled study on subjects with chloracne from the Seveso area. Neuroepidemiology. 1988;7:29-37.

108. ISC. Special Office for Seveso, International Steering Committee. Minutes of the Sixth Meeting of the International Steering Committee, February 19–21, 1984. Final report and recommendations. Regione Lombardia, Milan, Italy; 1984:1-17. In In Consonni D, Pesatori AC, Zocchetti C, et al. Mortality in a population exposed to dioxin after the Seveso, Italy, accident in 1976: 25 years of follow-up. American Journal of Epidemiology. 2008;167:847-858.

109. Bertazzi PA, Consonni D, Bachetti S, et al. Health effects of dioxin exposure: A 20-year mortality study. American Journal of Epidemiology. 2001;153:1031-1044.

110. Pesatori AC, Consonni D, Bachetti S, et al. Short- and long-term morbidity and mortality in the population exposed to dioxin after the "Seveso accident". Industrial Health. 2003;41:127-138.

111. Henck JM, New MA, Kociba RJ, Rao KS. 2,3,7,8-tetrachlorodibenzo-p-dioxin: acute oral toxicity in hamsters. Toxicology & Applied Pharmacology. 1981;59:405-407.

112. Weber LW, Lebofsky M, Stahl BU, Smith S, Rozman KK. Correlation between toxicity and effects on intermediary metabolism in 2,3,7,8-tetrachlorodibenzo-p-dioxin-treated male C57BL/6J and DBA/2J mice. Toxicology & Applied Pharmacology. 1995;131:155-162.

113. Kociba RJ, Keyes DG, Beyer JE, et al. Results of a two-year chronic toxicity and oncogenicity study of 2,3,7,8-tetrachlorodibenzo-p-dioxin in rats. Toxicology & Applied Pharmacology. 1978;46:279-303.

114. EPA. (Environmental Protection Agency). Health assessment document for 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) and related compounds. EPA/600/bp-92/001a,b,c; 1994. In ATSDR, (Agency for Toxic Substances and Disease Registry) Toxicological profile for chlorinated dibenzo-p-dioxins (CDDs). 1998, Department of Health and Human Services, Public Health Service. Agency for Toxic Substances and Disease Registry: Atlanta, Ga.

115. Smith AH, Lopipero P. Evaluation of the toxicity of dioxins and dioxin-like PCBs: a health risk appraisal for the New Zealand population. ME number 351. Wellington: Minsitry for the Environment; 2001.

116. EPA. (Environmental Protection Agency) Exposure and Human Health Reassessment of 2,3,7,8-Tetrachlorodibenzo-p-Dioxin (TCDD) and Related Compounds National Academy Sciences (NAS) Review Draft . Available at <http://www.epa.gov/ncea/pdfs/dioxin/nas-review/> (Last accessed on Jan 09 2009). Washington, DC; 2003.

117. NAS. (National Acandemy of Sciences). Health Risks from Dioxin and Related Compounds: Evaluation of the EPA Reassessment. Washington, DC: The National Academies Press; 2006.

122

118. SCF. Opinion on the Risk Assessment of Dioxins and Dioxin-like PCBs in Food (22 November 2000), Available at <http://europa.eu.int/comm/food/fs/sc/scf/out78_en.pdf>. (Last on accessed Oct 22 2008). Brussels, Belgium: European Commission, Scientific Committee on Food; 2001.

119. Wilson CL, Bodnar JA, Brown BG, Morgan WT, Potts RJ, Borgerding MF. Assessment of dioxin and dioxin-like compounds in mainstream smoke from selected US cigarette brands and reference cigarettes. Food & Chemical Toxicology. 2008;46:1721-1733.

120. Buckland SJ, Ellis HK, Salter RT. Organochlorines in New Zealand: Ambient concentrations of selected organochlorines in air. Organochlorines Programme Ministry for the Environment. Available at <http://www.mfe.govt.nz/publications/hazardous/ambient-air-report-dec99/ambient-air-report-main-document-dec99.pdf> (Last accessed on Oct 27 2008). Wellington: Ministry for the Environment; 1999.

121. IARC. Polychlorinated Dibenzo-para-dioxin and Polychlorinated Dibenzofurans. IARC Monographs on the Evaluation of Carcinogenic Risks to Humans. Lyon, France: World Health Organization, International Agency for Research on Cancer; 1997:191-195.

122. Fingerhut MA, Halperin WE, Marlow DA, et al. Cancer mortality in workers exposed to 2,3,7,8-tetrachlorodibenzo-para-dioxin. New England Journal of Medicine. 1991;324:212-218.

123. Becher H, FleschJanys D, Kauppinen T, et al. Cancer mortality in German male workers exposed to phenoxy herbicides and dioxins. Cancer Causes & Control. 1996;7:312-321.

124. Hooiveld M, Heederik DJ, Bueno-de-Mesquita HB. Preliminary results of the second follow-up of a Dutch cohort of workers occupationally exposed to phenoxy herbicides. Organohalogen Compounds. 1996;30:185-189.

125. Ott MG, Zober A. Cause specific mortality and cancer incidence among employees exposed to 2,3,7,8-TCDD after a 1953 reactor accident. Occupational & Environmental Medicine. 1996;53:606-612.

126. COC. (Committee on Cancer). 2,3,7,8-tetrachlorodibenzo-p-dioxin: consideration of 1997 IARC monograph COC statement COC/99/S1 - May 1998. Available at <http://www.advisorybodies.doh.gov.uk/coc/tcdd.htm>. (Last accessed on Jan 08 2009); 1998.

127. JEFCA. Joint FAO/WHO Expert Committee on Food Additives. Fifty-seventh meeting Rome, 5–14 June 2001. Summary and Conclusions. Available at <http://www.who.int/ipcs/food/jecfa/summaries/en/summary_57.pdf>. (Last accessed on Oct 27 2008). Rome; 2001.

128. ECSCF. (European Commission Scientific Committee on Foods) Opinion of the Scientific Committee on Food on the risk assessment of dioxins and dioxin-like PCBs in food. Health and Consumer Protection Directorate-General Document SCF/CS/CNTM/DIOXIN/8, adopted on 22 November. Brussels, Belgium.; 2000.

129. ATSDR. (Agency for Toxic Substances and Disease Registry) Toxicological profile for chlorinated dibenzo-p-dioxins (CDDs). Atlanta, Ga: Department of Health and Human Services, Public Health Service. Agency for Toxic Substances and Disease Registry; 1998.

130. Pohl HR, Hicks HE, Jones DE, Hansen H, De Rosa CT. Public health perspectives on dioxin risks: Two decades of evaluations. Human and Ecological Risk Assessment. 2002;8:233-250.

123

131. Squire RA. Pathologic evaluations of selected tissues from the Dow Chemical TCDD and 2,4,5-T rat studies. Submitted to Carcinogen Assessment Group, U.S. Environmental Protection Agency on August 15 under contract no. 68-01-5092; 1980. In EPA, (Environmental Protection Agency) Exposure and Human Health Reassessment of 2,3,7,8-Tetrachlorodibenzop-Dioxin (TCDD) and Related Compounds. Part II: Health Assessment of 2,3,7,8-Tetrachlorodibenzo-p-Dioxin (TCDD) and Related Compounds . Chapter 6: Carcinogenicity of TCDD in Animals. Available at <http://www.epa.gov/ncea/pdfs/dioxin/part2/drich6.pdf> (Last accessed on Oct 11 2008). 2000.

132. Sauer RM. 2,3,7,8-Tetrachlorodibenzo-p-dioxin in Sprague-Dawley rats. Submitted to the Maine Scientific Advisory Panel by Pathco, Inc., Jamsville, MD. March 13; 1990. In EPA, (Environmental Protection Agency) Exposure and Human Health Reassessment of 2,3,7,8-Tetrachlorodibenzop-Dioxin (TCDD) and Related Compounds. Part II: Health Assessment of 2,3,7,8-Tetrachlorodibenzo-p-Dioxin (TCDD) and Related Compounds . Chapter 6: Carcinogenicity of TCDD in Animals. Available at <http://www.epa.gov/ncea/pdfs/dioxin/part2/drich6.pdf> (Last accessed on Oct 11 2008). 2000.

133. Goodman DG, Sauer RM. Hepatotoxicity and carcinogenicity in female Sprague-Dawley rats treated with 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD): a pathology working group reevaluation. Regulatory Toxicology & Pharmacology. 1992;15:245-252.

134. Giri AK. Mutagenic and genotoxic effects of 2,3,7,8-tetrachlorodibenzo-p-dioxin, a review. Mutation Research. 1986;168:241-248.

135. Wassom JS, Huff JE, Loprieno N. A review of the genetic toxicology of chlorinated dibenzo-p-dioxins. Mutation Research. 1977-1978;47:141-160.

136. Randerath K, Putman KL, Randerath E, Mason G, Kelley M, Safe S. Organ-specific effects of long term feeding of 2,3,7,8-tetrachlorodibenzo-p-dioxin and 1,2,3,7,8-pentachlorodibenzo-p-dioxin on I-compounds in hepatic and renal DNA of female Sprague-Dawley rats. Carcinogenesis. 1988;9:2285-2289.

137. Turteltaub KW, Felton JS, Gledhill BL, et al. Accelerator mass spectrometry in biomedical dosimetry: relationship between low-level exposure and covalent binding of heterocyclic amine carcinogens to DNA. Proceedings of the National Academy of Sciences of the United States of America. 1990;87:5288-5292.

138. Park JY, Shigenaga MK, Ames BN. Induction of cytochrome P4501A1 by 2,3,7,8-tetrachlorodibenzo-p-dioxin or indolo(3,2-b)carbazole is associated with oxidative DNA damage. Proceedings of the National Academy of Sciences of the United States of America. 1996;93:2322-2327.

139. Aschengrau A, Monson RR. Paternal military service in Vietnam and risk of spontaneous-abortion. Journal of Occupational and Environmental Medicine. 1989;31:618-623.

140. Wolfe WH, Michalek JE, Miner JC, et al. Paternal serum dioxin and reproductive outcomes among veterans of Operation Ranch Hand. Epidemiology. 1995;6:17-22.

141. Smith AH, Fisher DO, Pearce N, Chapman CJ. Congenital-defects and miscarriages among New-Zealand 2,4,5-T sprayers. Archives of Environmental Health. 1982;37:197-200.

142. Mocarelli P, Brambilla P, Gerthoux PM, Patterson DG, Jr., Needham LL. Change in sex ratio with exposure to dioxin. Lancet. 1996;348:409.

124

143. Basharova GR. Reproduction in families of workers exposed to 2,4,5-T intoxication. Organohalogen Compounds. 1996;30:315-318.

144. Schnorr TM, Lawson CC, Whelan EA, et al. Spontaneous abortion, sex ratio, and paternal occupational exposure to 2,3,7,8-tetrachlorodibenzo-p-dioxin. Environmental Health Perspectives. 2001;109:1127-32.

145. Hanify JA, Metcalf P, Nobbs CL, Worsley KJ. Aerial spraying of 2,4,5-T and human birth malformations - an epidemiological investigation. Science. 1981;212:349-351.

146. Stockbauer JW, Hoffman RE, Schramm WF, Edmonds LD. Reproductive outcomes of mothers with potential exposure to 2,3,7,8-tetrachlorodibenzo-p-dioxin. American Journal of Epidemiology. 1988;128:410-419.

147. Erickson JD, Mulinare J, McClain PW, et al. Vietnam veterans' risks for fathering babies with birth defects. JAMA. 1984;252:903-912.

148. CDC. Serum 2,3,7,8-tetrachlorodibenzo-p-dioxin levels in US Army Vietnam-era veterans. Centers for Disease Control. JAMA. 1988;260:1249-1254.

149. Giavini E, Prati M, Vismara C. Embryotoxic effects of 2,3,7,8 tetrachlorodibenzo-p-dioxin administered to female rats before mating. Environmental Research. 1983;31:105-110.

150. Huuskonen H, Unkila M, Pohjanvirta R, Tuomisto J. Developmental toxicity of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) in the most TCDD-resistant and -susceptible rat strains. Toxicology & Applied Pharmacology. 1994;124:174-180.

151. Gray LE, Jr., Ostby JS. In utero 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD) alters reproductive morphology and function in female rat offspring. Toxicology & Applied Pharmacology. 1995;133:285-294.

152. Bjerke DL, Peterson RE. Reproductive toxicity of 2,3,7,8-tetrachlorodibenzo-p-dioxin in male rats: different effects of in utero versus lactational exposure. Toxicology & Applied Pharmacology. 1994;127:241-249.

153. Schantz SL, Ferguson SA, Bowman RE. Effects of 2,3,7,8-tetrachlorodibenzo-p-dioxin on behavior of monkeys in peer groups. Neurotoxicology & Teratology. 1992;14:433-446.

154. Gifford JS, Buckland SJ, Judd MC, McFarlane PN, Anderson SM. Pentachlorophenol (PCP), PCDD, PCDF and pesticide concentrations in a freshwater lake catchment. Chemosphere. 1996;32:2097-2113.

155. Liu Y, Peng P, Li X, Zhang S, Ren M. Polychlorinated dibenzo-p-dioxins and dibenzofurans (PCDD/Fs) in water and suspended particulate matter from the Xijiang River, China. Journal of Hazardous Materials. 2008;152:40-47.

156. Suarez MP, Rifai HS, Palachek R, Dean K, Koenig L. Distribution of polychlorinated dibenzo-p-dioxins and dibenzofurans in suspended sediments, dissolved phase and bottom sediment in the Houston Ship Channel. Chemosphere. 2006;62:417-429.

157. Lohmann R, Nelson E, Eisenreich SJ, Jones KC. Evidence for dynamic air-water exchange of PCDD/Fs: a study in the Raritan Bay/Hudson River estuary. Environmental Science & Technology. 2000;34:3086 -3093.

158. Götz R, Enge P, Friesel P, et al. Sampling and analysis of water and suspended particulate matter of the River Elbe for polychlorinated dibenzo-p-dioxins (PCDDs) and dibenzofurans (PCDFs). Chemosphere. 1994;28:63-74.

159. Tysklind M, Fangmark I, Marklund S, Lindskog A, Thaning L, Rappe C. Atmospheric transport and transformation of polychlorinated dibenzo-p-dioxins and dibenzofurans. Environmental Science & Technology. 1993;27:2190-2197.

160. Broman D, Naf C, Zebuhr Y. Long-term high-volume and low-volume air sampling of polychlorinated dibenzo-para-dioxins and dibenzofurans and

125

polycyclic aromatic-hydrocarbons along a transect from urban to remote areas on the Swedish Baltic Coast. Environmental Science & Technology. 1991;25:1841-1850.

161. Wevers M, De Fre R, Cleuvenbergen RV, Rymen T. Concentrations of PCDDs and PCDFs in ambient air at selected locations in Flanders. Organohalogen Compounds. 1993;12:123-126.

162. Sugita K, Asada S, Yokochi T, Ono M, Okazawa T. Polychlorinated dibenzo-p-dioxins, dibenzofurans, co-planar PCBs and mono-ortho PCBs in urban air. Organohalogen Compounds. 1993;12:127-130.

163. Lucy R, Proffitt G. Dioxin Concentrations in Residential Soil, Paritu, New Plymouth. Available from <http://www.mfe.govt.nz/publications/hazardous/taranaki-dioxin-report-sep02/taranaki-dioxin-report-sep02.zip> (Last accessed on Nov 2 2008). Wellington: Ministry for the Environment; 2002.

164. Buckland SJ, Eltzer BD, Salter RT. Organochlorines in New Zealand: Ambient concentrations of selected organochlorines in soils. ME318. Organochlorines Programme Ministry for the Environment. Available at <http://www.mfe.govt.nz/publications/hazardous/organochlorine-concentration-in-soil-dec98.pdf> (Last accessed on Nov 02 2008). Wellington: Ministry for the Environment; 1998.

165. Rotard W, Christmann W, Knoth W. Background Levels of PCDD/F in Soils of Germany. Chemosphere. 1994;29:2193-2200.

166. Gälli R, Krebs J, Kraft M, Good M. PCDDs and PCDFs in soil samples from Switzerland. Chemosphere. 1992;24:1095-1102.

167. Broman D, Naf C, Rolff C, Zebuhr Y. Analysis of polychlorinated dibenzo-para-dioxins (PCDD) and polychlorinated dibenzofurans (PCDF) in soil and digested sewage-sludge from stockholm, Sweden. Chemosphere. 1990;21:1213-1220.

168. Fielder H, Cooper K, Andersson R, et al. PCDD/PCDF in soil and pine needle samples in a rural area in the United States of America. Organohalogen Compounds. 1995;24:285-292.

169. Sund KG, Carlo GL, Crouch RL, Senefelder BC. Background soil concentrations of phenolic-compounds, chlorinated herbicides, PCDDs and PCDFs in the Melbourne metropolitan-area. Australian Journal of Public Health. 1993;17:157-161.

170. Krauss P, Mahnke K, Freire L. Determination of PCDD/F and PCB in forest soils from Brazil. Organohalogen Compounds. 1995;24:357-361.

171. Zook DR, Rappe C. Environmental sources, distribution, and fate of polychlorinated dibenzodioxins, dibenzofurans, and related organochlorines. In: Schecter A, ed. Dioxins and health. New York: Plenum Press; 1994:80-113.

172. Hites R, Harless R. Atmospheric transport and deposition of polychlorinated dibenzo-p-dioxins and dibenzofurans. Office of Research and Development, US Environmental Protection Agency, Research Triangle Park, NC. EPA/600/3-91/002; 1991. In ATSDR, (Agency for Toxic Substances and Disease Registry) Toxicological profile for chlorinated dibenzo-p-dioxins (CDDs). 1998, Department of Health and Human Services, Public Health Service. Agency for Toxic Substances and Disease Registry: Atlanta, Ga.

173. Atkinson R. Atmospheric lifetimes of dibenzo-p-dioxins and dibenzofurans. Science of the Total Environment. 1991;104:17-33.

126

174. Fletcher CL, McKay WA. Polychlorinated dibenzo-p-dioxins (PCDDs) and dibenzofurans (PCDFs) in the aquatic environment - a literature-review. Chemosphere. 1993;26:1041-1069.

175. Muir DCG, Lawrence S, Holoka M, et al. Partitioning of polychlorinated dioxins and furans between water, sediments and biota in Lake Mesocosms. Chemosphere. 1992;25:119-124.

176. Webster GRB, Muldrew DH, Graham JJ, Sarna LP, Muir DCG. Dissolved organic mater mediated aquatic transport of chlorinated dioxins. Chemosphere. 1986;15:1379-1386.

177. Servos MR, Muir DCG, Webster GRB. The effect of dissolved organic-matter on the bioavailability of polychlorinated dibenzo-para-dioxins. Aquatic Toxicology. 1989;14:169-184.

178. Eltzer BD. Comparison of Point and Nonpoint Sources of Polychlorinated Dibenzo-P-Dioxins and Polychlorinated Dibenzofurans to Sediments of the Housatonic River. Environmental Science & Technology. 1993;27:1632-1637.

179. Wölz J, Engwall M, Maletz S, et al. Changes in toxicity and Ah receptor agonist activity of suspended particulate matter during flood events at the rivers Neckar and Rhine - a mass balance approach using in vitro methods and chemical analysis. Environmental Science & Pollution Research. 2008;15:536-553.

180. Orazio CE, Kapila S, Puri RK, Yanders AF. Persistence of chlorinated dioxins and furans in the soil environment. Chemosphere. 1992;25:1469-1474.

181. Osako M, Kim YJ. Influence of coexisting surface-active agents on leachability of dioxins in raw and treated fly ash from an MSW incinerator. Chemosphere. 2004;54:105-116.

182. Frankki S, Persson Y, Shchukarev A, Tysklind M, Skyllberg U. Partitioning of chloroaromatic compounds between the aqueous phase and dissolved and particulate soil organic matter at chlorophenol contaminated sites. Environmental Pollution. 2007;148:182-190.

183. Hofmann T, Wendelborn A. Colloid facilitated transport of polychlorinated dibenzo-p-dioxins and dibenzofurans (PCDD/Fs) to the groundwater at Ma Da area, Vietnam. Environmental Science & Pollution Research. 2007;14:223-224.

184. Persson Y, Shchukarev A, Oberg L, Tysklind M. Dioxins, chlorophenols and other chlorinated organic pollutants in colloidal and water fractions of groundwater from a contaminated sawmill site. Environmental Science and Pollution Research. 2008;15:463-471.

185. EPA. (Environmental Protection Agency) Bioaccumulation testing and interpretation for the purpose of sediment quality assessment: status and needs. EPA-823-R-00-001. Available at <http://www.epa.gov/ost/cs/biotesting/bioaccum.pdf> (Last accessed on Nov 01 2008). Washington, DC; 2000.

186. Ferrario J, Byrne C, Cleverly DH. 2,3,7,8-Dibenzo-p-dioxins in mined clay products from the United States: evidence for possible natural origin. Environmental Science & Technology. 2000;34:4524-4532.

187. Rappe C, Bergqvist PA, Fiedler H, Cooper KR. PCDD and PCDF contamination in catfish feed from Arkansas, USA. Chemosphere. 1998;36:2705-2720.

188. Sinkkonen S, Paasivirta J. Degradation half-life times of PCDDs, PCDFs and PCBs for environmental fate modeling. Chemosphere. 2000;40:943-949.

189. Crosby DG, Wong AS. Environmental degradation of 2,3,7,8-tetrachlorodibenzo-p-dioxin (TCDD). Science. 1977;195:1337-1338.

127

190. Miller GC, Hebert VR, Zepp RG. Chemistry and photochemistry of low-volatility organic-chemicals on environmental surfaces. Environmental Science & Technology. 1987;21:1164-1167.

191. Hutzinger O, Blumich MJ, Berg M, Olie K. Sources and fate of PCDDs and PCDFs: An overview. Chemosphere. 1985;14:581-600.

192. Choudhry GG, Webster GRB. Environmental Photochemistry of PCDDs .2. Quantum yields of the direct phototransformation of 1,2,3,7-tetrachlorodibenzo-para-dioxin, 1,3,6,8-tetrachlorodibenzo-para-dioxin, 1,2,3,4,6,7,8-heptachlorodibenzo-para-dioxin and 1,2,3,4,6,7,8,9-octachlorodibenzo-para-dioxin in aqueous acetonitrile and their sunlight half-lives. Journal of Agricultural and Food Chemistry. 1989;37:254-261.

193. Arthur MF, Frea JI. 2,3,7,8-Tetrachlorodibenzo-p-dioxin - aspects of its important properties and Its potential biodegradation in soils. Journal of Environmental Quality. 1989;18:1-11.

194. Field JA, Sierra-Alvarez R. Microbial transformation and degradation of polychlorinated biphenyls. Environmental Pollution. 2008;155:1-12.

195. Des Rosiers P. Methodologies for materials contaminated with PCDDs and related-compounds. Chemosphere. 1986;15:1513-1528.

196. McPeters AL, Overcash MR. Demonstration of photodegradation by sunlight of 2,3,7,8-tetrachlorodibenzo-p-dioxin in 6 cm soil columns. Chemosphere. 1993;27:1221-1234.

197. Srogi K. Overview of analytical methodologies for dioxin analysis. Analytical Letters. 2007;40:1647-1671.

198. Patterson DG, Jr., Hampton L, Lapeza CR, Jr., et al. High-resolution gas chromatographic/high-resolution mass spectrometric analysis of human serum on a whole-weight and lipid basis for 2,3,7,8-tetrachlorodibenzo-p-dioxin. Analytical Chemistry. 1987;59:2000-2005.

199. Nygren M, Hansson M, Sjostrom M, et al. Development and validation of a method for determination of PCDs and PCDFs in human-blood plasma - a multivariate comparison of blood and adipose-tissue levels between Vietnam veterans and matched controls. Chemosphere. 1988;17:1663-1692.

200. Van Den Berg M, Van Der Wielen FWM, Olie K, Van Boxtel CJ. The presence of PCDDs and PCDFs in human-breast milk from the Netherlands. Chemosphere. 1986;15:693-706.

201. Gross ML, Lay JO, Jr., Lyon PA, et al. 2,3,7,8-Tetrachlorodibenzo-p-dioxin levels in adipose tissue of Vietnam veterans. Environmental Research. 1984;33:261-268.

202. Chang RR, Jarman WM, Hennings JA. Sample cleanup by solid-phase extraction for the ultratrace determination of polychlorinated dibenzo-p-dioxins and dibenzofurans in biological samples. Analytical Chemistry. 1993;65:2420-2427.

203. MoH. Establishment of a maximum intake for dioxin. Public Health Perspectives 5(4): 7. Wellington, New Zealand: Ministry of Health; 2002.

204 IOM. Institute of Medicine. Veterans and Agent Orange. Available from <http://www.nap.edu/catalog.php?record_id=12662>. (Last accessed on Aug 14 2009). Washington: National Academies Press; 2008.

128

5.0 Tolerable Daily Intakes for a range of organochlorine pesticides as defined by various international bodies The Nelson Marlborough Health Board has requested a summary of TDIs for a range of OCPs, as derived by the various international bodies, and a comparison of the most conservative derived guidance values with those employed by Dr Craig Stevenson in his recalculation of the respective Total Hazard Index (THI) values.1 The THI values represent the total chronic daily intake of the respective OCP via all exposure pathways over the time interval of the Mapua remediation project divided by the TDI of that contaminant. For Mapua, THI’s were calculated monthly. Tables 5.1 to 5.8 summarise the calculation of tolerable daily intakes (TDIs) of various organochlorine pesticides (OCPs) that a consumer may ingest for the duration of his life without appreciable health risk, as determined by four major agencies: World Health Organisation (WHO), United States Environmental Protection Agency (EPA), National Institute of Public Health and the Environment, the Netherlands (RIVM), Agency for Toxic Substances and Disease Registry (ATSDR). There is some degree of variability amongst these agencies, as they have not always chosen the same study to determine their NOAELs, the applied safety factors and the derived guidance values. Nevertheless, all were determined with conservative safety factors to ensure safety is maintained for exposed populations. Table 5.9 in this section summarises these tables to present the lowest value for each organochlorine. EPA values were used by Dr Craig Stevenson to recalculate the Total Hazard Index (THI) for exposures to organochlorine pesticides.1 In the summary table (table 5.9), only the figures for lindane (RIVM), heptachlor (WHO) and hexachlorobenzene (ATSDR), are lower than the respective EPA guidance values. Caution must be raised with the derived RIVM guidance value for lindane. The Joint Meeting of the Food and Agriculture Organization and World Health Organisation on Pesticide Residues (JMPR), for example, noted the study underpinning this guidance value employed impure samples of lindane to demonstrate immunotoxic effects in rats; JMPR concluded, on the basis of another investigation, that lindane is not immunotoxic.2 Given the uncertainty surrounding the derivation of the respective guidance value, Dr Stevenson was therefore correct in not using it in his recalculations. The guidance values established by the EPA and the ATSDR for hexachlorobenzene were derived from the same study.3 Rats exposed to hexachlorobenzene for the duration of their lifetime had evidence of hepatic injury at doses 0.016, 0.08, 0.4 and 2 µg/kg (though only at doses greater than 0.08 was the incidence of injury significant when compared to the incidence in respective controls). The ATSDR incorrectly used the lowest dose as a LOAEL, even though the results were not statistically significant for this dosage. In contrast, the EPA interpreted the dose at 0.08 µg/kg as the correct NOAEL as the 0.4 ug/kg was the first dose to show statistical evidence of increased risk of hepatic injury. Therefore, the ATSDR guidance value should not be considered as an appropriately derived value. The WHO value was derived from a number of investigations involving rats and pigs. The derived guidance value, based on investigations that had no serious design flaws and findings that were correctly interpreted, involved rat chronic studies and should therefore be used to derive the respective THI’s.

129

The WHO derived their guidance value for heptachlor (0.1 ug/kg/day) on investigations with beagle dogs. Concerns must be raised with the underlying investigation that employed a total of four dogs per sex of which 2 per sex were sacrificed for the histological evaluation. Such low number of animals questions the statistical viability of their derived NOAEL. Comparing the number of animals in the EPA derived guidance value, the underlying study employed six groups of rats containing 20 per sex, making this study statistically robust. Again, the EPA study was the correct choice for the THI values. Therefore, with the exception of hexachlorobenzene, the EPA values are representative of the figures calculated by the various agencies and can be recommended as acceptable for use in determining the respective THIs. Recommended values are summarised in table 5.9.

Table 5.1 Tolerable daily intake values for aldrinf, as defined by various regulatory authorities

Agency Type of Exposure

Value

Year Species Critical effect

NOAEL/LOAEL Uncertainty factor

Principal study

WHO/FAO4 a

Chronic ingestion PTDI

0.1 µg/kg/day

1994 rat liver changes

0.025 mg/kg/day LOAEL

250(10A, 10H, 2.5L)

5

EPA6 b Chronic ingestion RfD

0.03 µg/kg/day



1000(10A, 10H, 10L)

5

RIVM7,c, Chronic ingestion TDI

0.1 µg/kg/day

1999/2000 rat liver changes


250(10A, 10H, 2.5L)

5

TGA8,d, Chronic ingestion TDI

0.1 µg/kg/day



250(10A, 10H, 2.5L)

5

ATSDRe Chronic ingestion MRL

0.03 µg/kg/day



1000(10A, 10H, 10L)

5

aWHO/FAO – World Health Organisation / Food and Agriculture Organization of the United Nations bEPA – United States Environmental Protection Agency cRIVM - Rijksinstituut voor Volksgezondheid en Milieu (National Institute of Public Health and the Environment, the Netherlands) dTGA - Australian Therapeutic Goods Administration eATSDR - Agency for Toxic Substances and Disease Registry fTDI for aldrin can be determined separately or as the sum of aldrin and dieldrin

130

Table 5.2 Tolerable daily intake values for dieldrinf, as defined by various regulatory authorities

Agency Type of Exposure

Value Year Species

Critical effect NOAEL/LOAEL

Uncertainty factor

Principal study

WHO/FAO4,a Chronic ingestion PTDI

0.1 µg/kg/day

1994 dog/rat liver changes


250(10A, 10H, 2.5L)

5 EPA6,b Chronic

ingestion RfD

0.05 µg/kg/day

1990 rat

Liver parenchymal cell changes

0.005 mg/kg/day NOAEL

100(10A, 10H)

9 RIVM 7,c,


0.1 µg/kg/day

1999/2000 dog/rat liver changes


250(10A, 10H, 2.5L)

5 TGA8,d, Chronic

ingestion ADI

0.1 µg/kg/day

2003 dog/rat liver changes


250(10A, 10H, 2.5L)

5 ATSDRe Chronic

ingestion MRL

0.05 µg/kg/day

2002 rat

Liver parenchymal cell changes


100(10A, 10H)

9

aWHO/FAO – World Health Organisation / Food and Agriculture Organization of the United Nations bEPA – United States Environmental Protection Agency cRIVM - Rijksinstituut voor Volksgezondheid en Milieu (National Institute of Public Health and the Environment, the Netherlands) dTGA - Australian Therapeutic Goods Administration eATSDR - Agency for Toxic Substances and Disease Registry fTDI for aldrin can be determined separately or as the sum of aldrin and dieldrin

Table 5.3 Tolerable daily intake values for lindane, as defined by various regulatory authorities

Agency Type of Exposure Value Year Species Critical effect NOAEL/LOAEL

Uncertainty factor

Principal study

WHO/FAO4a Chronic ingestion PTDI

5 µg/kg/day

2002 rat

Periacinar hepatocellular hypertrophy


100(10A, 10H)

10 EPA6,b Chronic

ingestion RfD

0.3 µg/kg/day

1988 rat

Liver and kidney toxicity


1000(10A, 10H, 10D)

11 RIVM7,c Chronic

ingestion TDI

0.04 µg/kg/day

1999/2000 rat Immunotoxic effects


300(10A, 10H, 3L)

12 TGA8,d, Chronic

ingestion ADI

3 µg/kg/day

- -


- -

aWHO/FAO – World Health Organisation / Food and Agriculture Organization of the United Nations bEPA – United States Environmental Protection Agency cRIVM - Rijksinstituut voor Volksgezondheid en Milieu (National Institute of Public Health and the Environment, the Netherlands) dTGA - Australian Therapeutic Goods Administration

131

Table 5.4 Tolerable daily intake values for chlordane, as defined by various regulatory authorities

aWHO/FAO – World Health Organisation / Food and Agriculture Organization of the United Nations bEPA – United States Environmental Protection Agency cTGA - Australian Therapeutic Goods Administration dATSDR - Agency for Toxic Substances and Disease Registry

Table 5.5 Tolerable daily intake values for DDT/DDE/DDDe, as defined by various regulatory authorities


Uncertainty factor

Principal study


10 µg/kg/day

2000 rat Development toxicology

1 mg/kg/day NOAEL

100(10A, 10H)

16 EPA6,b

Chronic ingestion RfD

0.5 µg/kg/day

1996 rat Liver lesions


100(10A, 10H)

17 RIVM 7,c


0.5 µg/kg/day

1999/2000 rat Liver lesions 0.05 mg/kg/day NOAEL

100(10A, 10H)

17 TGA8,d

Chronic ingestion ADI

2 µg/kg/day

2003 rat -


100(10A, 10H)

-

aWHO/FAO – World Health Organisation / Food and Agriculture Organization of the United Nations bEPA – United States Environmental Protection Agency cRIVM - Rijksinstituut voor Volksgezondheid en Milieu (National Institute of Public Health and the Environment, the Netherlands) dTGA - Australian Therapeutic Goods Administration eTDI for DDT can be determined separately or as the sum of DDT, DDE and DDD


Uncertainty factor

Principal study


0.5 µg/kg/day

1994 rat

Hepatocellular swelling/ necrosis

50 µg/kg/day NOAEL

100(10A, 10H)

13 EPA6,b

Chronic ingestion RfD

0.5 µg/kg/day

1998 mice Hepatic necrosis


300(10A, 10H, 3D)

14 TGA8,c

Chronic ingestion ADI

0.5 µg/kg/day

2003 -

Hepatocellular swelling/ necrosis

50 µg/kg/day NOAEL

100(10A, 10H)

13 ATSDR d Chronic

ingestion MRL

0.6 µg/kg/day

1994 rat Hepatic effects 55 µg/kg/day NOAEL

100(10A, 10H)

15

132

Table 5.6 Tolerable daily intake values for heptachlor, as defined by various regulatory authorities


Uncertainty factor

Principal study

WHO/FAO4a Chronic ingestion PTDI

0.1 µg/kg/day

1994 dog

Histopathological changes in the liver

0.025mg/kg/day NOAEL

200(10A, 10H, 2D)

18 EPA6,b Chronic

ingestion RfD

0.5 µg/kg/day

1991 rat Liver weight increases


300(10A, 10H, 3D)

19 TGA8,c Chronic

ingestion ADI

0.5 µg/kg/day

2003 rat Liver weight increases


- -

aWHO/FAO – World Health Organisation / Food and Agriculture Organization of the United Nations bEPA – United States Environmental Protection Agency cTGA - Australian Therapeutic Goods Administration

Table 5.7 Tolerable daily intake values for heptachlor epoxidec, as defined by various regulatory authorities


Uncertainty factor

Principal study

WHO/FAO4,

a Chronic ingestion PTDI

0.1 µg/kg/day

1994 dog

Histopathological changes in the liver


200(10A, 10H, 2D)

18 EPA6,b Chronic

ingestion RfD

0.013 µg/kg/day

1991 dog Liver changes


1000(10A, 10H, 10L)

20

aWHO/FAO – World Health Organisation / Food and Agriculture Organization of the United Nations bEPA – United States Environmental Protection Agency cWHO/FAO use values for heptachlor to determine heptachlor epoxide TDI.

Table 5.8 Tolerable daily intake values for hexachlorobenzene, as defined by various regulatory authorities

Agency Type of Exposure Value Year Species

Critical effect NOAEL/LOAEL

Uncertainty factor

Principal study


0.17 µg/kg/day

1997 Rat &pig

hepatic effects


300(10A, 10H, 3D)

21 EPA6,b Chronic

ingestion RfD

0.8 µg/kg/day

1991 rat hepatic effects


100(10A, 10H)

3 RIVM7,c Chronic

ingestion TDI

0.5 µg/kg/day



100(10A, 10H)

22, 23 ATSDR d Chronic

ingestion MRL

0.05 µg/kg/day



300(10A, 10H, 3D)

3

aWHO/FAO – World Health Organisation / Food and Agriculture Organization of the United Nations bEPA – United States Environmental Protection Agency cRIVM - Rijksinstituut voor Volksgezondheid en Milieu (National Institute of Public Health and the Environment, the Netherlands) dATSDR - Agency for Toxic Substances and Disease Registry

133

Table 5.9 Summary and recommendations of the lowest guidance values for the respective organochlorines

Pesticide Lowest TDI (µg/kg/day)

Lowest recommended values (µg/kg/day)

Aldrin 0.03 0.03 Dieldrin 0.05 0.05 Lindane 0.04 0.3 Chlordane 0.5 0.5 DDT/DDE/DDD 0.5 0.5 Heptachlor 0.1 0.5 Heptachlor epoxide 0.013 0.013 Hexachlorobenzene 0.05 0.17

134

5.1 References 1. Stevenson C. Recalculation of Hazard Indices and Cancer Risks for the Mapua

FCC Contaminated Site Remediation Project. Auckland, New Zealand: Air Environmental Science Ltd; 2009.

2. JMPR. Pesticide residues in food . Report of the Joint Meeting of the FAO Panel of Experts on Pesticide Residues in Food and the Environment and the WHO Core Assessment Group on Pesticide Residues. Available at <http://www.fao.org/ag/AGP/AGPP/Pesticid/JMPR/Download/2002_rep/2002JMPRReport2.pdf> (Last accessed on Aug 25 2009). Rome, Italy ; 2002.

3. Arnold DL, Moodie CA, Charbonneau SM, et al. Long-Term Toxicity of Hexachlorobenzene in the Rat and the Effect of Dietary Vitamin-a. Food and Chemical Toxicology . 1985;23:779-793.

4. FAO/WHO. Pesticide residues in food — 1994. Report of the Joint Meeting of the FAO Panel of Experts on Pesticide Residues in Food and the Environment and a WHO Expert Group on Pesticide Residues (FAO Plant Production and Protection Paper 127). Rome: Food and Agriculture Organization of the United Nations; 1995.

5. Fitzhug OG, Nelson AA, Quaife ML. Chronic oral toxicity of aldrin and dieldrin in rats and dogs. Food & Cosmetics Toxicology . 1964;2:551-562.

6. EPA. (U.S. Environmental Protection Agency) Integrated Risk Information System (IRIS). Available at <http://cfpub.epa.gov/ncea/iris/index.cfm> (Last accessed on Nov 22 2008). Washington, DC: U.S. Environmental Protection Agency; 2008.

7. RIVM. Re-Evaluation of Human Toxicological Maximum Permissible Risk Levels. Available at <http://www.rivm.nl/bibliotheek/rapporten/711701025.pdf> (Last accessed on Nov 22 2008). Netherlands: Rijksinstituut voor Volksgezondheid en Milieu (RIVM; National Institute of Public Health and the Environment, the Netherlands); 2001.

8. TGA. Acceptable Daily Intakes for Agricultural and Veterinary Chemicals. Available at <http://www.health.gov.au/internet/main/publishing.nsf/Content/E8F4D2F95D616584CA2573D700770C2A/$File/Amended%20ADI%20List%20-%20April%202008.pdf> (Last accessed on Nov 22 2008). Canberra, Australia: Office of Chemical Safety, Therapeutic Goods Administration; 2007.

9. Walker AIT, Stevenson DE, Robinson J, Thorpe R, Roberts M. The toxicology and pharmacodynamics of dieldrin (HEOD): Two-year oral exposures of rats and dogs. Toxicology & Applied Pharmacology . 1969;15:345-373.

10. Amyes SJ. Lindane: Combined oncogenicity and toxicity study by dietary administration to Wistar rats for 104 weeks. Unpublished report No. 90/CIL002/0839 from Life Science Research Ltd. Suffolk, England: Submitted to WHO by CIEL, Brussels, Belgium; 1990. In FAO/WHO. Pesticide residues in food — 1994. Report of the Joint Meeting of the FAO Panel of Experts on Pesticide Residues in Food and the Environment and a WHO Expert Group on Pesticide Residues (FAO Plant Production and Protection Paper 127). Rome: Food and Agriculture Organization of the United Nations; 1995.

135

11. Zoecon corporation Z. Rat, subchronic oral bioassay. MRID No. 00128356. Available from U.S. EPA FOI, Washington, DC; 1983. In EPA. (U.S. Environmental Protection Agency) Integrated Risk Information System (IRIS). Available at <http://cfpub.epa.gov/ncea/iris/index.cfm> (Last accessed on Nov 22 2008). Washington, DC: U.S. Environmental Protection Agency; 2008.

12. Meera P, Rao PR, Shanker R, Tripathi O. Immunomodulatory effects of gamma-HCH (Lindane) in mice. Immunopharmacology and Immunotoxicology . 1992;14:261-282.

13. Ihui S, Hayakawa T, Yonemura T, Takamura F, Takahashi Y, Hashizume M. Thirty-month chronic toxicity and tumourigenicity test in rats with chlordane technical. Unpublished report by Research Institute for Animal Science in Biochemistry and Toxicology. Chicago, IL: Submitted to WHO by Velsicol Chemical Corp; 1983. In FAO/WHO. Pesticide residues in food — 1994. Report of the Joint Meeting of the FAO Panel of Experts on Pesticide Residues in Food and the Environment and a WHO Expert Group on Pesticide Residues (FAO Plant Production and Protection Paper 127). Rome: Food and Agriculture Organization of the United Nations; 1995.

14. Khasawinah AM, Grutsch JF. Chlordane: 24-month tumorigenicity and chronic toxicity test in mice. Regulatory Toxicology and Pharmacology . 1989;10:244-254.

15. Khasawinah AM, Grutsch JF. Chlordane - 30-Month Tumorigenicity and Chronic Toxicity Test in Rats. Regulatory Toxicology and Pharmacology . 1989;10:95-109.

16. FAO/WHO. Pesticide residues in food — 2000. Evaluations — 2000. Part II — Toxicology (WHO/PCS/01.3). Geneva: World Health Organization, Joint FAO/WHO Meeting on Pesticide Residues; 2001.

17. Laug EP, Nelson AA, Fitzhugh OG, Kunze FM. Liver cell alteration and DDT storage in the fat of the rat induced by dietary levels of 1 to 50 p.p.m. DDT. Journal of Pharmacology & Experimental Therapeutics . 1950;98:268-273.

18. Wazeter FX, Geil RG, Goldenthal EI, Cookson KM, Howell DG. Two-year oral study in beagle dogs. Unpublished report No. 163-048 by International Research and Development Corp. Chicago, IL: Submitted to WHO by Velsicol Chemical Corp; 1971. In IPCS. Heptachlor. International Programme on Chemical Safety. Concise International Chemical Assessment Document 70. Geneva: World Health Organisation; 2006. Available at <http://www.inchem.org/documents/cicads/cicads/cicad70.htm> (Last accessed on August 30 2009).

19. Velsico Chemical Corporation VC. MRID No. 00062599. Available from EPA. Write to FOI, EPA, Washington, DC 20460; 1955. In EPA. (U.S. Environmental Protection Agency) Integrated Risk Information System (IRIS). Available at <http://cfpub.epa.gov/ncea/iris/index.cfm> (Last accessed on Nov 22 2008). Washington, DC: U.S. Environmental Protection Agency; 2008.

20. Dow Chemical Company DC. MRID No. 00061912. Available from EPA. Write to FOI, EPA, Washington, DC 20460; 1958. In EPA. (U.S. Environmental Protection Agency) Integrated Risk Information System (IRIS). Available at <http://cfpub.epa.gov/ncea/iris/index.cfm> (Last accessed on Nov 22 2008). Washington, DC: U.S. Environmental Protection Agency; 2008.

136

21. IPCS. Hexachlorobenzene. International Programme on Chemical Safety. Environmental Health Criteria 195 . Geneva: World Health Organisation; 1997.

22. Mollenhauer HH, Johnson JH, Younger RL, Clark DE. Ultrastructural changes in liver of the rat fed hexachlorobenzene. American Journal of Veterinary Research . 1975;36:1777-1781.

23. Mollenhauer HH, Johnson JH, Younger RL, Clark DE. A unique intracellular aberration related to hexachlorobenzene ingestion. American Journal of Veterinary Research . 1976;37:847-850.

137

6.0 Appendices Appendix 1 Abbreviations and Glossary Acute exposure: Exposure to a chemical for a duration of 14 days or less Ah receptor: Aryl hydrocarbon receptor ADI: Acceptable Daily Intake. It represents the acceptable daily intake

of the amount of a substance in food (such as additives in food, residues of pesticides and veterinary drugs in food) a consumer may ingest for the duration of their life without appreciable risk. It is based on the NOAEL (see NOAEL) to which appropriate safety or uncertainty factors are applied. See tolerable daily intake

Association: In statistics, an association comes from two variables that are

related; it is often confused with causality. An association does not imply a causal relationship. In statistics, correlation and association are related but not entirely overlapping concepts

ATSDR: Agency for Toxic Substances and Disease Registry Bioaccumulation: A process whereby chemicals are retained in body tissue and may

increase in time Biomagnification: The increase in tissue accumulation in species higher in the

natural food chain as a result of the ingestion of species that are contaminated

Bio-concentration Factor: A measure of a chemical’s ability to accumulate in a living

organism; it is the ratio of concentration in that living organisms tissue (defined as mg/kg) to the concentration in the surrounding environment (mg/L – if that environment is aquatic)

Body Burden: The total amount of accumulated chemical present in an

organism and represents the result of multiple years of exposure. For dioxin it represents the burden of the 17 toxic congeners, predominantly accumulated in adipose tissue (and serum lipids) and is typically expressed as picogram TEQ per gram of serum lipid (or alternatively defined per kg body weight)

CAS: Chemical Abstract Service Registry Number Chronic exposure: See Repeated exposure

138

Clearance: The rate at which a given chemical is removed (typically from

the kidney) from a solution; this is defined as the number of millilitres of solution that is completely cleared of a given chemical over a defined period of time

Congener Molecules that belong to a closely related chemical family or

have a similar chemical structure Cytochrome P450: Comprises of a large family of distinct yet related heme-

containing enzymes that catalyse a wide variety of chemical biotransformation reactions, predominantly in the liver and gut epithelial cells

Developmental Effects Adverse effects occurring to an offspring of an animal from pre-

natal development to their sexual maturity; effects are due to exposure of a chemical to either parent prior to the offspring’s conception

EDI: Estimated daily intake represents the summation of the TEQs

present in all foods in a average diet EPA: Environmental Protection Agency (United States) FAO: Food and Agriculture Organization Hazard: Capability of a substance to cause an adverse effect. It is the

inherent property of an agent having the potential to cause adverse effects.

IARC: International Agency for Research on Cancer (World Health

Organisation) IUPAC: The International Union of Pure and Applied Chemistry, whose

responsibility is to establish official names of chemical elements and compounds

JMPR: Joint Meeting of the FAO Panel of Experts on Pesticide Residues

in Food and the Environment and the WHO Core Assessment Group on Pesticide Residues

LD50: The statistically derived single dose of a chemical that can be

expected to cause death in 50% of a given population of test animals under a defined set of experimental conditions.

LOAEL: Lowest-Observed-Adverse Effect Level. This is the lowest dose

of a chemical, which causes an observable adverse effect on a test animal or organism. Any dose or concentration that is below this value will (theoretically) not result in an adverse effect

139

log Kow: The log oil water partition coefficient defines the ratio of

concentrations of a substance in equilibrium in a two-phasic system i.e. oil and water. The coefficient is defined as a logged ratio, indicating if a molecule is hydrophilic (values less than one) or lipophilic (values greater than 3)

Meta Analysis: A method of analysis that combines the result of a number of

surveys to determine the underlying processes MCL: Maximum Contaminant Level. The highest amount of a

contaminant allowed by EPA in water supplied by a municipal water system; also referred to as "drinking water standard."

MRL: Minimal Risk Levels, devised by ATSDR, is an estimate of the

daily human exposure to a hazardous substance that is likely to be without appreciable risk of adverse noncancerous health effects over a specified duration of exposure (acute, subacute and chronic). These estimates are intended to serve as a screening tool, (predominantly used by ATSDR health assessors) to identify contaminants and potential health effects that may be of concern at hazardous waste sites

NOAEL: No Observable Adverse Effect Level. This is the highest dose of

a chemical, which doesn’t cause an observable adverse effect on a test animal or organism. Any dose or concentration that exceeds this value will (theoretically) result in an adverse effect

Operation Ranch Hand: This describes a US Air Force operation to distribute the

herbicide Agent Orange in Vietnam from 1962 to 1971. Consequent to this war, cohort studies have closely monitored personnel who served in units involved in this operation; they had elevated exposure to TCDD, a contaminant of this herbicide. Presently the US Air Force maintains an exposure registry of about 1,200 personnel

PCB Polychlorinated biphenyl PCDD: polychlorinated dibenzo-p-dioxin PCDF: Polychlorinated dibenzofuran PPM Parts per million. A unit of measurement that expresses the

number of parts of a substance contained within a million parts of either gas, liquid or solid

Relative risk: A measure of the risk of disease in an exposed group compared

with the risk of an unexposed group. A relative risk of 1.0 means

140

that risks in the two groups are the same. If that risk is doubled for an exposed population, the relative risk is 2.0

Repeated exposure: Repeated exposure is divided into 3 categories:

• Subacute exposure refers to repeated exposure to a chemical for one month or less;

• Subchronic for 1 to 3 months • Chronic for more than 3 months

RfD: Reference dose; an estimate of daily exposure a consumer may

ingest for the duration of their life without appreciable risk. It is derived from an appropriate NOAEL (see NOAEL) with applied safety or uncertainty factors.

Risk: Chance or probability of adverse effects occurring under specific

exposure conditions. This can be expressed as risk (R) equals toxicity (T) times exposure (E), or R = T * E.

Sub-acute Exposure: See Repeated exposure Sub-chronic Exposure: See Repeated exposure TCDD: 2,3,7,8-tetrachlorodibenzo-p-dioxin TEF: Toxic equivalency factor. This term compares the potential

toxicity of each dioxin-like compound with the most toxic congener, TCDD, which has been defined as 1

Temporary tolerances: A temporary legal limit, established by EPA, for the maximum

amount of a pesticide residue which may be present in or on a food.

TEQ: Toxic equivalence. In a given mixture of dioxin congeners, the

TEQ represents the summation of each congener’s TEF multiplied by its respective mass concentration

Teratogen: Describes a substance that can cause non-heritable birth defects TDI: Tolerable Daily Intake. A TDI is defined as an estimate of the

intake of a substance over a lifetime considered to be without appreciable health risk. TDI is similar in definition and intent to terms such as the ADI. Like an ADI, they are based on extrapolated animal data to which appropriate safety or uncertainty factors are applied. See acceptable daily intake

THI: Total Hazard Index. The THI represent the total chronic daily

intake of a substance via all exposure pathways divided its TDI.

141

Toxicity: A chemical’s capacity to cause direct harm to an organism. Acute

toxicity is the poisoning that occurs following a single exposure whereas chronic toxicity is the result of repeated or low-level long term exposures to a toxic agent causing injuries such as cancer, reproductive disorders etc

TLV: Threshold Limit Value. The concentration of an airborne

substance that a healthy person can be exposed to for a 40-hour work week without adverse effect; a workplace exposure standard

UF: Uncertainty factor. An UF is a mathematical expression of

uncertainty that is used to protect populations from hazards which cannot be assessed with high precision. A maximum UF of 10 is assigned for interspecies extrapolation, etc.

142

Appendix 2 U.S. Environmental Protection Agency maximum contaminant levels goals for organic contaminants in drinking water The EPA has formulated guidelines for the acceptable concentrations of various contaminants in drinking water, based solely on possible health risks and exposure (Code of Federal Regulations. 40 CFR 141.50). They are known as Maximum Contaminant Level Goals (MLCGs), which are set to zero for the following compounds (as the EPA believes this level of protection would not cause any of the potential health problems). They apply to community water systems and non-transient, non-community water systems. As this level of protection is almost unattainable, the EPA has also formulated guidelines for the acceptable concentrations of these agents in drinking water (see appendix 4). MCLGs are zero for the following contaminantsa

Benzene Vinyl chloride Carbon tetrachloride 1,2-dichloroethane Trichloroethylene Acrylamide Alachlor Chlordane Dibromochloropropane 1,2-Dichloropropane Epichlorohydrin Ethylene dibromide Heptachlor Heptachlor epoxide Pentachlorophenol Polychlorinated biphenyls (PCBs) Tetrachloroethylene Toxaphene Benzo[a]pyrene Dichloromethane (methylene chloride) Di(2-ethylhexyl)phthalate Hexachlorobenzene 2,3,7,8-TCDD (Dioxin)

a http://law.justia.com/us/cfr/title40/40-22.0.1.1.3.6.16.1.html

143

Appendix 3 U.S. Environmental Protection Agency maximum contaminant levels for organic contaminants The following maximum contaminant levels (MCL) for synthetic organic contaminants (Code of Federal Regulations: 40 CFR 141.61) apply to community water systems and non-transient, non-community water systemsa. The EPA suggests such a value is associated with little to no risk of adverse health effects or cancer. CAS No. Contaminant MCL (mg/L) (1) 15972-60-8 Alachlor 0.002 (2) 116-06-3 Aldicarb 0.003 (3) 1646-87-3 Aldicarb sulfoxide 0.004 (4) 1646-87-4 Aldicarb sulfone 0.002 (5) 1912-24-9 Atrazine 0.003 (6) 1563-66-2 Carbofuran 0.04 (7) 57-74-9 Chlordane 0.002 (8) 96-12-8 Dibromochloropropane 0.0002 (9) 94-75-7 2,4-D 0.07 (10) 106-93-4 Ethylene dibromide 0.00005 (11) 76-44-8 Heptachlor 0.0004 (12) 1024-57-3 Heptachlor epoxide 0.0002 (13) 58-89-9 Lindane 0.0002 (14) 72-43-5 Methoxychlor 0.04 (15) 1336-36-3 Polychlorinated biphenyls 0.0005 (16) 87-86-5 Pentachlorophenol 0.001 (17) 8001-35-2 Toxaphene 0.003 (18) 93-72-1 2,4,5-TP 0.05 (19) 50-32-8 Benzo[a]pyrene 0.0002 (20) 75-99-0 Dalapon 0.2 (21) 103-23-1 Di(2-ethylhexyl)adipate 0.4 (22) 117-81-7 Di(2-ethylhexyl)phthalate 0.006 (23) 88-85-7 Dinoseb 0.007 (24) 85-00-7 Diquat 0.02 (25) 145-73-3 Endothall 0.1 (26) 72-20-8 Endrin 0.002 (27) 1071-53-6 Glyphosate 0.7 (28) 118-74-1 Hexacholorbenzene 0.001 (29) 77-47-4 Hexachlorocyclopentadiene 0.05 (30) 23135-22-0 Oxamyl (Vydate) 0.2 (31) 1918-02-1 Picloram 0.5 (32) 122-34-9 Simazine 0.004 (33) 1746-01-6 2,3,7,8-TCDD (Dioxin) 3.0 x10-8 a http://law.justia.com/us/cfr/title40/40-22.0.1.1.3.7.16.2.html

144

Appendix 4 Agency for Toxic Substances and Disease Registry (ATSDR) Minimal Risk Levels The following list of minimal risk levels (MRL), pertinent for this review, have been reproduced from the ATSDR documenta Chemical Name Route Duration MRL (mg/kg/day) Atrazine Oral Acute 0.01 Subacute 0.003 DDT, P,P'- Oral Acute 0.0005 Subacute 0.0005 Dieldrin Oral Subacute 0.0001 Chronic 0.00005 PCP Oral Acute 0.005 Subacute 0.001 Chronic 0.001 PCB Oral Subacute 0.03 Chronic 0.02 PeCDF Oral Acute 0.001 Subacute 0.00003 TCDD Oral Acute 0.0002 Subacute 0.00002 Chronic 0.000001

aMinimal Risk Levels for Hazardous Substances (2007). Agency for Toxic Substances & Disease Registry. Department of Health and Human Sevices, Atlanta, GA. (http://www.atsdr.cdc.gov/mrls/pdfs/mrllist_11_07.pdf ). Last accessed 27 September, 2008.

Toxicological Profiles of Chemicals of Concern at Mapua finalMapua, a small coastal village located...

Documents

Transcript of Toxicological Profiles of Chemicals of Concern at Mapua finalMapua, a small coastal village located...