Trichloroethene in drinking-water

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WHO/HEP/ECH/WSH/2020.10

Trichloroethene in drinking-water

Background document for development of WHO Guidelines for drinking-water quality

This document replaces document reference number WHO/SDE/WSH/05.08/22

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WHO/HEP/ECH/WSH/2020.10

© World Health Organization 2020

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Preface

Access to safe drinking-water is essential to health, a basic human right and a component of effective policy for health protection. A major World Health Organization (WHO) function to support access to safe drinking-water is the responsibility “to propose ... regulations, and to make recommendations with respect to international health matters ...”, including those related to the safety and management of drinking-water.

The first WHO document dealing specifically with public drinking-water quality was published in 1958 as International standards for drinking-water. It was revised in 1963 and 1971 under the same title. In 1984–1985, the first edition of the WHO Guidelines for drinking-water quality (GDWQ) was published in three volumes: Volume 1, Recommendations; Volume 2, Health criteria and other supporting information; and Volume 3, Surveillance and control of community supplies. Second editions of these volumes were published in 1993, 1996 and 1997, respectively. Addenda to Volumes 1 and 2 of the second edition were published in 1998, addressing selected chemicals. An addendum on microbiological aspects, reviewing selected microorganisms, was published in 2002. The third edition of the GDWQ was published in 2004, the first addendum to the third edition was published in 2006, and the second addendum to the third edition was published in 2008. The fourth edition was published in 2011, and the first addendum to the fourth edition was published in 2017.

The GDWQ are subject to a rolling revision process. Through this process, microbial, chemical and radiological aspects of drinking-water are subject to periodic review, and documentation relating to aspects of protection and control of drinking-water quality is accordingly prepared and updated.

Since the first edition of the GDWQ, WHO has published information on health criteria and other information to support the GDWQ, describing the approaches used in deriving guideline values, and presenting critical reviews and evaluations of the effects on human health of the substances or contaminants of potential health concern in drinking-water. In the first and second editions, these constituted Volume 2 of the GDWQ. Since publication of the third edition, they comprise a series of free-standing monographs, including this one.

For each chemical contaminant or substance considered, a background document evaluating the risks to human health from exposure to that chemical in drinking-water was prepared. The draft health criteria document was submitted to a number of scientific institutions and selected experts for peer review. The draft document was also released to the public domain for comment. Comments were carefully considered and addressed, as appropriate, taking into consideration the processes outlined in the Policies and procedures used in updating the WHO guidelines for drinking-water quality and the WHO Handbook for guideline development. The revised draft was submitted for final evaluation at expert consultations.

During preparation of background documents and at expert consultations, careful consideration was given to information available in previous risk assessments carried out by the International Programme on Chemical Safety, in its Environmental Health Criteria monographs and Concise International Chemical Assessment Documents; the International Agency for Research on Cancer; the Joint Food and Agriculture Organization of the United Nations (FAO)/WHO Meeting on Pesticide Residues; and the Joint FAO/WHO Expert Committee on Food Additives (which evaluates contaminants such as lead, cadmium, nitrate and nitrite, in addition to food additives).

Further up-to-date information on the GDWQ and the process of their development is available on the WHO website and in the current edition of the GDWQ.

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Acknowledgements

The update of this background document on trichloroethene in drinking-water for the development of the World Health Organization (WHO) Guidelines for drinking-water quality (GDWQ) was led by Emanuela Testai of the Istituto Superiore di Sanità of Italy. The contributions of Professor John Fawell, of Cranfield University, United Kingdom, and France Lemieux, Health Canada, who led the update of section 7 on practical considerations, are gratefully acknowledged.

The work of the following experts was crucial in the development of this document and others in the second addendum to the fourth edition:

Dr M Asami, National Institute of Public Health, Japan Dr RJ Bevan, independent consultant, United Kingdom Mr R Carrier, Health Canada, Canada

Dr J Cotruvo, Joseph Cotruvo & Associates and NSF International WHO Collaborating Centre, United States of America

Dr D Cunliffe, South Australia Department of Health, Australia

Dr L d’Anglada, Environmental Protection Agency, United States of America Dr A Eckhardt, Umweltbundesamt (Federal Environment Agency), Germany Professor JK Fawell, Cranfield University, United Kingdom

Dr A Hirose, National Institute of Health Sciences of Japan

Dr A Humpage, University of Adelaide (formerly South Australian Water Corporation), Australia Dr P Marsden, Drinking Water Inspectorate, United Kingdom

Professor Y Matsui, Hokkaido University, Japan

Dr E Ohanian, Environmental Protection Agency, United States of America Professor CN Ong, National University of Singapore, Singapore

Dr J Strong, Environmental Protection Agency, United States of America Dr E Testai, National Institute of Health, Italy

The draft text was discussed at the expert consultations for the second addendum to the fourth edition of the GDWQ, held on 28–30 March 2017 and 13–14 July 2018. The final version of the document takes into consideration comments from both peer reviewers and the public, including N Kobayashi, National Institute of Health Science, Japan; B Lampe, NSF International, United States of America; G Miller, Environmental Protection Agency, United States of America; and M Templeton, Imperial College London, United Kingdom.

The coordinator was Ms J De France, WHO, with support from Dr V Bhat, formerly of NSF International, United States of America. Strategic direction was provided by Mr B Gordon, WHO. Dr A Tritscher, formerly of WHO, and Dr P Verger, WHO, provided liaisons with the Joint FAO/WHO Expert Committee on Food Additives and the Joint FAO/WHO Meeting on Pesticide Residues. Dr R Brown and Ms C Vickers, WHO, provided liaisons with the International Programme on Chemical Safety. Dr M Perez contributed on behalf of the WHO Radiation Programme. Dr Andina Faragher, Biotext, Australia, was responsible for the scientific editing of the document.

Many individuals from various countries contributed to the development of the GDWQ. The efforts of all who contributed to the preparation of this document are greatly appreciated.

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Acronyms and abbreviations

BMD benchmark dose

BMDL lower 95% confidence limit of the benchmark dose

BMDL01 lower 95% confidence limit on the benchmark dose for a 1% response

bw body weight

CAS Chemical Abstracts Service

CH chloral hydrate

CI confidence interval

CNS central nervous system

CYP cytochrome P450

DCA dichloroacetic acid

DCVC S-dichlorovinyl-L-cysteine DCVG S-dichlorovinyl glutathione DNA deoxyribonucleic acid

FAO Food and Agriculture Organization of the United Nations GAC granular activated carbon

GD gestation day

GDWQ Guidelines for drinking-water quality

GSH glutathione

GST glutathione-S-transferase

GV guideline value

HED human equivalent dose

Leq litre-equivalent

LOAEL lowest-observed-adverse-effect level NOAEL no-observed-adverse-effect level

OR odds ratio

PBPK physiologically based pharmacokinetic (modelling) PCE perchloroethylene (tetrachloroethene)

POD point of departure

PPAR peroxisome proliferator activated receptor

RR relative risk

TCA trichloroacetic acid

TCE trichloroethene

TCOG trichloroethanol glucuronide TCOH trichloroethanol

TDI tolerable daily intake USA United States of America

US EPA United States Environmental Protection Agency

VHL Von Hippel–Lindau

VOC volatile organic compound WHO World Health Organization

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Executive summary

Trichloroethene (TCE) is primarily, if not exclusively, a groundwater contaminant, because it volatilizes to the atmosphere from surface waters. The primary cause of groundwater contamination is poor handling and disposal practices, which result in soil contamination; in vulnerable aquifers, soil contamination can result in groundwater contamination. Commercial utility of TCE is decreasing as a result of increasing regulations.

TCE is a data-rich compound, with many high-quality studies available on kinetics and toxicity in humans and laboratory animals. These include studies on TCE-induced neurological effects in humans and animals; effects on kidney, liver and body weight in animals; immunological effects in animals; reproductive effects in humans and animals; and developmental effects in animals.

Selection of multiple critical effects, rather than the lowest point of departure, as the basis of the guideline value (GV) of 8 g/L helped overcome possible limitations of individual studies.

The GV is achievable using currently available treatment technologies. Source control measures should include improved handling and disposal practices.

TCE monitoring requirements in drinking-water regulations and standards should be limited to groundwater sources where a catchment risk assessment indicates the possibility of presence of TCE. Source control should be the primary mitigation measure; however, this is not feasible where there is historical contamination. Effective treatment techniques include aeration, including packed tower aeration, and granular activated carbon. Ozone and advanced oxidation processes with ozone may also be effective. Surface water sources do not need to be monitored or treated, since TCE volatilizes to the atmosphere.

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1 General description

1.1 Identity

Trichloroethene (TCE) is also known as trichloroethylene, acetylene trichloride, 1-chloro-2,2- dichloroethylene, 1,1-dichloro-2-chloroethylene, ethylene trichloride or 1,1,2- trichloroethylene.

CAS No.: 79-01-6

Molecular formula: C2HCl3

Chemical structure:

1.2 Physicochemical properties

Table 1.1. Physicochemical properties of trichloroethene

Property Value

Molecular weight 131.39

Boiling point 87.2 °C

Melting point –84.7 °C

Density at 20 °C 1.4642 g/cm³

Vapour density (air = 1) 4.53

Vapour pressure at 25 °C 69 mm Hg

Solubility

Water at 25 °C 1280 mg/L

Organic solvents Soluble in ethanol, diethyl ether, acetone and chloroform

Partition coefficients

Log Kow 2.61

Log Koc 49–460

Henry’s law constant at 25 °C 9.85 × 10^–3 atm-m³/mol Note: Conversion factors: 1 ppm = 5.46 mg/m³; 1 mg/m³ = 0.18 ppm (ATSDR, 2019) Source: ATSDR (2019)

1.3 Organoleptic properties

TCE is a liquid with a sweet ether-like and chloroform-like odour. The odour thresholds for TCE are 546–1092 mg/m³ in air and 0.31 mg/L in water (Amoore & Hautala, 1983; Ruth, 1986).

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1.4 Major uses and sources

TCE is used primarily in metal degreasing operations. It is also used as a solvent for greases, oils, fats and tars; in paint removers, coatings and vinyl resins; and by the textile processing industry to scour cotton, wool and other fabrics. Historically, the most important use of TCE has been vapour degreasing of metal parts in the automotive and metals industries. This use has been declining since the 1990s, as a result of increased environmental regulations governing TCE emissions (ATSDR, 2019). For example, the use of TCE as a solvent in Europe dropped by 85% from 1984 to 2006, and by a further 60% from 2006 to 2010 (ECSA, 2012;

IARC, 2014).

Currently, the main use of TCE is as a feedstock material to produce other chemicals, such as fluorinated hydrocarbons and fluorinated polymers, which are being phased out under the Montreal Protocol on Substances that Deplete the Ozone Layer. About 80% of current production in the European Union is used for this purpose (ECSA, 2012). TCE may be used as a chemical intermediate in the production of polyvinyl chloride, flame-retardant chemicals and insecticides.

Most of the TCE used for degreasing is believed to be emitted to the atmosphere (US EPA, 1985a). TCE may also be introduced into surface water and groundwater in industrial effluents (IPCS, 1985). Poor handling, and improper disposal of TCE in landfills, have been the main causes of groundwater contamination. Biodegradation of another volatile organic pollutant, tetrachloroethene (also called perchloroethylene, PCE), in groundwater may also lead to the formation of TCE (Major, Hodgins & Butler, 1991).

2 Environmental levels and human exposure

TCE is widely distributed in the environment as a result of industrial emissions.

Potential environmental exposure to TCE in the air, rainwater, surface waters and drinking- water has been reviewed (IARC, 2014; ATSDR, 2019). The partitioning tendency of TCE in the environment has been estimated as follows: air, 97.7%; water, 0.3%; soil, 0.004%; sediment, 0.004% (Boutonnet et al., 1998).

TCE in the atmosphere is highly reactive and persists for an estimated half-life of 6.8 days. It is transformed in the atmosphere by reaction with photochemically produced hydroxyl radicals (ATSDR, 2019).

In surface water, volatilization is the principal route of degradation; photodegradation and hydrolysis play minor roles. In groundwater, TCE is degraded slowly by microorganisms.

2.1 Water

TCE has been detected frequently in natural water and drinking-water in various countries (IARC, 2014). Because of the high volatility of TCE, it is normally present at low or undetectable concentrations in surface water (≤1 μg/L; Health Canada, 2005). A 2000 survey of 68 First Nations community water supplies (groundwater and surface water) in Manitoba, Canada, found that TCE concentrations were undetectable (<0.5 μg/L) (Yuen & Zimmer, 2001).

The United States Environmental Protection Agency (US EPA, 2014) noted that very low concentrations of TCE are anticipated in surface water, based on TCE releases to water and wastewater treatment reported to the Toxics Release Inventory, as well as the fate of TCE in wastewater treatment.

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The major route of removal of TCE from water is volatilization. The estimated volatilization half-life is 1.2 hours from a model river (1 m deep, flowing at 1 m/second, with a wind velocity of 5 m/second) and 4.6 days from a model lake (1 m deep, flowing at 0.05 m/second, with a wind velocity of 0.5 m/second) (US EPA, 2010). However, in groundwater systems where volatilization and biodegradation are limited, concentrations are higher if contamination has occurred in the vicinity and leaching has taken place.

Data from Canada – New Brunswick (1994–2001), Alberta (1998–2001), Yukon (2002), Ontario (1996–2001) and Quebec (1985–2002) – for raw water (surface water and groundwater), and for treated and distributed water indicated that more than 99% of samples contained TCE at concentrations ≤1.0 μg/L. Most samples with detectable TCE concentrations were from groundwater, with the highest concentration being 81 μg/L (Alberta Department of Environmental Protection, New Brunswick Department of Health and Wellness, Ontario Ministry of Environment and Energy, Yukon Department of Health and Social Services and Quebec Ministry of the Environment, personal communications, 2002). In England and Wales, for about 5000 raw water (surface water and groundwater) samples taken in 2017 from about 800 abstraction points, the mean concentration of TCE was 0.55 µg/L and the maximum was 17.4 µg/L (P. Marsden, UK Drinking Water Inspectorate, personal communication, 28 August 2018). Also in England and Wales, for more than 11 000 drinking-water samples analysed in 2003, the mean concentration of TCE was 0.39 µg/L and the maximum was 21.8 µg/L (P.

Marsden, personal communication, 28 Aug 2018).

Contamination of drinking-water supplies with TCE varies with location and with the drinking- water source:

• Contamination is more likely in locations with relevant industrial activities, and improper handling and disposal.

• Generally higher levels of TCE are expected in groundwater because of the lack of volatilization that occurs compared with surface water.

Because analytical methods have improved since TCE was first assayed, concentrations that were once considered “nondetectable” are now quantifiable. This confounds the use of historical TCE data, because the values for “nondetectable” have changed over time. Since the use of TCE continues to decrease, more recent data on the concentration of TCE in drinking- water is required to provide an accurate assessment of human exposure to TCE via drinking- water and its contribution to the total body burden.

2.2 Food

The daily intakes of TCE in food for Canadian adults (20–70 years old) and children (5–

11 years old) were estimated to range from 0.004 to 0.01 μg/kg body weight (bw)/day and from 0.01 to 0.04 μg/kg bw/day, respectively (Canadian Department of National Health and Welfare, 1993). These numbers were based on TCE concentrations from food surveys in the United States of America from the mid- to late 1980s, as well as Canadian food consumption data. In recent decades, the severe restrictions on the use of TCE in North America and Europe suggest that levels in food have been decreasing.

As part of the Total Diet Study in the USA, TCE was found in 30 out of 70 (43%) food items purchased in supermarkets or restaurants in 1996–2000 at concentrations in the low microgram-per-kilogram range (Fleming-Jones & Smith, 2003). Food was sampled four times per year on a regional basis over a 5-year period. Based on 20 samples of each food item, PCE

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was most frequently detected in raw avocado (n = 6; 2–75 μg/kg). Potato chips (n = 4; 4–

140 μg/kg) had the highest level of TCE detection, and beef frankfurters (n = 5; 2–105 μg/kg) had the second most frequent and second highest level of detection. Potential sources of the contamination were not investigated (Fleming-Jones & Smith, 2003).

Among 17 samples of brown grease from grease traps in food preparation facilities, TCE was detected in three of the samples, with a mean TCE concentration of 321.3 μg/L (range 146–

600 μg/L (Ward, 2012).

2.3 Air

TCE has been detected worldwide in outdoor and indoor air. In the USA, the results of 1200 measurements in 25 states suggest a general downward trend in mean concentrations of TCE in air, from about 1.5 μg/m³ in the late 1980s to 0.8 μg/m³ in the late 1990s (IARC, 2014).

TCE concentrations in air have continued to decrease steadily in the USA: an analysis by McCarthy et al. (2007) of Air Quality System data over three trend periods (1990–2005, 1995–

2005, and 2000–2005) suggested a decrease of about 4–7% for median trichloroethylene levels annually. Data available on ambient air measurements obtained from EPA’s Air Quality System database, as reported by ATSDR (2019), indicate that, during the period 2010–2018, annual mean 50th percentile TCE airborne concentrations from various sampling sites across the USA ranged from 0 to 0.021 μg/m³. The mean 95th percentile TCE airborne concentrations across all sampling sites in 2002 and 2018 were 0.25 μg/m³ and 0.0128 μg/m³, respectively (ATSDR, 2019).

Data on TCE concentrations in air measured in different remote, rural, suburban and urban sites indicate a similar decreasing trend (IARC, 2014). Concentrations in urban air and in commercial/industrial areas were about three times higher than in rural areas (Wu & Schaum, 2000).

Modelling suggests that concentrations of TCE in indoor air can increase when TCE- contaminated water is used domestically – for example, during showering (Ömür-Özbek, Gallagher & Dietrich, 2011).

Brenner (2010) measured median and maximum TCE concentrations of 0.895 and 1.69 μg/m³ (0.16 and 0.31 ppb), respectively, for 541 indoor air samples from four large buildings at the southern end of San Francisco Bay. The levels were attributed to vapour intrusion from underlying contaminated groundwater and soil (US EPA, 2011c; Burk & Zarus, 2013).

2.4 Bioaccumulation

Bioconcentration of TCE in aquatic species is low, with bioconcentration factor values ranging between 3 and 100 in aquatic organisms (ATSDR, 2019) and some plants (Schroll et al., 1994).

2.5 Occupational exposure

The great majority of data regarding worker exposure to TCE were obtained from degreasing operations, which is the primary industrial use of TCE (ATSDR, 2019).

Worker exposure is likely to vary, although in most workplaces TCE concentration is regulated by time-weighted averages (TWA). The United States Occupational Safety and Health Administration allows an 8-hour TWA permissible exposure limit of 100 ppm and a 15-minute TWA exposure of 300 ppm, which should not be exceeded at any time during a work day (OSHA, 1993; Rosa, 2003).

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Worker exposure in the dry-cleaning industry is a notable route for exposure to TCE. This is generally evaluated using the relationship between concentrations of TCE in urine and concentrations in air collected in the breathing zone of workers in the workplace. In one study comparing exposed and non-exposed workers in a dry-cleaning centre, the mean values for exposure to TCE in the breathing zone were 1.56, 1.75 and 2.40 mg/m³ (0.28, 0.32 and 0.43 ppm based on the conversation factor in table 1) for sites with dry-cleaning machine capacities of 8, 12 and 18 kg, respectively. The mean value for exposure to TCE in the breathing zone for the occupationally non-exposed participants was 0.98 mg/m³ (0.18 ppm).

Mean urinary concentrations before and after work shifts were measured. Levels before work were 2.38, 5.53 and 8.18 μg/L (ppb), and levels after work were 4.46, 11.31 and 4.46 μg/L (ppb) at sites with dry-cleaning machine capacities of 8, 12 and 18 kg, respectively. For occupationally non-exposed participants, levels were 0.31 μg/L (ppb) before work and 0.29 μg/L (ppb) after work (Rastkari, Yunesian & Ahmadkhaniha, 2011).

2.6 Estimated total exposure, biomonitoring studies and relative contribution of drinking-water

Most people are exposed to TCE through drinking-water or air. Exposure is likely to have decreased in North America and Europe as a result of restricted use of TCE in the past several decades in these regions. TCE has been detected in human body fluids such as blood (Brugnone et al., 1994; Skender et al., 1994) and breast milk (Pellizzari, Hartwell & Harris, 1982). Several studies have examined blood concentrations of TCE in the general population. The number of individuals with measurable concentrations of TCE is generally low and has declined in recent years (IARC, 2014).

In the United States National Health and Nutrition Examination Survey 1999–2000, blood samples were taken from 290 subjects; 88% of samples were below the limit of detection of TCE, and the mean TCE concentration in the positive samples was 0.013 μg/L. In an update of this survey, for 2001–2014, blood concentrations were usually below the limit of detection of 0.012 ng/mL for 17,419 subjects from the USA general population, including different ethnic groups and age groups. The most recent data reported include results from 923 cigarette smokers and 2,054 nonsmokers within the USA general population surveyed during 2013 and 2014: again, the levels were below the limit of detection (ATSDR, 2019).

Exposure of the general population from air, water and food was several orders of magnitude lower than occupational exposure.

As a result of the volatility and lipid solubility of TCE, exposure can also occur dermally and through inhalation, especially through bathing and showering (Krishnan & Carrier, 2008).

These indirect exposures are evaluated in terms of litre-equivalents per day (Leq/day). For example, an inhalation exposure of 1.7 Leq/day means that the daily exposure to TCE via inhalation is equivalent to a person drinking an extra 1.7 L of water per day. The use of Leq as a metric of exposure is the most appropriate approach for systemically acting contaminants that do not exhibit portal-of-entry effects but are likely to induce the same adverse effect by various exposure routes (Krishnan & Carrier, 2008).

McKone (1987) has estimated that the indoor-air exposure attributable to tap water is 1.5–

6 times the exposure attributable to the consumption of 2 L/day of tap water. Bogen et al. (1988) proposed lifetime Leq/day values for 70 kg adults of 2.2 (ingestion), 2.9 (inhalation) and 2 (dermal). Weisel & Jo (1996) have reported that approximately equivalent amounts can enter the body by inhalation, dermal absorption and ingestion. Lindstrom & Pleil (1996) calculated,

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using a TCE concentration of 4.4 μg/L in water, that the ingested dose was more important than the inhaled dose for a 10-minute shower, which, in turn, was greater than the dermal dose.

Krishnan (2003) determined Leq/day values for dermal and inhalation exposures of adults and children to TCE (5 μg/L) in drinking-water on the basis of the methodological approach of Lindstrom & Pleil (1996), the use of physiologically based pharmacokinetic models and consideration of the fraction absorbed (Laparé, Tardif & Brodeur, 1995; Lindstrom & Pleil, 1996; Poet et al., 2000). Bathing in water for 30 minutes resulted in 2.39 Leq (1.67 inhalation and 0.72 dermal) for adults. In Japan, the median indoor-air exposure to TCE attributable to tap water for a Japanese lifestyle was estimated to be 3.1 Leq/day (Akiyama et al., 2018). Thus, the indirect exposure rates depend on exposure scenarios, such as the duration and frequency of showering. These scenarios are associated with local lifestyles. Since most people do not take a daily 30-minute bath, the values here are considered to be conservative. Overall, indirect exposure attributable to tap water may equal direct exposure from water intake. However, estimates could be improved by considering bioavailability, target tissue dose, and extent of absorption via all routes and media (Krishnan & Carrier, 2013).

3 Toxicokinetics and metabolism in animals and humans

3.1 Absorption

TCE is readily absorbed following both oral and inhalation exposure. Dermal absorption is also possible, but information on this route of exposure is limited. Significant variability between and within species in TCE absorption following all routes of exposure has been well documented.

In animals, TCE is rapidly and extensively absorbed from the gastrointestinal tract into the systemic circulation. Mass balance studies using radiolabelled TCE indicated that mice and rats metabolized TCE at 38–100% and 15–100%, respectively, following oral administration in corn oil vehicle. For both species, the lower values were obtained following treatment with large doses, in excess of 1000 mg/kg bw, implying that the rate of absorption was higher at low doses than at high doses (Daniel, 1963; Parchman & Magee, 1982; Dekant & Henschler, 1983;

Dekant, Metzler & Henschler, 1984; Buben & O’Flaherty, 1985; Mitoma et al., 1985; Prout, Provan & Green, 1985; Rouisse & Chakrabarti, 1986). Different vehicles affect the rate of absorption: the rate is almost 15 times greater following dosing in water than following dosing in corn oil. Overall, absorption of TCE through the gastrointestinal tract is considerable and, at very low concentration, nearly complete.

Although human exposure studies investigating oral absorption of TCE were not identified, numerous case studies of accidental or intentional ingestion of TCE suggest that absorption from the gastrointestinal tract in humans is likely to be extensive (Kleinfeld & Tabershaw, 1954; DeFalque, 1961; Brüning et al., 1998). Following ingestion accidents, TCE and its metabolites were reported in blood and/or urine at the first sampling times after exposure, the earliest of which was 13 hours, with peak amounts in blood within the first 24 hours (Brüning et al., 1998; Perbellini et al., 1991; Yoshida et al., 1996).

Pulmonary uptake of TCE into the systemic circulation is rapid in animals, after both administration through the nose only and exposure of the whole body to TCE vapour (IARC, 2014). Blood:gas partition coefficients in rodents vary between species, strains and sexes (Lash, Parker & Scott, 2000). After inhalation exposure to radiolabelled TCE at 54 or 3200 mg/m³ over a 6-hour period, net pulmonary uptake was 10 times greater at the higher concentration than at the lower concentration in rats, whereas it was similar at both exposure concentrations

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in mice (Stott, Quast & Watanabe, 1982). In humans, TCE is rapidly and extensively absorbed by the lungs and into the alveolar capillaries. The blood:air partition coefficient of TCE has been estimated to be approximately 1.5- to 2.5-fold lower in humans than in rodents (Sato et al., 1977; Monster, 1979; Clewell et al., 1995, Simmons et al., 2002; Mahle et al., 2007). TCE pulmonary uptake is rapid during the first 30–60 minutes of exposure, and decreases significantly as TCE concentrations in tissues approach steady state (Fernandez et al., 1977;

Monster, Boersma & Duba, 1979).

Dermal absorption has been demonstrated in mice (Tsuruta, 1978) and guinea-pigs (Jakobson et al., 1982). Dermal absorption has also been demonstrated in human volunteers (Stewart &

Dodd, 1964; Sato & Nakajima, 1978), with variability in absorption rates between individuals (Kezic et al., 2000).

3.2 Distribution

Once absorbed, TCE diffuses readily across biological membranes, and is widely distributed to tissues and organs via the circulatory system. Studies in animals (e.g. Fernandez et al., 1977;

Dallas et al., 1991; Fisher et al., 1991) and humans (De Baere et al., 1997) have found TCE or its metabolites in most major organs and tissues. Primary sites of distribution include the lungs, liver, kidneys and central nervous system (CNS). TCE may accumulate in adipose tissue because of its lipid solubility. in humans, reported tissue:blood partition coefficients were highest for fat (52–64); the range for all other tissues and organs is much lower, at 0.5–6.0 (IARC, 2014). Slow release of TCE from adipose stores might act as an internal source of exposure, ultimately resulting in longer mean residence times and bioavailability of TCE (Fernandez et al., 1977; Dallas et al., 1991; Fisher et al., 1991).

Age-dependent factors may influence TCE distribution in humans (Pastino, Yap & Carroquino, 2000).

3.3 Metabolism

Adverse health effects of TCE are attributed to some of its metabolites (except for solvent effects that occur at extremely high exposures to the parent compound).

TCE metabolism is quite complex, yielding multiple intermediates and end products (IARC, 2014; Lash et al., 2014; ATSDR, 2019). Experimental animal and human data indicate that TCE metabolism occurs through two major pathways: cytochrome P450 (CYP)-dependent oxidation and glutathione (GSH) conjugation catalysed by glutathione S-transferases (GSTs).

Flux through the CYP-dependent oxidation pathway far exceeds that through the GSH conjugation pathway in all species studied, including humans. Metabolites generated by the CYP-dependent oxidation pathway are mostly chemically stable. In contrast, the GSH conjugation pathway generates several highly reactive metabolites. Chemical stability of the metabolite is an important determinant of systemic availability and fate. Relatively stable TCE metabolites may be transported from their site of formation into the bloodstream and delivered to other potential target organs.

TCE metabolism by the oxidative pathway occurs mainly in the liver. Other tissues that are sites of CYP-mediated TCE metabolism include the lungs, kidneys and male reproductive organs. Different isozymes of cytochrome P450 oxidize TCE; the highest contribution is by CYP2E1 (Lash et al., 2014). In the CYP-dependent oxidation pathway, TCE is metabolized to an epoxide intermediate (TCE epoxide), which spontaneously rearranges to chloral

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(trichloroacetaldehyde, trichloroethanal). Chloral is further metabolized to trichloroethanol (TCOH), trichloroethanol glucuronide (TCOG) and trichloroacetic acid (TCA) as the principal metabolites. Under certain conditions, TCE epoxide forms dichloroacetyl chloride, which rearranges to dichloroacetic acid (DCA) (Goeptar et al., 1995). DCA is then further metabolized by GST. This occurs at a higher rate than metabolism of TCA and TCE; as a result, measurable DCA concentrations are not often generated in vivo.

TCE metabolism by the GST-catalysed GSH conjugation pathway occurs more slowly than metabolism by the CYP-catalysed pathway. The initial GSH conjugation step occurs primarily (but not exclusively) as first-pass metabolism in the liver, which has a high content of GSTs.

The liver is very efficient at excreting GSH conjugates as the S-dichlorovinyl glutathione (1,1- DCVG and 1,2-DCVG) into either bile or plasma. Subsequently, through enterohepatic and renal–hepatic circulation, S-dichlorovinyl-L-cysteine (DCVC) or the mercapturate N-acetyl-S- dichlorovinyl-L-cysteine are delivered to the kidneys for further metabolism or excretion.

DCVG may undergo N-acetylation and be excreted in the urine or metabolized by a lyase enzyme to reactive metabolites, including a thioacetaldehyde and a thioketene (Clewell et al., 2001). Additionally, in situ GSH conjugation of TCE can occur within the kidneys themselves, primarily the proximal tubules, establishing an intra-organ cycle of GSH conjugate transport and metabolism (Lash et al., 2014).

Exposure to TCE clearly results in exposure of tissues to a complex mixture of metabolites (US EPA, 2011b).

The very high clearance of TCE seen at low oral doses, which is associated with first-pass metabolism in the liver, essentially favours the oxidative metabolism pathway. This is the main reason that the GSH conjugation pathway does not seem to contribute much to the clearance of TCE at low doses. In addition, the enterohepatic circulation of TCOG is believed to play a very important role in maintaining TCA levels, and therefore has a major impact on the oxidative metabolite dosimetry (Stenner et al., 1997, 1998; Barton et al., 1999). The oxidative metabolites are clearly responsible for the effects on the liver (both cancer and noncancer; see sections 4 and 5). This implies that the oral route is most important for liver effects, whereas other routes of exposure may preferentially affect other organs (e.g. kidney).

There are several interspecies differences in TCE metabolism. For example, human hepatic microsomes have less activity towards TCE than rat or mouse hepatic microsomes (Nakajima et al., 1993), and humans are less efficient at metabolizing TCE than rodents. Furthermore, a comparison of renal β-lyase activities in the kidney indicates that rats are more efficient than humans at metabolizing DCVC to reactive metabolites (Clewell et al., 2000).

TCE metabolism also differs within species. In humans, variations between individuals have been reported in enzyme expression and activity – for example, in the activity of CYP2E1 and GST. These reflect differences between the sexes, pathological status, genetic polymorphisms, or induction and inhibition of the enzymes (Lash, Parker & Scott, 2000; Lash et al., 2014). For example, chronic exposure to ethanol, a CYP2E1 inducer, is expected to increase TCE metabolism (Lash et al., 2014). In addition, genetic polymorphisms of OAT1 and OAT3 have been reported to result in different capacity to accumulate DCVG or DCVC (Lash, Putt &

Parker, 2006), which is likely to affect nephrotoxicity.

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3.4 Elimination

The database pertaining to the elimination of TCE is large, and TCE clearance is well characterized in both animals and humans. In humans, it has been estimated that during and up to 5 days after a 4-hour inhalation exposure period, pulmonary excretion accounts for 19–35%

of TCE intake, urinary excretion of metabolites accounts for 24–39% of TCE intake, and the balance is retained in the body (Monster, Boersma & Duba, 1976; Opdam, 1989; Chiu et al., 2007). Although the elimination kinetics of TCE and its metabolites vary by route of exposure, elimination pathways appear to be similar for ingestion and inhalation. The half-life of TCE in alveolar air has been estimated as about 6–44 hours. Half-lives of trichloroethanol and TCA in urine are 15–50 and 36–73 hours, respectively (IARC, 2014). No data were found regarding elimination of TCE and its metabolites following dermal exposure.

TCE is eliminated either unchanged in expired air or as metabolites, primarily in urine. The excreted metabolites are TCA, TCOH or TCOG (following oxidative metabolism), or DCVG or the cysteine conjugate N-acetyl-S-dichlorovinyl-L-cysteine (following GSH conjugation).

Studies in human volunteers have shown that urinary TCOH is first produced more quickly and in larger amounts than urinary TCA. However, over time, TCA production eventually exceeds that of TCOH. Small amounts of metabolized TCE are excreted in the bile or as TCOH in exhaled air. The total radioactivity recovered in mouse and rat faeces after oral exposure to radiolabelled TCE accounted for about 1–5% of total radiolabel administered (Dekant, Metzler

& Henschler, 1984; Kim & Ghanayem, 2006), although higher values (up to 24%) were also reported (Green & Prout, 1985) in another strain of mice. TCE may also be excreted in breast milk (Pellizzari, Hartwell & Harris, 1982; Fisher et al., 1987; Fisher, Whittaker & Taylor, 1989).

Elimination is more rapid in mice than in rats (Lash, Parker & Scott, 2000), but formation of TCA is approximately 10 times faster in mice than in rats. These observations help explain interspecies differences in toxicity associated with TCE, given that the toxicity of TCE is linked to the formation of its metabolites (Parchman & Magee, 1982; Stott, Quast & Watanabe, 1982;

Dekant, Metzler & Henschler, 1984; Buben & O’Flaherty, 1985; Mitoma et al., 1985; Prout, Provan & Green, 1985; Rouisse & Chakrabarti, 1986). In humans, differences between individuals have been seen in the metabolism and elimination of TCE (Nomiyama &

Nomiyama, 1971; Fernandez et al., 1975; Monster, Boersma & Duba, 1976).

3.5 Physiologically based pharmacokinetic modelling

Toxicity studies have been conducted for the inhalation route in humans (occupationally exposed individuals) and in experimental animals. In contrast, the database on TCE ingestion via drinking-water is limited. Therefore, many targets of toxicity from chronic exposure to TCE largely focus on the inhalation route of exposure.

Considering the main features of TCE kinetics, as summarized above, a linear extrapolation from high-dose studies in rodents to low-dose human exposures seems not be appropriate, for the following reasons:

• TCE is rapidly and well absorbed by both the oral and inhalation routes of exposure (ATSDR, 2019).

• The metabolic pathways and kinetics of excretion for oral exposure are similar to those for inhalation exposure (ATSDR, 2019).

• Data for oral exposure indicate a pattern of effects similar to that of inhalation exposure.

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• Differences in first-pass effects (affecting systemic bioavailability) between oral and inhalation exposures can be adequately accounted for by a physiologically based pharmacokinetic (PBPK) model.

• Quantitative differences in TCE metabolism between humans and rodents exist.

• Metabolite production is not linear because the oxidative pathway is saturated at the high doses at which the GST pathway starts to be active.

The use of PBPK modelling allows a route-to-route extrapolation, as well as estimation of the internal exposure.

Several PBPK models of TCE have been developed, progressively increasing in complexity to address specific problems in extrapolation of kinetics from rats and mice to humans (Fisher, 2000; Poet et al., 2000; Thrall & Poet, 2000; Simmons et al., 2002; Keys et al., 2003; Hack et al., 2006; Chiu, Okino & Evans, 2009; Evans et al., 2009; US EPA, 2011b). Recent models include the kinetics of the relevant TCE oxidative metabolites (chloral hydrate [CH], TCA, TCOH and trichloroethanol-glucuronide conjugate) (Fisher, 2000; Hack et al., 2006), as well as the metabolites formed via GSH conjugation in the liver or kidney leading to the appearance of DCVC (Clewell et al., 2000). The US EPA has derived its own model, based on previous models but incorporating newer data (Chiu, Okino & Evans, 2009; Evans et al., 2009), and has applied the updated model to dosimetry extrapolations to support its toxicological review of TCE (US EPA, 2011b). The model features have been extensively described (ATSDR, 2019).

Starting from the lowest-observed-adverse-effect level (LOAEL) and no-observed-adverse- effect level (NOAEL) or benchmark dose (BMD) values, PBPK modelling was used to apply a route-to-route extrapolation and calculate an internal dose based on present understanding of the role that different TCE metabolites play and the mode of action for TCE toxicity (US EPA, 2011b). The PBPK model was also used to estimate interspecies and intraspecies pharmacokinetic variability. This resulted in 99th percentile estimates of human equivalent dose (HED99) for the critical effects.

The PBPK model simulated 100 weeks of human exposure. This was considered representative of continuous lifetime exposure because longer simulations did not add substantially to the average (e.g. doubling the simulated exposure time resulted in a change in the resulting HED of less than a few percent).

4 Effects on humans

4.1 Acute exposure

CNS effects were the primary effects noted from acute inhalation exposure to TCE in humans.

Symptoms included sleepiness, fatigue, headache, confusion and feelings of euphoria (ATSDR, 2019). Simultaneous exposure to TCE and ethanol results in a marked inhibition of the metabolism of TCE, which leads to accumulation of TCE in the blood and increases the extent of CNS depression (Muller, Spassovski & Henschler, 1975). Effects on the liver, kidneys, gastrointestinal system and skin have also been noted (ATSDR, 2019). In its past use as an inhalant anaesthetic drug in humans, concentrated solutions of TCE have proved quite irritating to the gastrointestinal tract, and have caused nausea and vomiting (DeFalque, 1961).

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4.2 Short-term exposure

Information from medium-term (to long-term) TCE exposures via inhalation and the dermal route has been reviewed (ATSDR, 2019). These studies indicated that the CNS is the most sensitive organ for toxicity, followed by the liver and kidneys. Case reports of intermediate and chronic occupational exposures included effects such as dizziness, headache, sleepiness, nausea, confusion, blurred vision, facial numbness, and weakness. Liver effects noted included liver enlargement and increases of liver enzymes in serum. Kidney effects included increased N-acetyl-β-D-glucosaminidase. Cardiovascular, immunological, reproductive and carcinogenic effects were also observed (ATSDR, 2019).

4.3 Long-term exposure 4.3.1 Systemic effects

The systemic effects elicited by TCE are not specific to the exposure route; similar effects can be elicited via oral and inhalation routes.

There is some evidence for TCE-induced hepatic effects (e.g. changes in blood and urine indices of liver function, enlarged liver) in occupationally exposed humans. However, study limitations include lack of quantifiable exposure data and confounding due to concomitant exposure to other chemicals.

Renal toxicity was reported in occupationally exposed humans (although workers were sometimes also exposed to other chemicals in the workplace). No clear evidence of kidney effects has been reported in studies examining the association between long-term exposure to TCE in drinking-water and adverse health effects.

4.3.2 Neurological effects

Reported neurological effects, as described in section 4.2, have been associated with relatively high exposure to TCE.

4.3.3 Reproductive and developmental effects

Most epidemiological studies have found no convincing association between adverse reproductive effects in humans and exposure to TCE in contaminated drinking-water (IPCS, 1985; ATSDR, 2019). Epidemiological data are typically limited by concomitant exposure to other potentially hazardous substances, and case–control studies are limited by small numbers of cases.

Although an epidemiological study of 2000 male and female workers exposed to TCE via inhalation found no increase in infant malformations following exposure (IPCS, 1985), an association was found between the occurrence of congenital heart disease in children and a drinking-water supply contaminated with TCE and similar chemicals (IPCS, 1985). These studies were confounded by several factors, including potential exposure to many other contaminants or compounds that produce similar metabolites, a lack of characterization of the exposure levels and the exposed populations, and failure to characterize the nature of the

“congenital heart disease” (which may not necessarily be equivalent to cardiac anomalies).

Therefore, use of these studies to indicate a causal association between TCE and congenital cardiac anomalies remains very limited.

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Other epidemiological studies of women exposed to degreasing solvents, including TCE, have reported elevated risks of cardiac anomalies in their offspring (Goldberg et al., 1990; Ferencz, Loffredo & Correa-Villaseñor, 1997; Wilson et al., 1998). Large, statistically significant excesses were observed for specific cardiac defects: left-sided obstructive defects (odds ratio [OR] = 6.0; 95% confidence interval [CI] = 1.7–21.3) and hypoplastic left heart (OR = 3.4; 95%

CI = 1.6–6.9), with an attributable risk¹ of 4.6% (Wilson et al., 1998). Neural tube defects have also been noted with either occupational or drinking-water exposure to solvents, including TCE (Holmberg & Nurminen, 1980; Holmberg et al., 1982; Bove, Fulcomer & Klotz, 1995).

In a study in which semen parameters of workers exposed to TCE were evaluated (Chia et al., 1996), sperm density showed a significant difference between low- and high-exposure subjects.

In a recent study involving a small number of subjects, TCE and its metabolites were identified in seminal fluids of workers exposed to TCE (Forkert et al., 2003), suggesting that TCE may play a role in the observed effects on sperm parameters.

Overall, epidemiological studies are plagued by lack of clarity on background coexposure. For example, in the Wilson et al. (1998) study, the investigators asked subjects about their exposure to “solvents/de-greasing compounds” but not specifically TCE. Subjects at airforce bases are exposed to jet fuels as well as other solvents on a daily basis (Stewart, Lee & Marano, 1991), but it is unlikely that they know the exact compounds contained in the degreasing compounds or solvents. This means that, based on currently available human studies, TCE cannot be specifically implicated; however, these studies can be used as supporting evidence, complementary to developmental and reproductive effects reported in animal studies.

4.3.4 Immunological effects

Studies in humans reported some associations between occupational exposure to TCE and immunotoxicological end-points. In workers, onset of scleroderma (a systemic autoimmune disease) has been reported, although a meta-analysis indicated that the available data did not allow clear conclusions, because of the very low incidence of systemic sclerosis (IARC, 2014).

Some changes in levels of inflammatory cytokines were reported in degreasers using TCE, as well as case reports of hypersensitivity skin disorder (IARC, 2014).

4.3.5 Genotoxicity and carcinogenicity

Studies examining TCE-induced genotoxicity in humans have been largely inconclusive. Four studies using peripheral lymphocyte cultures from exposed workers showed no, or only minor, effects on frequency of sister chromatid exchange (Gu et al., 1981a, b; Nagaya, Ishikawa &

Hata, 1989; Brandom et al., 1990; Seiji et al., 1990). As reviewed by IARC (2014), no further studies of genotoxicity of TCE in humans have been published.

The carcinogenicity of TCE has been investigated in several types of epidemiological studies, including cohort and case–control studies in occupationally exposed workers and in the general population exposed via different routes (inhalation, oral and dermal), in addition to ecological studies of environmental exposures.

1 Attributable risk is the risk or rate difference that may be attributable to the exposure (Rothman, 1986).

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The focus has mainly been on tumours of the kidney and liver, and non-Hodgkin lymphoma.

A clear association between any specific type of cancer and exposure to TCE has not been consistently observed in these studies. Cancer occurrence in populations exposed to drinking- water contaminated with various concentrations of TCE has been examined in several studies, but the interpretation of these studies is complicated by methodological problems.

The evidence for TCE-induced cancers in humans has been reviewed in depth by IARC (2014) and Rusyn et al. (2014). Three cohort studies were available. Two of these studies, in Sweden and Finland (Axelson et al., 1994; Anttila et al., 1995), involved people who had been monitored for exposure to TCE by measurement of TCA in urine.

The third study, in the USA (Spirtas et al., 1991), covered 14,444 workers (10 730 men and 3725 women) exposed to TCE during maintenance of military aircraft and missiles for at least 1 year between 1952 and 1956. Radican et al. (2008) extended the follow-up of this cohort until 2000. These workers were also exposed to other solvents and chemicals, including other potential carcinogens. Personal and area samples were available for some chemicals, including TCE (Stewart, Lee & Marano, 1991). Exposure frequency and exposure patterns (intermittent and continuous) for TCE were assessed based on information on job tasks. TCE was used in degreasers until 1968, when it was replaced by 1,1,1-trichloroethane (Stewart, Lee & Marano, 1991).

In none of these three cohort studies was it possible to control for potential confounding factors, such as smoking (IARC, 2014). As of 31 December 2000, 68.1% of cohort members had died.

The Cox model hazard ratio for all cancers was 1.03 (95% CI = 0.91–1.17; 854 deaths). No significantly increased hazard ratio appeared for any specific cancer in either men or women (IARC, 2014).

Overall, an elevated risk for liver and biliary tract cancer was observed, in addition to a modestly elevated risk for non-Hodgkin lymphoma seen in cohort studies. A marginally increased risk for non-Hodgkin lymphoma was suggested to exist in areas where groundwater is contaminated with TCE (IARC, 1995, 2014).

The occurrence of renal cancer was not elevated in the cohort studies. However, a study of German workers exposed to TCE yielded five cases of renal cancer compared with none in a control comparison group (Henschler et al., 1995). This latter study, conducted on 169 workers in a cardboard factory in Germany who were exposed to TCE for at least 1 year between 1956 and 1975, claimed a causal link between kidney cancer and TCE exposure (Henschler et al., 1995). By the close of the study in 1992, 50 members of the study group had died, 16 from malignant neoplasms. In two of these 16 cases, kidney cancer was the cause of death (standardized mortality ratio = 3.28, versus local population). Five workers were diagnosed with kidney cancer: four with renal cell cancer and one with a urothelial cancer of the renal pelvis (standardized incidence ratio = 7.77; 95% CI = 2.50–18.59). After the close of the observation period, two additional kidney tumours (one renal and one urothelial) were diagnosed in the study group. For the seven cases of kidney cancer, the average exposure duration was 15.2 years (range 3–19.4 years). By the end of the study, 52 members of the control group, which consisted of 190 unexposed workers from the same plant, had died, 16 from malignant neoplasms, but none from kidney cancer. No case of kidney cancer was diagnosed in the control group. Although this study received some criticism (McLaughlin &

Blot 1997), it is hard to ignore its findings.

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A positive association between renal cancer and prolonged occupational exposure to high levels of TCE was reaffirmed in a case–control study in Germany involving 134 renal cell cancer patients and 410 controls, comprising workers from industries with and without TCE exposure (Brüning et al., 2003). When the results were adjusted for age, sex and smoking, a significant excess risk was determined for the longest-held job in industries with TCE exposure (OR = 1.80; 95% CI = 1.01–13.32). Any exposure to degreasing agents was found to be a risk factor for renal cell cancer (OR = 5.57; 95% CI = 2.33–13.32). Self-reported narcotic symptoms, an indication of peak exposures, were associated with an excess risk for renal cell cancer (OR

= 3.71; 95% CI = 1.80–7.54). However, the levels of occupational exposure in that study were very high and unlikely to be reached from environmental exposure. The prolonged exposure to high levels probably affects the metabolism of TCE, with the net production of active metabolites underlying the development of renal cell cancer in occupationally exposed industrial workers.

A more recent case–control study in Montreal, Canada (Christensen et al., 2013), included histologically confirmed cases of cancer in men (n = 3730; participation rate, 82%; for control, n = 533) occurring between 1979 and 1985 from 18 of the largest hospitals in the Montreal metropolitan area. On the basis of job history reported by study subjects, exposure was estimated for 294 substances; only about 3% of the control individuals were exposed to TCE, limiting the power of the study. A total of 177 cases of cancer of the kidney were included. For exposure to TCE, the OR was 0.9 (95% CI = 0.4–2.4) when considering any level of exposure, and 0.6 (95% CI = 0.1–2.8) for substantial exposure, after adjustment for age, income, education, ethnicity, questionnaire response and smoking.

The GST gene family encodes multifunctional enzymes that catalyse several reactions between GST and electrophilic as well as hydrophobic compounds (Raunio et al., 1995). Certain defective GST genes are known to be associated with an increased risk of different kinds of cancer. A case–control study (Brüning et al., 1997b) investigated the role of GST polymorphisms in the incidence of renal cell cancer in two occupational groups exposed to high levels of TCE. The data indicate a higher risk for development of renal cell cancer if TCE- exposed people carry either the GSTT1 or GSTM1 gene, compared with individuals lacking the enzyme. These results, which are supported by the study of Henschler et al. (1995), support the view of the mode of action of TCE-induced kidney cancer as involving metabolites derived from the GSH-dependent pathway, at least in humans. Involvement of GST-dependent metabolites is further supported by a hospital-based case–control study on TCE exposure and renal cell cancer between 1999 and 2003 in seven central and eastern European cities (Moore et al., 2010). The final study population included 1097 cases and 1476 controls, who were interviewed to collect information about exposure and other possible confounders (e.g. smoking habits). A slight increased risk of renal cell cancer was observed among subjects ever exposed to TCE (OR = 1.63; 95% CI = 1.04–2.54). No increase in risk of renal cell cancer was observed among subjects with two deleted GSTT1 alleles: the ORs were 0.93 (95% CI = 0.35–2.44) in ever-exposed subjects, 0.81 (95% CI = 0.24–2.72) in subjects with below- average exposure, and 1.16 (95% CI = 0.27–5.04) in subjects with above-average exposure intensity. The presence of at least one copy of the GSTT1 gene did not significantly affect the OR.

Mutations in the Von Hippel–Lindau (VHL) tumour suppressor gene have been associated with increased risk of renal cell carcinoma (Brüning et al., 1997a; Brauch et al., 1999). Brüning et al. (1997a) examined VHL mutation by single-stranded conformation polymorphism in 23 renal cell carcinoma patients with documented high occupational TCE exposure. All TCE-

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exposed renal cell carcinoma patients had VHL mutations; this was higher than the background frequency (33–55%) among unexposed renal cell carcinoma patients. Brauch et al. (1999), in a follow-up study in 44 TCE-exposed renal cell carcinoma patients, found that 75% of TCE- exposed patients had VHL mutations and 39% had a C to T mutation at nucleotide 454. All the C to T transitions in the control renal cell carcinoma patients were relatively rare (6% of the total incidence).

The US EPA conducted a meta-analysis of epidemiological studies, focusing on non-Hodgkin lymphoma and cancers of the kidney and liver, as part of its evaluation of the carcinogenicity of TCE (Scott & Jinot, 2011). Twenty-four studies met the inclusion criteria: two studies with a high relative risk (RR) for renal cancer (i.e. Henschler et al., 1995, and Vamvakas et al., 1998) were not included in the meta-analysis because they did not meet the inclusion criteria as a result of incomplete cohort identification or potential selection bias. Overall meta-RRs for those exposed to TCE were 1.27 (95% CI = 1.13–1.43) for cancer of the kidney, 1.29 (95% CI

= 0.07–1.56) for cancer of the liver and intrahepatic bile ducts, and 1.23 (95% CI = 1.07–1.42) for non-Hodgkin lymphoma. An adjustment technique to control for possible publication bias reduced the meta-RR for non-Hodgkin lymphoma to 1.15 (95% CI = 0.97–1.36). A meta- analysis largely overlapping with that by Scott & Jinot (2011) was conducted by Karami et al.

(2012). The meta-RR for cancer of the kidney from cohort studies was 1.41 (95% CI = 0.98–

2.05), and 1.26 (95% CI = 1.02–1.56) when the study by Henschler et al. (1995) was excluded.

The meta-RR for case–control studies was 1.55 (95% CI = 1.18–2.05), and 1.35 (95% CI = 1.17–1.57) when the study by Vamvakas et al. (1998) was excluded. The combined RR for cohort and case–control studies was 1.41 (95% CI = 1.16–1.70). IARC (2014) noted that meta- RRs were stronger when more recent publications were included; it was suggested that this might reflect improved exposure assessment and less exposure misclassification. In a meta- analysis of 18 studies (14 cohort and four case–control) of non-Hodgkin lymphoma, Mandel et al. (2006) reported meta-RRs of 2.33 (95% CI = 1.39 to 3.91) for non-Hodgkin lymphoma from studies with higher-quality exposure data, 0.84 (95% CI = 0.73–0.98) from studies with lower- quality exposure data, and 1.39 (95% CI = 0.62–3.10) from case–control studies.

In conclusion, studies in humans show consistent evidence of an association between occupational TCE exposure and kidney cancer. Associations reported for liver cancer and non- Hodgkin lymphoma, although positive, are less consistent.

5 Effects on experimental animals and in vitro test systems

Many studies of a wide range of toxic end-points using repeated oral exposures to TCE have been reviewed (WHO, 2005). Because of the poor solubility of TCE in water, few studies used water as a vehicle (Tucker et al., 1982), although some drinking-water or water gavage studies have used emulsifying agents. Many of the studies are therefore confounded by the use of corn oil as a vehicle, which has been found to alter the pharmacokinetics of TCE, and to affect lipid metabolism and other pharmacodynamic processes.

The best documented systemic effects are neurotoxicity, hepatotoxicity, nephrotoxicity and pulmonary toxicity in adult animals. Reproductive and developmental effects have also been extensively studied.

5.1 Acute exposure

Neurological, lung, kidney and heart effects have been reported in animals acutely exposed to TCE (US EPA, 2011b; IARC, 2014; ATSDR, 2019). Tests involving acute exposure of rats

Trichloroethene in drinking-water