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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
WHO/HEP/ECH/WSH/2020.10
© World Health Organization 2020
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Suggested citation. Trichloroethene in drinking-water. Background document for development of
WHO Guidelines for drinking-water quality. Geneva: World Health Organization; 2020
(WHO/HEP/ECH/WSH/2020.10). Licence: CC BY-NC-SA 3.0 IGO.
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iii
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.
iv
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.
v
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
vii
Contents
Executive summary .................................................................................................................. 1
1 General description ...................................................................................................... 2
1.1 Identity ............................................................................................................... 2
1.2 Physicochemical properties ............................................................................... 2
1.3 Organoleptic properties ...................................................................................... 2
1.4 Major uses and sources ...................................................................................... 3
2 Environmental levels and human exposure ............................................................... 3
2.1 Water .................................................................................................................. 3
2.2 Food ................................................................................................................... 4
2.3 Air ...................................................................................................................... 5
2.4 Bioaccumulation ................................................................................................ 5
2.5 Occupational exposure ....................................................................................... 5
2.6 Estimated total exposure, biomonitoring studies and relative
contribution of drinking-water ........................................................................... 6
3 Toxicokinetics and metabolism in animals and humans .......................................... 7
3.1 Absorption.......................................................................................................... 7
3.2 Distribution ........................................................................................................ 8
3.3 Metabolism ........................................................................................................ 8
3.4 Elimination ....................................................................................................... 10
3.5 Physiologically based pharmacokinetic modelling .......................................... 10
4 Effects on humans ...................................................................................................... 11
4.1 Acute exposure................................................................................................. 11
4.2 Short-term exposure ......................................................................................... 12
4.3 Long-term exposure ......................................................................................... 12
4.3.1 Systemic effects ................................................................................... 12
4.3.2 Neurological effects ............................................................................. 12
4.3.3 Reproductive and developmental effects ............................................. 12
4.3.4 Immunological effects ......................................................................... 13
4.3.5 Genotoxicity and carcinogenicity ........................................................ 13
5 Effects on experimental animals and in vitro test systems ..................................... 16
5.1 Acute exposure................................................................................................. 16
5.2 Short-term exposure ......................................................................................... 17
viii
5.3 Long-term exposure ......................................................................................... 18
5.3.1 Systemic effects ................................................................................... 18
5.3.2 Neurological effects ............................................................................. 18
5.3.3 Reproductive and developmental effects ............................................. 18
5.3.4 Immunological effects ......................................................................... 20
5.3.5 Genotoxicity and carcinogenicity ........................................................ 21
5.4 Mode of action ................................................................................................. 23
6 Overall database and quality of evidence ................................................................ 25
6.1 Summary of health effects ............................................................................... 25
6.2 Quality of evidence .......................................................................................... 27
7 Practical considerations............................................................................................. 27
7.1 Analytical methods and achievability .............................................................. 27
7.2 Source control .................................................................................................. 28
7.3 Treatment methods and performance ............................................................... 28
8 Conclusion .................................................................................................................. 29
8.1 Derivation of the guideline value ..................................................................... 29
8.1.1 Noncancer effects................................................................................. 31
8.1.2 Guideline value .................................................................................... 32
8.2 Considerations in applying the guideline value ............................................... 32
References ............................................................................................................................... 34
Trichloroethene in drinking-water
1
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.
Trichloroethene in drinking-water
2
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).
Trichloroethene in drinking-water
3
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.
Trichloroethene in drinking-water
4
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
Trichloroethene in drinking-water
5
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).
Trichloroethene in drinking-water
6
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,
Trichloroethene in drinking-water
7
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
Trichloroethene in drinking-water
8
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
Trichloroethene in drinking-water
9
(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.
Trichloroethene in drinking-water
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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.
Trichloroethene in drinking-water
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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).
Trichloroethene in drinking-water
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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.
Trichloroethene in drinking-water
13
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).
Trichloroethene in drinking-water
14
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.
Trichloroethene in drinking-water
15
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-
Trichloroethene in drinking-water
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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
17
and mice have shown TCE to have low toxicity from inhalation exposure and moderate toxicity
from oral exposure (ATSDR, 2019). The 14-day acute oral median lethal dose (LD50) values
for TCE were determined to be 2400 mg/kg bw in mice (Tucker et al., 1982) and
4920 mg/kg bw in rats (IPCS, 1985; ATSDR, 2019). The 4-hour inhalation median lethal
concentration (LC50) was calculated to be 67 600 mg/m3 in rats (Siegel et al., 1971) and
54 700 mg/m3 in mice (Fan, 1988). A review of studies of dermal exposure of TCE in rabbits
indicated that skin irritation occurs after 24 hours at 0.5 mL, and degenerative skin changes
occur within 15 minutes at 1 mL in guinea-pigs (Fan, 1988). Instillation of 0.1 mL to rabbit
eyes caused conjunctivitis and keratitis, with complete recovery within 2 weeks.
5.2 Short-term exposure
In a 13-week oral study, Fischer 344/N rats and B6C3F1 mice (10 per sex per dose) were
administered TCE in corn oil by gavage at doses of up to 1000 mg/kg bw/day in female rats,
up to 2000 mg/kg bw/day in male rats and up to 6000 mg/kg bw/day in mice of both sexes for
5 days per week (NTP, 1990). Body weights were decreased in male rats at 2000 mg/kg bw/day.
Pulmonary vasculitis involving small veins was reported in female rats at 1000 mg/kg bw/day.
Mild to moderate cytomegaly and karyomegaly of the renal tubular epithelial cells occurred in
rats at 1000 mg/kg bw/day (females) or 2000 mg/kg bw/day (males). The NOAEL in rats was
reported as 1000 mg/kg bw/day (males) and 500 mg/kg bw/day (females). Among the mice,
there were decreases in survival in both sexes and body weight gain in males at
750 mg/kg bw/day and above. In both sexes, doses of 3000 mg/kg bw/day and above were
associated with centrilobular necrosis and multifocal calcification in the liver, as well as mild
to moderate cytomegaly and karyomegaly of the renal tubular epithelial cells. A NOAEL was
set at 375 mg/kg bw/day for mice.
In drinking-water studies (Sanders et al., 1982; Tucker et al., 1982), CD-1 and ICR outbred
albino mice (140 per sex per dose) were administered TCE in a 1% solution of Emulphor in
drinking-water at doses of 0, 0.1, 1.0, 2.5 or 5.0 mg/L (equivalent to 0, 18.4, 216.7, 393 or
660 mg/kg bw/day) for 4 or 6 months. Females at 5.0 mg/L and males at and above 2.5 mg/L
consumed less water than the controls. A decrease in body weight gain in both sexes and an
increase (P < 0.05) in kidney weight in males occurred at 5.0 mg/L. In addition, at 5.0 mg/L,
there were elevated urinary protein and ketone levels in both sexes, decreases in leukocyte and
red blood cell counts in males, altered coagulation times in both sexes and shortened
prothrombin times in females. At 2.5 mg/L, enlargement of the liver, and an increase in urinary
protein and ketone levels in males were observed. Inhibition of humoral immunity, cell-
mediated immunity and bone marrow stem cell colonization was seen among females at
2.5 mg/L and above. The LOAEL was considered to be 2.5 mg/L (equivalent to 393 mg/kg
bw/day) based on decreased water consumption, enlargement of the liver, increases in urinary
protein and ketone levels in males (an indication of renal effects), and changes in
immunological parameters in females. A NOAEL of 1.0 mg/L (equivalent to
216.7 mg/kg bw/day) was determined from these studies. Several previous oral studies in
animals had not found evidence of renal toxicity in mice or rats exposed to TCE (Stott, Quast
& Watanabe, 1982).
Several studies have evaluated the toxicity of TCE to rodents following short-term inhalation
exposure. In a 14-week inhalation study, rats were exposed to TCE at 0, 270, 950 or
1800 mg/m3 for 4 hours per day, 5 days per week, for 14 weeks. Another group was exposed
to TCE cat 300 mg/m3 for 8 hours per day, 5 days per week, for 14 weeks. There were
significant increases (P < 0.01) in the absolute and relative liver weights in treated animals
compared with controls, although liver and kidney function tests of treated animals remained
Trichloroethene in drinking-water
18
within normal limits (Kimmerle & Eben, 1973). In a study in which mice, rats and gerbils
(unspecified strains) were exposed to TCE continuously by inhalation at 810 mg/m3 for 30 days,
there was a significant increase (P < 0.05) in the liver weights of all three species (Kjellstrand
et al., 1981). Renal effects of inhaled TCE have also been reported (Kjellstrand et al., 1981,
1983a, b). Male and female gerbils exposed to TCE at 810 mg/m3 continuously for 30 days had
increased (P < 0.05) kidney weight. NMRI mice exposed to TCE at 200, 410, 810 or
1600 mg/m3 continuously for 30 days had significantly increased (P < 0.05) kidney weight, at
410 mg/m3 in males and above 810 mg/m3 in females. No kidney effects were evident in the
remaining strains of mice (Kjellstrand et al., 1983a).
5.3 Long-term exposure
5.3.1 Systemic effects
Administration of high doses of TCE by gavage for long durations in rats and mice has been
associated with nephropathy, with characteristic degenerative changes in the renal tubular
epithelium (NCI, 1976). Toxic nephrosis, characterized by cytomegaly of the renal tubular
epithelium, has been reported in cancer bioassays in mice and rats (NTP, 1983, 1988, 1990).
The toxicity of TCE was investigated in F344 rats and B6C3F1 mice (50 per sex per dose)
given 0, 500 or 1000 mg/kg bw/day (rats) and 0 or 1000 mg/kg bw/day (mice) in corn oil,
5 days per week, for 103 weeks. Survival was reduced in male rats and mice but not in females
(NTP, 1983). Toxic nephrosis, characterized by cytomegaly of the renal tubular epithelium,
occurred in rats at 500 mg/kg bw/day and above, and in mice at 1000 mg/kg bw/day. LOAELs
of 500 mg/kg bw/day in rats and 1000 mg/kg bw/day in mice were defined for long-term
effects. A NOAEL was not determined (NTP, 1990).
5.3.2 Neurological effects
Reported neurological effects have been associated with relatively high exposure to parent TCE;
therefore, they are more frequent after acute and short-term exposure. Intermediate- duration
exposures to TCE have produced neurological effects similar to those found in acute-exposure
situations, including hearing loss (Crofton, Lassiter & Rebert, 1994), increased latency in
visual discrimination (Blain, Lachapelle & Molotchnikoff, 1992), and increased disinhibition
or excitability (ATSDR, 2019).
Most of these effects were found to be reversible when the exposure period ended. When the
neurological effects were evaluated at different doses and time of exposure, concentration was
more relevant than time of exposure in determining effects (Bushnell 1997; ATSDR, 2019).
5.3.3 Reproductive and developmental effects
In a reproductive toxicity study, Long–Evans rats were exposed by inhalation to TCE at
9700 mg/m3 for 6 hours per day, 5 days per week, for 12 weeks before mating; for 6 hours per
day, 7 days per week, only during pregnancy through gestation day (GD) 21; or for 6 hours per
day, 5 days per week, for 2 weeks before mating and for 6 hours per day, 7 days per week,
during pregnancy through GD 21.
Incomplete ossification of the sternum, indicative of delay in maturation, occurred in animals
exposed during pregnancy, and a significant decrease in postnatal weight gain occurred in
offspring of the premating exposed group. No maternal toxicity, teratogenicity or other effects
on reproductive parameters were observed (Dorfmueller et al., 1979).
Trichloroethene in drinking-water
19
In a two-generation reproductive toxicity study, male and female Fischer 344 rats were fed
diets containing microencapsulated TCE at doses of approximately 0, 75, 150 or
300 mg/kg bw/day from 7 days before mating through to the birth of the F2 generation.
Although left testicular and epididymal weights decreased in the F0 and F1 generations, no
associated histopathological changes were observed. The weight changes were attributed to
general toxicity, rather than reproductive toxicity (NTP, 1986). In a similar two-generation
reproductive toxicity study in CD-1 mice given TCE at up to 750 mg/kg bw/day, sperm
motility was reduced by 45% in F0 males and 18% in F1 males, but there were no treatment-
related effects on mating, fertility or reproductive performance in the F0 or F1 animals (NTP,
1985).
Numerous teratogenicity studies have been conducted using TCE administered by oral or
inhalation routes. Exposure of Swiss Webster mice to TCE by inhalation at 1600 mg/m3 for
7 hours per day on GD 6–15 did not result in treatment-related maternal toxicity or
teratogenicity (Leong, Schwetz & Gehring, 1975). When Swiss Webster mice and Sprague–
Dawley rats were exposed to TCE by inhalation at a concentration of 1600 mg/m3, 7 hours per
day on GD 6–15, a significant decrease (P < 0.05) in maternal weight gain and some evidence
of haemorrhages in the cerebral ventricles were observed, but no teratogenic or reproductive
effects were seen (Schwetz, Leong & Gehring, 1975). In contrast, a significant decrease in fetal
weight and some increase in fetal resorptions were reported in rats (strain not specified)
exposed to TCE at 540 mg/m3 for 4 hours per day during GD 8–21 (Healy, Poole & Hopper,
1982).
In a study of the effect of exposure to TCE on developmental/reproductive function, female
Sprague–Dawley rats were exposed to TCE in drinking-water at 0, 1.5 or 1100 mg/L (equal to
0, 0.18 or 132 mg/kg bw/day) in one of three dose regimens: for 3 months before pregnancy;
for 2 months before and 21 days during pregnancy; or for 21 days during pregnancy only
(Dawson et al., 1993). No maternal toxicity was observed at any dose level or regimen. An
increase in incidence of fetal heart defects was observed in treated animals at both dose levels
(8.2% at 0.18 mg/kg bw/day and 9.2% at 132 mg/kg bw/day, versus 3% in controls) in dams
exposed before and during pregnancy, and only at the high dose (132 mg/kg bw/day; 10.4%,
versus 3% in controls) in animals exposed only during pregnancy. The LOAEL was set at
0.18 mg/kg bw/day, based on the increased incidence of heart defects in fetuses born to dams
that were exposed before and during gestation. However, the study was limited in that it
expressed the incidence of malformations only as a proportion of the total number of fetuses in
the dose group and did not attempt to establish the incidence of heart defects on a per-litter
basis. Despite this shortcoming, the study lends support to similar findings of increased
congenital defects in epidemiological studies (Goldberg et al., 1990; Bove, Fulcomer & Klotz,
1995), although a clear dose–response relationship is lacking.
A subsequent study (Fisher et al., 2001) conducted with Sprague–Dawley rats treated with TCE,
TCA and DCA at dose levels as high as 400 mg/kg bw/day failed to reproduce the heart
malformations reported by Dawson et al. (1993). However, the two studies differed in design,
which may partly account for the incongruence of the results. First, the Fisher et al. (2001)
study used soybean oil as a vehicle, whereas the Dawson et al. (1993) study used water as a
vehicle. Second, Fisher et al. (2001) administered a very large dose of TCE (400 mg/kg bw/day)
in soybean oil in boluses on GD 5–16 only, whereas Dawson et al. (1993) administered TCE
in drinking-water at lower doses (maximum 1100 mg/L, or 129 mg/kg bw/day) ad libitum,
either during the entire gestation period (GD 1–21) or before and throughout pregnancy; both
the form of test agent and the timing of the dosage may partly account for the variations
Trichloroethene in drinking-water
20
between the two studies. Third, the Fisher et al. (2001) study had a very high background
incidence of heart malformations (52% on a per-litter basis) among the control fetuses that
were dosed with soybean oil only – this rate is much higher than the incidence of heart
malformations in the parallel water controls (37%). The Dawson et al. (1993) study reported a
much lower incidence of heart malformations (25% on a per-fetus basis) in the control fetuses
that were dosed with water only. The high background incidence of heart malformations
associated with the controls in the Fisher et al. (2001) study might have masked the effects in
the TCE treatment groups.
In another developmental toxicity study by Johnson et al. (2003), pregnant Sprague–Dawley
rats (9–13 per exposure level) were exposed to TCE throughout pregnancy in drinking-water
at concentrations of 0, 0.0025, 0.25, 1.5 or 1000 ppm. On GD 22, there was a significant
increase in the percentage of offspring with abnormal hearts in the treated groups. The
percentage of litters with abnormal hearts ranged from 0 to 66.7%, while 16.4% of control
litters had abnormal hearts. Although this study appears to suggest the presence of a dose–
response relationship, with the effects beginning to manifest at a dose of 250 μg/L (0.25 ppm;
corresponding to 0.048 mg/kg bw/day) and a NOAEL of 2.5 μg/L (0.00045 mg/kg bw/day),
the dose–response relationship is not as clear on closer examination of the data. However, US
EPA (2011b) calculated a rat BMDL01 (lower 95% confidence limit on the benchmark dose for
a 1% response) of 0.0207 mg/kg bw/day from the fetal heart malformation incidence data. The
BMDL01 was preferred to the NOAEL because the selected nested model accounts for
intralitter effects (i.e. the tendency for littermates to respond more similarly to one another than
to the other litters in a dose group), using pups as the unit of analysis since using litter as the
unit may not be optimal for detecting effects.
5.3.4 Immunological effects
Studies on animals indicate TCE as producing some immunotoxic effects, as reviewed by
IARC (2014). These include accelerated autoimmune responses in mice that are prone to
autoimmune disease, autoimmune hepatitis, inflammatory skin lesions and reduced thymus
weight.
The potential developmental immunotoxicity induced by TCE was studied in B6C3F1 mice
(Peden-Adams et al., 2006), by treating parents (C3H/HeJ male and C57BL/6N female mice;
five per sex per group) with TCE in the drinking-water at 0, 1.4 or 14 ppm, beginning at pairing
(1:1) and continuing for 7 days of mating, and throughout gestation and lactation. Pups were
evaluated for body length, timing of eye opening and ear unfolding. At weaning of the pups at
3 weeks of age, 5–7 pups per treatment group were weighed and sacrificed to assess kidney,
liver, thymus and spleen weights. TCE-related effects on the immune system were assessed by
measuring splenic lymphocyte proliferation, NK cell activity, plaque-forming cell (PFC)
response, splenic B220+ cells, and thymic and splenic T-cell immunophenotypes. The
remaining pups (4–5 pups per treatment group) were assessed at 8 weeks of age in a manner
similar to those assessed at 3 weeks of age, with additional assessments of autoantibodies to
dsDNA (double-stranded DNA) and delayed-type hypersensitivity response. At the lower dose
tested (estimated maternal dose of 0.37 mg/kg bw/day), a decreased PFC response was
observed in 3- and 8-week-old pups, and increased delayed-type sensitivity was noted in 8-
week-old pups. The LOAEL derived from the study was therefore 0.37 mg/kg bw/day.
In the study of Keil et al. (2009), developmental immunotoxicity was evaluated by treating
groups of 9-week-old female B6C3F1 mice (9–10 per group), which are not prone to
spontaneous autoimmune disorders, with TCE in drinking-water at 0, 1.4 or 14 ppm in 1%
Trichloroethene in drinking-water
21
Emulphor vehicle for 30 weeks (the low dose was estimated at 0.35 mg/kg bw/day). During
the exposure period, serum levels of total IgG and some autoantibodies (anti-ssDNA [single-
stranded DNA], anti-dsDNA and anti-glomerular antigen) were monitored. At sacrifice, spleen,
thymus, liver and kidney were weighed. Spleen and thymus were processed for assessment of
cell counts and activity, whereas kidneys were processed for histopathologic evaluation. TCE
did not alter NK cell activity, or T- and B-cell proliferation. At the low dose, decreased thymus
weight (30% lower than controls), and increased serum levels of IgG and selected
autoantibodies were observed, but overall there was no evidence that TCE accelerated the onset
of autoimmune disease (ATSDR, 2019).
5.3.5 Genotoxicity and carcinogenicity
A range of assays, covering a wide spectrum of genetic end-points, has been performed to
assess possible genotoxic effects of TCE or its metabolites. DNA- or chromosome-damaging
effects have been evaluated in bacteria, fungi, yeast, plants, insects, rodents and humans, using
many different end-points.
TCE-induced genotoxicity and its possible mechanism have been reviewed (WHO, 2005;
IARC, 2014; ATSDR, 2019).
Evidence for TCE genotoxicity is often conflicting, in part because of the presence of
impurities or mutagenic stabilizers in the test material. In fact, the information from many of
the early studies (until the mid-1990s) may not be adequate for complete evaluation of the
genotoxic potential of TCE, as few of the studies identified the grade and purity of the test TCE.
In addition, some TCE samples used contained a mutagenic stabilizer, and other assays used
pure samples without stabilizers; the TCE in the latter might have decomposed to chemicals
with mutagenic activity, further confounding the interpretation of the significance of the
findings.
TCE is weakly active both in vitro and in vivo, inducing recombination responses, including
sister chromatid exchange, and aneuploidies, including micronuclei; however, it appears to be
unable to induce gene mutations or structural chromosomal aberrations. TCE was also observed
to induce increased DNA synthesis and mitosis in mouse liver in vivo (WHO, 2005). The
potential for TCE to cause DNA strand breaks in rodent liver cells in vivo and in culture at
high concentrations (Bull, 2000) has been questioned (Styles, Wyatt & Coutts, 1991; Chang,
Daniel & DeAngelo, 1992), and was shown not to occur in the kidney (Clay, 2008).
Genotoxicity studies have been conducted for the major metabolites of TCE. CH, DCA and
TCA require very high doses to be genotoxic; insufficient information was available to draw
any conclusions for TCOH, and the conjugates DCVC and DCVG (Moore & Harrington-Brock,
2000). Nevertheless, the US EPA, in revising the genotoxicity of selected TCE metabolites,
concluded that there is relatively strong evidence for genotoxicity of CH and some evidence
for genotoxicity of other TCE metabolites, including DCA, DCVC and DCVG (US EPA,
2011b). Firm conclusions on whether TCE has a mutagenic mode of action cannot be drawn
from the available information.
Overall, results of testing in mammalian and nonmammalian test systems indicate a potential
for TCE to induce chromosomal damage. The weight of evidence suggests that TCE does not
act directly as a mutagenic agent, but that some metabolites have a genotoxic potential.
Trichloroethene in drinking-water
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Carcinogenicity in animal models has been extensively reviewed (US EPA, 2011b; IARC,
2014; ATSDR, 2019). After TCE exposure by the oral route, various types of cancers have
been found in rodents. However, in many studies, rats were administered the maximum
tolerated dose, resulting in a poor survival rate, which affects interpretation of the data. In
addition to other limitations in study design, stabilizers such as epichlorohydrin and other
epoxides were used to prevent TCE degradation when exposed to light. Coexposure to these
stabilizers can be a significant confounding factor; indeed, the observed forestomach tumours
were believed to be induced by the direct alkylating epoxides that were used as stabilizers
(Henschler et al., 1984).
Carcinogenicity studies of TCE by the oral route in mice have demonstrated treatment-related
liver tumours in mice in both sexes (NCI, 1976; NTP, 1983, 1988, 1990). Oral exposure to
TCE has also been shown to increase malignant lymphomas in female mice (US EPA, 2011b).
An increase in the incidence of testicular interstitial cell tumours was reported in male rats;
however, inadequacies in this study mean that the data could not be conclusively interpreted
(NTP, 1988).
In a carcinogenicity assay exposing rodents to TCE by gavage (NTP, 1983), there was a
significant increase in the incidence of hepatocellular carcinomas (P < 0.05) at
1000 mg/kg bw/day in male mice (13/49 relative to 8/48 in controls) and hepatocellular
adenomas (P < 0.05) in female mice (8/49 compared with 2/48 in controls). There were no
treatment-related liver tumours in rats. The male rats at 1000 mg/kg bw/day that survived until
the end of the study showed a higher (P = 0.028) incidence of renal tubular cell
adenocarcinomas (3/16 compared with 0/33 among controls). These kidney tumours were
considered biologically significant, given the rarity of kidney tumours in that rat strain.
A close audit of another carcinogenicity study (NTP, 1988) that exposed four different rat
strains (ACI, August, Marshall and Osborne–Mendel) to TCE by gavage indicated that the
study documentation was inadequate to support proper interpretation of the reported tumour
incidence data. No other treatment-related tumours were reported in these rat strains.
In the NTP (1990) carcinogenicity study, which exposed B6C3F1 mice and F344/N rats to TCE
by gavage, there was a significant (P < 0.05) increase in the incidence of combined
hepatocellular carcinomas and adenomas (P < 0.05) in female mice (22/49 at
1000 mg/kg bw/day, compared with 6/48 in untreated controls). No treatment-related kidney
tumours were observed in mice. Although the study authors considered the results equivocal
because of reduced survival in the treated groups, the kidney tumour incidences in rats were
statistically significant (P < 0.05) when adjusted for reduced survival (2/46 at
500 mg/kg bw/day and 3/33 at 1000 mg/kg bw/day, compared with none in controls); these
tumours were considered toxicologically significant because of the rarity of kidney tumours in
this rat strain.
Carcinogenicity studies of TCE by the inhalation route have shown treatment-related tumours
in the lungs of female and male mice (Fukuda, Takemoto & Tsuruta, 1983; Maltoni et al., 1986),
testes of rats (Maltoni et al., 1986), the lymphoid system (lymphomas) in female mice
(Henschler et al., 1980), the kidney in male rats and the liver in mice of both sexes (Maltoni,
Lefemine & Cotti, 1986).
Overall, animal carcinogenicity studies conducted using pure TCE showed that chronic
exposure to this compound by the oral route resulted in malignant liver tumours in mice of both
sexes and kidney tumours in male rats. Inhalation exposure led to lymphomas in female mice,
Trichloroethene in drinking-water
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malignant liver and lung tumours in mice of both sexes, and malignant kidney tumours in male
rats.
5.4 Mode of action
At TCE concentrations found in most occupational and environmental settings, TCE is
absorbed through the skin by diffusion. At very high concentrations, the absorption can be
enhanced by TCE-induced disruption of the phospholipidic structure of cell membranes
(solvent effect).
After absorption from the gastrointestinal tract, TCE is distributed to the liver first, where toxic
and nontoxic metabolites can be formed (first-pass effect). The toxicity of TCE does not seem
to be heavily dependent on its route of entry.
The similarity between carcinogenic effects induced by the parent compound and tested
metabolites suggests that TCE metabolites are mostly responsible for the liver and kidney
tumours observed in TCE bioassays. This seems to be particularly true for renal cell carcinoma;
additional supporting evidence for the involvement of metabolites derived from the GSH-
dependent pathway, at least in humans, is provided by the lower levels of DNA adduct
formation in individuals with the GST null genotype, who have limited formation of genotoxic
DCVC metabolites (Brüning et al., 1997b).
The mode of action for TCE-induced human renal carcinomas potentially involves mutation of
the VHL tumour suppressor gene, followed by induction of neoplasia (Brüning et al., 1997a).
Indeed, multiple mutations of the VHL tumour suppressor gene, primarily C to T changes,
including nucleotide 454, were found in renal carcinoma patients with prolonged exposure to
high levels of TCE (Brüning et al., 1997b; Brauch et al., 1999). These findings are consistent
with the finding that exposure to TCE at high levels, as in occupational settings (compared
with environmental levels), is highly likely to produce kidney cancer in humans.
The complexity of TCE metabolism and clearance complicates the identification of a
metabolite that is responsible for TCE-induced effects. More than one mode of action may
explain TCE-induced carcinogenicity, and several hypotheses have been put forward. A
number of events are likely to be significant to tumour development in rodents under bioassay
conditions. However, the events that may be more relevant to human exposure to TCE at
environmental levels are not known.
Peroxisome proliferation has been correlated with TCE-induced mouse liver carcinogenesis,
since tumours arise in parallel with peroxisome proliferation associated with TCE metabolites
(Elcombe, 1985; Elcombe, Rose & Pratt, 1985; Goldsworthy & Popp, 1987; Melnick et al.,
1987; DeAngelo et al., 1989; Cattley et al., 1998). TCA can activate the peroxisome proliferator
activated receptor-alpha (PPARα) and the subsequent cascade of responses, including effects
on gene transcription. However, peroxisome proliferation has not been observed in humans, so
chemicals acting with this mechanism in rodents would be unlikely to present a liver
carcinogenic hazard to humans.
Some observations suggest that PPARα activation is not the only mechanism for the
hepatocarcinogenicity of TCE (Rusyn et al., 2014):
• TCA also induces peroxisome proliferation in rats, although to a lower extent than in mice,
but TCA has been shown not to induce liver tumours in F344 rats (DeAngelo et al., 1997).
Trichloroethene in drinking-water
24
• The patterns of H-ras mutation frequency associated with DCA and other peroxisome
proliferators are different.
• Other TCA epigenetic effects (see below), including increased c-myc expression and
hypomethylation of DNA, are not specific to the PPARα activation mechanism and can
contribute to TCE-induced liver cancer (Rusyn et al., 2014).
The potential for peroxisome proliferation to play a role in TCE-induced kidney toxicity has
been assessed and is considered unlikely (Lash, Parker & Scott, 2000; Rusyn et al., 2014).
Although TCE has been reported to cause peroxisome proliferation in rat and mouse kidney,
with mice showing a greater response, it has not been shown to induce kidney cancer in mice.
In addition, studies indicate that renal peroxisomes are generally less responsive to peroxisome
proliferators than hepatic peroxisomes (Lash, Parker & Scott, 2000).
Since TCE does not cause α2µ-globulin accumulation (Goldsworthy et al., 1988), this
mechanism is not plausible. In addition, TCE has been identified as causing kidney damage in
both male and female rats (Barton & Clewell, 2000), whereas only male rats are known to
accumulate α2µ-globulin. As such, α2µ-globulin accumulation does not appear to be a mode of
action of TCE-induced kidney toxicity, as was previously thought.
The cysteine and GSH intermediates formed during the metabolism of TCE, DCVC and DCVG
have been shown to have some genotoxic potential, and also to induce the expression of proto-
oncogenes, including c-jun, c-fos and c-myc, in mouse liver tumours (Tao et al., 2000a, b). The
proto-oncogene c-myc is believed to be involved in control of cell proliferation and apoptosis,
which also points towards epigenetic mechanisms for the induction of liver tumours in mice.
Other evidence supports a cytotoxic mode of action. Most rats chronically exposed to TCE in
the National Cancer Institute (NCI) and National Toxicology Program (NTP) bioassays
developed toxic nephrosis, and more than 90% of rats (and mice) developed cytomegaly, which
was most evident in male rats. Associated with these findings, the incidence of kidney tumours
increased only in male rats. The TCE conjugates formed by the action of the β-lyase enzyme
produce proximal tubular necrosis and other lesions in rat kidney (Goeptar et al., 1995) that
lead to the production of reactive species. These reactive species may be responsible for
nephrotoxicity, as well as repair and proliferative responses along a continuum that may
ultimately result in tumorigenesis (Lash, Parker & Scott, 2000; Vaidya et al., 2003).
The formation of DCVG and DCVC in humans indicates that the mode of action associated
with their genotoxicity may be relevant at relatively high dose of exposure in humans. However,
it is uncertain what role it might play in human cancers induced by TCE at exposure levels
below the one activating the GST-dependent pathway or those expected to cause frank kidney
toxicity due to the different TCE metabolites.
Lung tumours were induced in female mice following exposure to TCE (Odum, Foster & Green,
1992). Accumulation of the TCE metabolite CH is thought to be the cause of TCE lung
carcinogenicity, as CH exposure results in lung lesions identical to TCE-induced tumours
(Green, Mainwaring & Foster, 1997; Green, 2000). Accumulation of CH in the Clara cells of
the lung is thought to lead to lung tumours by causing cell damage and compensatory cell
replication, which leads to tumour formation (Green, Mainwaring & Foster, 1997; Green,
2000). However, the mechanism by which CH results in tumour formation in animals may not
be pertinent to humans, as there is little CYP2E1 activity in human lungs (Green, Mainwaring
& Foster, 1997; Green, 2000). A specific lesion, characterized by vacuolization of Clara cells,
was seen only in mice; mice exposed to CH at 600 mg/m3 had similar lesions. Only mild effects
Trichloroethene in drinking-water
25
were seen with inhaled TCOH, and none with intraperitoneally administered TCA. These
results suggest that acute lung toxicity of TCE may be due to accumulation of chloral in Clara
cells in mice. Chloral is also genotoxic, and the toxicity observed with intermittent exposures
to TCE is likely to exacerbate any genotoxic effect through compensatory cell proliferation in
rodents.
In conclusion, the mode of action for tumour induction by TCE may be attributed to:
• nongenotoxic processes related to cytotoxicity, peroxisome proliferation, production of
reactive oxygen species and altered cell signalling; or
• genotoxic processes, such as the production of genotoxic metabolites (e.g DCVC) in the
kidney.
The latter possibility cannot be fully ignored, considering evidence of human DNA adducts
formed from genotoxic DCVC metabolites and the presence of VHL tumour suppressor gene
mutations in TCE-exposed kidney cancer patients (Brüning et al., 1997a).
Information on the mode of action for noncancer effects of TCE is more limited, and support
for hypotheses is largely based on observations of actions of other agents. The major endocrine
system effects associated with TCE exposure include the development of testicular (Leydig
cell) tumours in rats (Maltoni et al., 1988; NTP, 1988). TCE and its metabolites TCA and
TCOH have been found to concentrate in the male reproductive organs of rats following
inhalation exposure (Zenick et al., 1984). They have also been found in seminal fluids of
humans occupationally exposed to TCE (Forkert et al., 2003). Peroxisome proliferating
chemicals have been shown to induce Leydig cell tumours via a modulation of growth factor
expression by estradiol (Cook et al., 1999). The occurrence of Leydig cell tumours in rats
exposed to TCE may therefore act as a signal for disturbance of the endocrine system. This
could point to potential endocrine disturbances in humans as a result of TCE exposure. The
effect of endocrine disruption in human populations exposed to TCE is an area requiring further
research.
Studies of the mode of action for observed developmental effects seen with TCE, TCA and
DCA exposure, and data specific to TCE exposure are also scant. Again, a possible role for
peroxisome proliferation with PPARα activation in the development of eye anomalies
following TCE exposure has been hypothesized, although no data currently support it
(Narotsky & Kavlock, 1995; Narotsky et al., 1995).
The TCE metabolites TCA and DCA both produce cardiac anomalies in rats (WHO, 2005).
DCA also concentrates in rat myocardial mitochondria (Kerbey et al., 1976), freely crosses the
placenta (Smith, Randall & Read, 1992) and has known toxicity to tissues dependent on
glycolysis as an energy source (WHO, 2005).
More research into TCE and its metabolites is needed to more fully elucidate possible modes
of action for the effects observed in standard developmental protocols.
6 Overall database and quality of evidence
6.1 Summary of health effects
TCE has anaesthetic properties when inhaled at high concentrations.
Trichloroethene in drinking-water
26
Available human and animal data obtained after repeated exposure to TCE identify the kidney,
liver, immune system, male reproductive system and developing fetus as potential targets of
TCE toxicity and/or carcinogenicity.
The systemic effects elicited by TCE are not exposure or route specific; indeed, similar effects
can be elicited via oral and inhalation exposure routes. PBPK models have been developed for
extrapolation from the inhalation route to the oral route, and also for predicting human exposure
levels that would result in effects similar to those observed in rodents.
There is some evidence for TCE-induced hepatic effects (e.g. changes in blood and urine
indices of liver function, enlarged livers) in occupationally exposed humans. Study limitations
include lack of quantifiable exposure data, and concomitant exposure to other chemicals. Dose-
related increases in liver weight, and hepatocellular hypertrophy and peroxisome proliferation
have been consistently reported in TCE-exposed animals.
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. Epidemiological data are limited by
concurrent exposure to other organic solvents; case–control studies are limited by small
numbers of cases. Studies in animals demonstrate the toxicity of TCE to the male reproductive
system.
Effects such as decreases in litter size and perinatal survival have been reported in rats at
maternally toxic oral doses. Effects were seen in pups exposed at quite high doses
(≥37 mg/kg bw/day). Cardiac arrhythmias were reported in rats exposed to TCE, but not in
humans, unless at lethal doses.
Increased incidences of tumours of the kidney, liver and lymphoid tissues have been reported
in chronic bioassays of rats and mice exposed to very high levels of TCE via inhalation and
oral exposure. Available human data on occupationally exposed subjects provide strong
support for TCE-induced kidney cancer; there are indications that the individuals carrying a
deletion of the GSTT1 gene are much less susceptible to TCE-induced renal tumours. This
suggests that, in humans at high exposure doses, the oxidative pathway is saturated and the
GST-mediated pathway is actively forming reactive metabolites in the kidney.
There is some evidence in humans for an association between exposure to TCE and non-
Hodgkin lymphoma. Evidence for TCE-induced liver cancer in humans is less convincing and,
according to NRC (2009), inadequate.
Associations between the incidence of leukemia and other cancers and oral exposure to TCE
are suggestive, but not definitive, as a result of confounding factors, co-exposures, and
inconsistency between study results.
NRC (2009) concluded that there is limited/suggestive evidence of an association between TCE
exposure and risk of kidney cancer, and inadequate/insufficient evidence for determining
whether associations exist between exposure to TCE and cancer risk at other sites. US EPA
(2011b) concluded that TCE is “carcinogenic to humans by all routes of exposure” based on
convincing evidence of a causal association between TCE exposure and kidney cancer in
humans. IARC (2014) classified TCE as carcinogenic to humans (Group 1), concluding that
the epidemiologic data provide sufficient evidence for an association between exposure to TCE
Trichloroethene in drinking-water
27
and human kidney cancer, whereas the associations reported for liver cancer and non-Hodgkin
lymphoma, although positive, are less consistent and thus characterized as limited for these two
cancers. The evidence for other tumours was classified as inadequate (IARC, 2014).
6.2 Quality of evidence
TCE is a data-rich compound: many good-quality studies are available on its kinetics and toxic
effects in both animals and humans. Studies of workers and volunteers have provided most of
the data on health effects of inhaled TCE in humans. Although available data on oral exposure
have been considered of questionable validity as a result of co-exposure to other contaminants,
the availability of PBPK modelling allows inhalation-to-oral route extrapolation.
Aspects that are not fully elucidated include:
• uncertainty about the mode of action for tumour induction, for which many mechanisms
have been proposed, but none of them definitely established; the potential genotoxicity of
some GST-mediated TCE metabolites at high exposure doses make the picture even more
complex; and
• regardless of the mode of action, the relevance for humans, which could not be completely
assessed for some tumours.
Kidney cancer has been reported in humans in studies considered to be of adequate quality, but
only at high levels of exposure, such as in occupational settings. Liver effects have been
consistently reported, although TCE has been shown to have multiple targets in animals and
humans. None of the studies can be considered optimal, but they all concur to build a weight-
of-evidence approach. For this reason, an overall reference value was derived by the US EPA
(2011b), with tolerable daily intake (TDI) values derived from some recent studies falling
within a narrow range of 0.0003–0.0006 mg/kg bw/day (see section 8.1), thus enhancing the
strength of the findings.
7 Practical considerations
7.1 Analytical methods and achievability
TCE can be analysed together with trichloroethene by gas chromatography using
ISO 10301:1997, which has a limit of quantification of 0.1 μg/L (ISO, 1997).
Four methods for measuring TCE in drinking-water have been approved by the US EPA:
• Method 502.2, which employs purge and trap capillary gas chromatography with
photoionization detectors and electrolytic conductivity detectors in series, has a detection
limit in the range of 0.01–3.0 μg/L (US EPA, 1999).
• Method 524.2, which uses purge and trap capillary gas chromatography with mass
spectrometric detectors in series, has a detection limit of 0.5 μg/L (US EPA, 1999).
• Method 503.1, which uses purge and trap capillary gas chromatography with
photoionization conductivity detectors, has a detection limit of 0.01–3.0 μg/L (US EPA,
1999).
• Method 551.1, which uses liquid–liquid extraction and gas chromatography with electron
capture detectors, has a method detection limit of 0.01 μg/L (US EPA, 1999).
Trichloroethene in drinking-water
28
7.2 Source control
TCE is primarily, if not exclusively, a groundwater contaminant because it is lost to the
atmosphere from surface waters. The primary cause of contamination is poor handling and
disposal practices, which result in soil contamination and, in vulnerable aquifers, subsequent
water contamination. This may occur some distance from the source, but contaminants may be
drawn into the source by pumping. Control at source should be relatively cheap and
straightforward by improving handling and disposal practices.
7.3 Treatment methods and performance
Treatment of surface water sources is not needed because TCE volatilizes to the atmosphere.
The most effective technique for removal of TCE from groundwater is aeration, including
packed tower aeration. Granular activated carbon (GAC) adsorption has also been shown to be
effective. TCE concentrations below 2 μg/L should be achievable by air stripping (Duan, Ito &
Ohkawa, 2001). Combining air stripping and activated carbon has been shown to improve TCE
removal.
Aeration has been used to treat contaminated well water (27 μg/L) at pilot scale. For an air-to-
water ratio of 10, a rate of 25 m/hour and a 3.75 m contact height, the process achieved a 67%
reduction in TCE (Simon & Mitchell, 1992).
Pilot-scale tests using air stripping achieved TCE removals from water with an influent
concentration of 204 μg/L of 82% and 87% for air-to-water ratios of 75:1 and 125:1,
respectively (McKinnon & Dyksen, 1984). Other pilot-scale studies using diffused aeration
have achieved removals of 70–92% using an air-to-water ratio of 4:1 and a 10 minute contact
time (Kruithof et al., 1985). One study investigated the effect of media depth on the removal
rate. A packed tower with a media depth of 4.5 m, an air-to-water ratio of 30:1 and a liquid
loading rate of 13.8 l L/m2∙s achieved a removal of 98.2%, whereas a packed tower with a
media depth of 1.2 m achieved a removal of 45% under the same conditions (Amy, Narbaitz
& Cooper, 1987).
Gross & TerMaath (1985) studied the performance of packed tower aeration for stripping TCE
from groundwater at full scale. The groundwater contained TCE concentrations ranging from
50 to 8000 µg/L. During the study, two towers were operated in both series and parallel
configuration at air-to-water ratios of 10, 18 and 25. Higher removals were observed in the
series configuration (96–99.9%) than in the parallel configuration (86–99%). The highest
removal was observed at an air-to-water ratio of 25. Hand et al. (1988) studied the effectiveness
of air stripping for removing TCE from contaminated well water and found that, at an air-to-
water ratio of 60:1, approximately 98% of the influent TCE was removed (influent
concentration of 72 µg/L). Cummins (1985) found that, with a packing height of 5.5 m and an
air-to-water ratio of 39:1, approximately 98.5% of the influent TCE was removed (influent
concentrations of 20–38 mg/L) from three wells used for source.
Full-scale spray aeration of well water containing TCE at up to 10 μg/L achieved 90% removal
to below 1 μg/L (Kruithof & Koppers, 1989).
GAC has been used to remove high concentrations of TCE at pilot scale. The carbon effectively
removed 100% of the influent concentration (approximately 2500 μg/L), for 30 bed volumes
at an empty bed contact time (EBCT) of 2.5 minutes and 40 bed volumes at an EBCT of 10 min
Trichloroethene in drinking-water
29
(Hand et al., 1994). The presence of humic substances can decrease GAC adsorption (Urano,
1991).
A study of 68 full-scale treatment facilities in the USA treating water for volatile organic
compounds (VOCs) found that facilities employing GAC achieved TCE removal efficiencies
of >99%. Influent concentrations ranged from 3 to 400 µg/L, and the treatment facilities
operated under a variety of different configurations and design parameters; loading rates varied
from 4.6 to 18.5 m/hour, and EBCT varied from 9 to 30 minutes (AWWA, 1991).
In a municipal-scale treatment plant combining air stripping and GAC, TCE was removed to
levels below 1 µg/L (US EPA, 1985b).
Ozone doses of 2, 6 and 20 mg/L achieved TCE removals of 39%, 76% and 95%, respectively
(Fronk, 1987). Pilot plant studies have shown that ozonation can almost completely remove
trace concentrations of TCE from groundwater (Slagle, 1990).
Karimi et al. (1997) investigated the use of a two contactor ozone/hydrogen peroxide advanced
oxidation process for TCE oxidation at full scale. The influent TCE concentration ranged from
32 to 477 µg/L. In all tests, the treated water concentration of TCE was below 5 µg/L in the
second contactor. Greater than 90% removal of TCE was achieved in all tests after the first
contactor, with effluent TCE concentrations <5 µg/L when the influent concentration was
<100 µg/L. The greatest reduction in TCE (≥99%) occurred for ozone and H2O2 doses of 4.0–
4.6 mg/L and 2.2–2.4 mg/L, respectively. The authors noted that the optimum ratio of
H2O2/ozone was 0.5–0.6.
A combination of H2O2 and ultraviolet (UV) irradiation has been used to treat groundwater
contaminated with VOCs, including TCE (0.89–1.30 mg/L). At 38 L/minute, with a reactor
volume of 57 L and with H2O2 dosed at 65 mg/L, the effluent concentration of TCE was
generally below detection limits (maximum removal efficiency was >99.9%) (Topudurti et al.,
1994). Other research has confirmed that TCE is readily removed from water by ozone and that
UV irradiation gave only a slight improvement; 75 mg/L was removed by an ozonation rate of
6 mg/L and UV fluence of 100 mW∙s/cm2 (Pailard, Brunet & Dore, 1987).
Microbial remediation of TCE has also been described (Pant & Pant, 2010).
8 Conclusion
8.1 Derivation of the guideline value
The previous World Health Organization (WHO) evaluation (WHO, 2005) considered both
cancer and noncancer end-points in deriving the guideline value (GV) for TCE in drinking-
water. The GV for TCE was ultimately based on the noncancer end-points, and was protective
for both cancer and noncancer end-points.
In the previous WHO evaluation (WHO, 2005), the developmental toxicity study from Dawson
et al. (1993) was chosen as the point of departure (POD), based on the appropriateness of the
route (drinking-water), the low dose at which the effects were observed (which coincides with
the LOAEL in all animal studies reviewed), the severity of the end-point, the evidence of
similar effects (e.g. cardiac anomalies) from epidemiological studies, and the observation of
similar malformations in studies of TCE metabolites. However, it has been recognized that the
Dawson et al. (1993) study has several significant methodological limitations, including the
spontaneous incidence of the critical end-point (heart malformations).
Trichloroethene in drinking-water
30
In the present evaluation, it was considered more appropriate to take into account potential
PODs for candidate chronic TDI values from various studies by using the LOAEL/NOAEL
approach, benchmark dose (BMD) analysis, and PBPK modelling of human and animal data
considered suitable for assessment of a dose–response relationship (US EPA, 2011b).
Therefore, all the possible candidate PODs, rather than a single key study, were included in the
derivation of the TDI, considering various end-points. End-points included 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.
PBPK modelling was used to calculate internal doses from a number of studies for which
plausible internal dose metrics were available, based on present understanding of the role that
different metabolites play in TCE toxicity. The exception was the Peden-Adams et al. (2006)
study, because model parameters were not available; therefore the LOAEL was used.
The PBPK model estimated interspecies and intraspecies pharmacokinetic variability, and
resulted in 99th percentile estimates of human equivalent dose (HED99) for candidate 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 exposure time resulted in a change of less than a
few percentage points in the resulting HED).
Among the available studies, three were considered critical to deriving the TDI:
• Keil et al. (2009), from which a LOAEL of 0.35 mg/kg bw/day was identified, based on
decreased thymus weight in female mice exposed to TCE in the drinking-water for
30 weeks. A PBPK model was used to derive an HED99 of 0.048 mg/kg bw/day for lifetime
continuous exposure, which was used as the POD.
• Peden-Adams et al. (2006), from which a LOAEL of 0.37 mg/kg bw/day was identified
and considered as the POD, based on developmental immunotoxicity effects: decreased
plaque-forming cell response (at 3 and 8 weeks of age) and increased delayed-type
hypersensitivity (at 8 weeks of age) in pups exposed from GD 0 until 3 or 8 weeks of age
through drinking-water (placental and lactational transfer, and pup ingestion). A BMD
could not be calculated because of inadequate model fit, and no PBPK modelling was
applied because of lack of appropriate models and parameters to account for fetal and pup
exposure patterns.
• Johnson et al. (2003), in which pregnant Sprague–Dawley rats were administered TCE in
drinking-water during GD 1–22 at concentrations ≥0.0025 ppm. Increased incidences of
fetal cardiac malformations at maternal exposure levels ≥0.25 ppm (estimated maternal
doses ≥0.048 mg/kg bw/day) were identified as the critical effect. A PBPK model was
applied to the rat BMDL01 external dose of 0.0207 mg/kg bw/day to calculate the rat
internal dose; this was converted to an HED99 of 0.0051 mg/kg bw/day.
No new carcinogenicity data were identified since the previous evaluation (WHO, 2005). The
key studies have been selected among the ones identifying noncancer effects, with a POD that
is lower than that used in the previous evaluation.
Trichloroethene in drinking-water
31
8.1.1 Noncancer effects
Limitations of individual studies were overcome through selection of multiple critical effects,
rather than selecting the lowest NOAEL or BMDL as a POD.
By using this approach, as described by US EPA (2011b), a PBPK model was used to calculate
an internal dose POD from a number of studies for which plausible internal dose metrics could
be determined, based on present understanding of the role that different metabolites play in
TCE toxicity and the mode of action for toxicity.
The PBPK model was used to estimate interspecies and intraspecies pharmacokinetic
variability, and resulted in an HED99 for candidate critical effects.
From the three critical studies, the TDI derivation was as follows:
• Keil et al. (2009)
HED99 = 0.048 mg/kg bw/day
Uncertainty factor (UF) = 10 to account for use of a LOAEL
UF = 2.5 to account for remaining uncertainty associated with interspecies
toxicodynamic differences, because a PBPK model was used to characterize interspecies
toxicokinetic differences
UF = 3.2 to account for remaining uncertainty associated with human variability in
toxicodynamics, because a PBPK model was used to characterize human toxicokinetic
variability
TDI = 0.048/80 = 0.0006 mg/kg bw/day
• Peden-Adams et al. (2006)
LOAEL = 0.37 mg/kg bw/day
UF = 10 to account for use of a LOAEL
UF = 10 for interspecies extrapolation (a default factor was used, because of lack of
adequate toxicokinetic data to develop a PBPK model)
UF = 10 for human variability (a default factor was used, because of lack of adequate
toxicokinetic data to develop a PBPK model)
TDI = 0.37/1000 = 0.00037 mg/kg bw/day
• Johnson et al. (2003)
HED99 = 0.0051 mg/kg bw/day (derived from a BMDL01)
UF = 2.5 to account for remaining uncertainty associated with interspecies
toxicodynamic differences, because a PBPK model was used to characterize interspecies
toxicokinetic differences
Trichloroethene in drinking-water
32
UF = 3.2 to account for remaining uncertainty associated with human variability in
toxicodynamics, because a PBPK model was used to characterize human toxicokinetic
variability
TDI = 0.0051/8 = 0.00064 mg/kg bw/day
The TDI values fall within a narrow range of 0.0003–0.0006 mg/kg bw/day. The PBPK model-
based TDI value is 0.0006 mg/kg bw/day for both heart malformations in rats (Johnson et al.,
2003) and decreased thymus weights in mice (Keil et al., 2009). The lowest TDI comes from a
third study (Peden-Adams et al., 2006), which allowed derivation of a TDI value of
0.00037 mg/kg bw/day, based on the applied dose LOAEL for developmental immunotoxicity.
Further supporting data in the database are toxic nephropathy in rats (NTP, 1988;
0.0003 mg/kg bw/day) and increased kidney weight in rats (Boverhof et al., 2013;
0.0008 mg/kg bw/day by using route-to-route extrapolation from the inhalation study).
An overall TDI of 0.0005 mg/kg bw/day (0.5 µg/kg bw/day) was considered appropriate, being
supported by multiple effects rather than an individual value. This approach is less sensitive to
limitations of individual studies.
8.1.2 Guideline value
Based on the TDI as described above, the GV is:
GV= 0.5 µg/kg bw/day × 60 kg bw × 0.5 = 7.5 µg/L (rounded to 8 µg/L or 0.008 mg/L)
2 L/day
where
• 0.5 µg/kg bw/day is the TDI, as derived above
• 60 kg is the average body weight of an adult
• 0.5 is the fraction of the total daily intake that is allocated to drinking-water
• 2 L is the daily volume of water consumed by an adult.
An allocation factor higher than the default 20% factor is used, since the occurrence of TCE in
food is low (see section 2.2) and human exposures to TCE overall have been declining (see
section 2.5), as a result of increased environmental regulations governing TCE emissions
(IARC, 2014; ATSDR, 2019).
8.2 Considerations in applying the guideline value
In certain circumstances, there may be a need to adapt the GV by adjusting the allocation factor
or considering the Leq/day corresponding to inhalation exposure from the domestic use of
water to account for local conditions, including inhalational exposure from high rates of
showering and bathing, and/or from living in poorly ventilated buildings. This indirect
exposure may be particularly relevant in such settings because of TCE evaporation rates; the
contribution of indirect exposure via indoor air might be similar to, or higher than, that from
oral water ingestion.
Requirements for monitoring TCE in drinking-water regulations and standards should be
limited to groundwater sources where a catchment risk assessment indicates the possibility of
presence of TCE, such as where TCE is, or was, used as a degreasing agent in large or small
Trichloroethene in drinking-water
33
industries with limited resources for safe handling and disposal. In some cases, this may include
TCE drawn in from contamination elsewhere in the aquifer. Monitoring can be conducted at
the treatment works. If concentrations are shown to be stable or effective treatment is in place,
the frequency of monitoring can be quite low.
If monitoring data show elevated levels of TCE, it is suggested that a plan be developed and
implemented to address these situations. Monitoring is not needed for surface water sources
because TCE volatilizes to the atmosphere.
Trichloroethene in drinking-water
34
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