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WHO Guidelines for Indoor Air Quality: Selected Pollutants. Geneva: World Health Organization; 2010.

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WHO Guidelines for Indoor Air Quality: Selected Pollutants.

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8Trichloroethylene

, , and .

General description

Trichloroethylene (TCE) (CAS Registry Number, 79-01-6; C6H3Cl3; molecular weight 131.40 g/mol) is a widely used industrial solvent. It is a volatile, colourless liquid with a sweet ethereal (chloroform-like) smell. It has a melting point of −84.8 °C, a boiling point of 86.7 °C, a Henry's Law constant of 1.03 × 10−2 atm-m3/mol at 20 °C, a vapour pressure of 7.8 kPa at 20 °C, a water solubility of 1.1 g/l at 20 °C and a log Kow (octanol–water partition coefficent) of 2.29 (1).

The synonyms of TCE include acetylene trichloride, ethinyl trichloride, trichloroethene, TRI, TRIC, 1-chloro-2,2-dichloroethylene, 1,1,2-trichloroethylene, Trilene and Triklone. Its structural formula is (1):

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Conversion factors

At 760 mmHg and 20 °C, 1 ppm = 5.465 mg/m3 and 1 mg/m3 = 0.183 ppm; at 25 °C, 1 ppm = 5.374 mg/m3 and 1 mg/m3 = 0.186 ppm (2).

Production

In 1990, the estimated industrial production of TCE was 131 kilotonnes in western Europe, 79 kilotonnes in the United States and 57 kilotonnes in Japan, compared to 210, 121 and 82 kilotonnes, respectively, in 1980. The estimated annual consumption in these regions is 65–103% of the production levels. No estimates of production and use are available for other parts of the world (3).

Applications and uses

TCE is mainly used for the vapour degreasing and cold cleaning of manufactured metal parts (80–95% of consumption). Other applications include industrial dry cleaning, printing, the production of printing ink, extraction processes, paint production and textile printing (24).

In the atmosphere, TCE may react with photochemically produced hydroxyl radicals yielding phosgene, dichloroacetyl chloride, formyl chloride and other degradation products. Its half-life in the atmosphere varies with latitude, season and concentration of hydroxyl radicals. Reported half-lives for the reaction with hydroxyl radicals range from 1 day to 2½ weeks (in polar regions, the half-life may be as long as several months in winter) (5,6).

Analytical methods in air

Common methods of measuring indoor concentrations of TCE include integrated active and passive sampling methods using tubes packed with a carbon-based sorbent or evacuated SUMMA canisters (passivated canister sampling apparatus), or diffusion samplers such as charcoal badges. Canister sampling involves controlling the flow of air into a pre-evacuated canister. Sorbent tubes and badges retain compounds according to the affinity of the sorbent for that compound. For sorbent or charcoal sampling, the analytes must first be extracted thermally or chemically, then separated using gas chromatography and identified using a detection method such as mass spectrometry (7). Personal exposure studies often use these sorbent-based methods.

Indoor sources and exposure pathways

Inhalation is the main route of exposure for the general population, but ingestion may significantly contribute to total exposure, particularly if drinking-water sources are highly contaminated. Because TCE can volatilize rapidly from surface water (4,6), contaminated water may be an additional source of indoor exposure, through showering or the use of washing machines and dishwashers, for example. Contaminated soil can also contribute to ambient air concentrations of TCE via vapour intrusion (where TCE in soil gas enters homes through cracks in the foundations) (2,8). Dermal exposure can contribute to the total exposure, especially through the use of detergent products or showering (1,9). In a risk assessment study, Fan et al. (10) showed that the three most important exposure pathways are water ingestion, dermal absorption when showering and breathing indoor air.

Consumer products

Consumers may be exposed to TCE through the use of wood stains, varnishes, finishes, lubricants, adhesives, typewriter correction fluid, paint removers and certain cleaners, where TCE is used as a solvent (11). Use of these products may result in elevated indoor air concentrations over background, although, as they are expected to be used intermittently rather than constantly, both short-term and long-term average concentrations are likely to be variable.

Groundwater and drinking-water

Groundwater levels are variable and subject to local contamination. In wide-ranging surveys, concentrations have been shown to be mostly low, of the order of < 0.2–2 μg/l. However, measurements in groundwater related to contaminated sites may show high levels; up to 950 mg/l has been found (1).

Levels of TCE in drinking-water are generally less than 1 μg/l, although higher levels (up to 49 μg/l) have been reported (1). In a survey carried out in Germany during the 1990s, TCE was detected in about 40% of the tested drinking-water samples, with a concentration of more than 1 μg/l for 5.5% of the supplied residents (concentration range < 0.001–21 μg/l) (6). A regular monitoring programme is in place for drinking-water in England and Wales, covering 31 areas. TCE levels were monitored at approximately 2700 sites in 1994. The vast majority of measured TCE levels were below the detection limit (range 0.1–3.0 μg/l), although higher levels (up to 25 μg/l) have occasionally been detected (1). The EPA Groundwater Supply Survey of drinking-water systems nationwide detected TCE in 91 out of 945 water systems in 1984. The median level of the positive samples was about 1 μg/l, with a maximum level of 160 μg/l (2). In Canada, the majority of drinking-water supplies contain less than 0.2 mg/l TCE (5).

Food

The presence of TCE has been demonstrated in a wide range of foodstuffs, and in some exceptional cases high concentrations have been detected. In some countries, the use of TCE as a solvent in the production of foodstuffs has been banned. In total diet studies carried out in the United States, TCE was detected in only a small proportion of the samples (2,5). In the United States, dairy products (particularly butter) and margarine have been found to have high levels of TCE. Levels of 73 μg/kg in butter and margarine were found, while cheese products had an average level of 3.8 μg/kg (11). TCE is lipophilic and has been detected in the breast milk of women living in urban areas, though no quantitative data are available (2). It is possible that this may be leading to a high daily intake for nursing infants.

Indoor air concentrations

Residential concentrations

Indoor concentrations of TCE have been found to be near or below 1 μg/m3 in many microenvironments where the presence of known sources is uncommon. In several United States cities (New York, Los Angeles, Chicago, Minneapolis, St Paul, Baltimore, Elizabeth and Houston), median indoor residential concentrations ranged from 0.1 to 0.5 μg/m3; in many cases, standard deviations were of the same order of magnitude as the medians (1219). Indoor : outdoor ratios in New York and Los Angeles, in a study in a low-income group, were between 1 and 2 at the median and did not have large standard deviations relative to the median (except for New York homes in the winter) (16). This indicates that indoor TCE was probably derived from outdoor sources in the homes in this study. In the United States studies mentioned, outdoor residential median levels ranged from 0.1 to 0.4 μg/m3 with little variability (1219). Sampling in different seasons did not show significant differences.

In the European EXPOLIS study, conducted in 1998–1999 using the same methods in six European cities, it was found that in cities with high outdoor residential concentrations such as Athens, Milan and Prague, indoor residential levels were also higher (20). In Athens, the median indoor concentration was 8.2 μg/m3 with a 90th percentile of 22.4 μg/m3, compared with a median of 4.3 μg/m3 outdoors and a 90th percentile of 33 μg/m3. In Milan, the median indoor concentration was 7.7 μg/m3 and the 90th percentile was 21.2 μg/m3, while the median outdoor concentration was 2.3 μg/m3 and the 90th percentile was 8 μg/m3. In Prague, the median indoor concentration was 13.6 μg/m3 and the 90th percentile was 28.9 μg/m3, while the median outdoor concentration was 3.7 μg/m3 and the 90th percentile was 7.7 μg/m3. These indoor levels are likely to be a reflection of the contribution of outdoor sources through infiltration, as the outdoor levels were much higher than in Helsinki, Basel and Oxford. In Helsinki, most of the indoor and outdoor samples were below the detection limit. Basel had similar indoor and outdoor levels to those in Helsinki, while Oxford had slightly higher levels (median 2.1 μg/m3 and 90th percentile 6.6 μg/m3 indoors, and median 2.5 μg/m3 and 90th percentile 4.9 μg/m3 outdoors). The higher indoor (median 7.7–13.6 μg/m3) than outdoor levels (median 2.3–4.3 μg/m3) overall in Milan, Athens and Prague indicate that there may be an additional contribution of indoor sources to indoor TCE concentrations.

A nationwide survey of residences in France (n = 567; monitoring carried out in 2003–2005) found median indoor concentrations of 1 μg/m3; the 95th percentile concentration was 7.4 μg/m3(21). Indoor concentration in the attached garage (n = 139) was below the detection limit at the median and 12.9 μg/m3 at the 95th percentile, suggesting the presence of indoor sources. Outdoor concentrations (n = 517) were below the detection limit at the median and 2.3 μg/m3 at the 95th percentile.

A study of 25 homes in Shimizu, Japan found indoor geometric mean concentrations of 0.22 μg/m3 (geometric SD = 2.16) in the summer and 0.36 μg/m3 (geometric SD = 1.64) in the winter (22). These were similar to the corresponding outdoor geometric mean concentrations of 0.23 μg/m3 (geometric SD = 2.14) measured in the summer of 2002 and 0.36 μg/m3 (geometric SD =1.61) in winter 2002 (22), indicating that again, infiltration from the outdoors was likely to have been the main source of TCE indoors.

Non-residential microenvironments

A survey of 70 office buildings in the United States (without any reported complaint) found a median TCE concentration of 0.29 μg/m3 with a 95th percentile of 2.6 μg/m3(23). A study of mechanically ventilated office sector and non-office sector spaces in 20 public buildings in Hong Kong SAR, China found arithmetic means of 5.6 μg/m3 (SD = 9.6) for offices and 8.8 μg/m3 (SD = 10.7) in non-office areas (24). Other studies of offices, restaurants and stores found TCE median or mean levels below 1 μg/m3, although there was a high degree of variability in some cases (2527). In general, it appears that, for the most part, TCE in non-residential indoor environments is associated with infiltration from outdoors, as common indoor environments are unlikely to have sources of TCE in amounts that would contribute enough to the indoor concentration to result in significantly elevated levels compared to outdoors. Studies have found that most non-industrial areas do not have significantly high outdoor levels compared to industrial areas.

Modelled air concentrations of TCE near different types of vapour degreasing machine have been estimated to range from around 5 mg/m3 for newer, closed-loop machines to over 1000 mg/m3 for older, open-top machines. In Germany, levels of TCE in degreasing applications based on occupational personal air concentrations in the 1990s were found to be around 20–50 mg/m3 on average (28).

Ambient air

Concentrations of TCE in ambient outdoor air may fluctuate widely over relatively short periods of time, depending on the strength of the emission source, variations in wind direction and velocity, and scavenging and photodecomposition (5). Arithmetic mean rural concentrations in Canada were found to be 0.02 μg/m3 with a maximum value of 4 μg/m3, based on passive month-long measurements in 2001 and 2002 (29).

Short-term measurements made during peak traffic hours on busy roads in industrial, commercial, residential and central business districts in Hong Kong SAR, China gave arithmetic mean TCE concentrations of 48.5 μg/m3 (SD = 77.8), 3.6 μg/m3 (SD = 3.4), 0.4 μg/m3 (SD = 0.5) and 1.3 μg/m3 (SD = 1.8), respectively (30). These patterns indicate that areas of heavy industry have significantly higher, although more variable, concentrations of TCE, probably due to the presence of high-emitting sources. Residential and non-industrial areas where few sources are present have much lower concentrations. Based on results from the European EXPOLIS study in Athens, Milan and Prague, it is likely that small industry sources are either mixed in with residential and business areas or that heavy industry is near these areas. Non-residential indoor environments in Athens, Milan and Prague were found to have similar TCE levels to indoor home environments. Median workplace concentrations were 6.4 μg/m3 in Athens, 4.7 μg/m3 in Milan and 4.4 μg/m3 in Prague (20).

Toxicokinetics and metabolism

Identification of studies

Considering the possible overlaps in the relevant evidence on health effects of trichloroethylene, the search of the literature supporting the “Toxicokinetics and metabolism” and “Health effects” sections was conducted in one process. The electronic searches were made in PubMed between August and September 2008, with an update in 2009. The keywords used were: “trichloroethylene” and “health effects” or “risk assessment” or “metabolism/biotransformation/kinetics”. We selected all relevant papers on this subject. Around 100 publications were selected; 23% concerned the metabolism of TCE and mostly the development of PBPK models, 16% focused on molecular mechanisms and mode of action, and 29% were related to toxicological tests in mammals. One third discussed reproductive and developmental effects while 25% concerned human studies, mostly the exposure of workers and pregnant women in the general population. Finally, two reviews on toxicological effects, three on carcinogenic risk in human and two on risk assessment were also selected.

A complementary Internet search was made in August 2008 on toxicological databases (hazardous substances databank, TOXNET) and on the web site of international or national health assessment agencies, including WHO, the European Commission, IPCS, ATSDR, Health Canada, the French Agency for Environmental and Occupational Health Safety (Affset) and USEPA. Twelve reports from these agencies, published between 1985 and 2009, were selected.

Toxicokinetics

Two update reviews of the pharmacokinetics of TCE have recently been published (31,32).

Absorption

In humans and in animals, TCE is readily absorbed via the oral, inhalation and dermal routes.

Because TCE is an uncharged, nonpolar and highly lipophilic molecule, gastrointestinal absorption is extensive and occurs by passive diffusion (31).

In humans, TCE is known to be highly and rapidly absorbed by inhalation (25–55%) (33), with a high uptake during the first few minutes and steady-state blood levels reached within 2 hours. This high absorption rate is due to the high blood–air partition coefficient, which ranges from 9 to 15 (34). The absorption after inhalation is also high in animals, but significant differences in blood–air partition coefficient exist between species (31).

TCE can also penetrate intact human skin. Four human male volunteers had TCE blood concentrations of 2 mg/l immediately following the immersion of one hand in TCE for 30 minutes. Levels fell to 0.34 mg/l 30 minutes after the end of the immersion period and to 0.22 mg/l after 60 minutes (35).

Distribution

After absorption, TCE is widely distributed in the body via the circulatory system. Because of its high liposolubility, it is predominantly found in adipose tissue and then in the liver, kidneys, cardiovascular system and nervous system (2). TCE crosses the blood–brain barrier and the placenta (36). It has been shown in lactating rats that TCE and its metabolite trichloroacetic acid are excreted in milk. In goats, TCE and its metabolite trichloroethanol were found to be transferred to milk to a slight degree only (37).

Biotransformation

In humans, 40–75% of the retained dose of inhaled TCE is metabolized. The metabolic pathways are qualitatively similar in all species. TCE is metabolized to multiple metabolites either locally or in systemic circulation (e.g. in the liver and by Clara cells in the lung). Many of these metabolites are thought to have toxicological importance (38). The general pattern of enzymatic metabolism occurs through two main pathways: oxidation via the microsomal mixed-function oxidase system (CYP450) and, to a lesser extent, conjugation with glutathione (32).

  • TCE is principally and rapidly transformed by CYP 2E1 (39) into an epoxide intermediate, which spontaneously rearranges to trichloroacetaldehyde and then chloral hydrate. Chloral hydrate acts as a substrate for alcohol dehydrogenase and chloral hydrate dehydrogenase, leading to the formation of trichloroethanol and trichloroacetic acid, respectively. The main metabolites, which are primarily present in the urine, are therefore trichloroethanol, its glucuronide conjugate and trichloroacetic acid. Metabolism appears to be qualitatively identical, irrespective of the exposure route (2).
  • Other minor metabolites of TCE have been identified, including the mercapturic acid N-acetyl-S-(dichlorovinyl)-L-cysteine (DCVC), which is formed in the kidneys from the glutathione conjugate of TCE (previously formed in the liver as a minor biotransformation product). The presence of DCVC in the urine has been demonstrated in rats and also in workers exposed to TCE (37) (Fig. 8.1).
Fig. 8.1. Metabolism of trichloroethylene.

Fig. 8.1

Metabolism of trichloroethylene. Sources: ATSDR (2); Chiu et al. (32).

The metabolism of TCE is quantitatively dependent on the concentration and species tested: it is concentration-dependent in rats but not in mice. In humans, no saturation has been demonstrated. As a consequence of differences in blood flow and overall metabolic rate, species differences exist in the fraction of administered dose of TCE that is available for conversion to toxic metabolites in the target organs (31). In mice, oxidative metabolites are formed in greater quantities than glutathione conjugate metabolites, and dichloroacetic acid is produced to a very limited extent relative to trichloroacetic acid, while most S-(1,2-dichlorovinyl)glutathione (DCVG) is converted into DCVC (40).

TCE is well metabolized by human hepatocytes in culture, with a Km of 266 (± 202) ppm and a Vmax of 16.1 (± 12.9) nmol/hour per 106 viable hepatocytes. Lipscomb et al. (41) predicted a Vmax of approximately 1400 nmol/hour per gram of human liver.

The link between the various metabolites of TCE and diverse types of toxicity is known to be highly complex, making understanding of the toxicological mechanisms of action more complicated. Animal-to-human extrapolation is a source of a high level of uncertainty (42). There is inconclusive evidence suggesting that the glutathione biotransformation route, leading to DCVC production, is more important in humans than in rodents (43).

When ten volunteers were exposed to 250–380 ppm of TCE for 160 minutes, 16% of the retained TCE was eliminated through respiration after exposure. Trichloroacetic acid excretion in females was 2–3 times more than in males for the first 24 hours. However, twice as much trichloroethanol was excreted in males. These observations suggest a sex difference in human metabolism (Nomiyama & Nomiyama 1971, cited in HSDB (36)).

Elimination

In humans and animals, non-metabolized TCE can be eliminated via expired air (11–40%). The main metabolites are eliminated by the kidneys: urinary elimination of trichloroethanol and trichloroacetic acid is complete 5 and 13 days, respectively, after the end of exposure (35,44). In humans, the half-times for renal elimination of trichloroethanol and its glucuronide are about 10 hours. Urinary excretion of trichloroacetic acid is slower, with a reported half-time of about 52 hours (2). Male volunteers were administered chloral hydrate in three separate experiments. Chloral hydrate, dichloroacetic acid, trichloroacetic acid, and trichloroethanol and its glucuronide were measured in blood and urine over a 7-day period. Trichloroacetic acid had the highest plasma concentration and the largest area under the curve of any metabolite. The trichloroacetic acid elimination curve displayed an unusual concentration–time profile that contained three distinct compartments. This complex elimination pattern may result from the enterohepatic circulation of trichloroethanol glucuronide and its subsequent conversion to trichloroacetic acid, as shown in rats (45).

Biomarkers of human exposure

Biomonitoring of TCE is possible by measuring levels of the parent compound or its main metabolite, trichloroacetic acid, in expired air, blood and urine.

Several studies have demonstrated a correlation between levels of TCE in ambient air and in exhaled air (2). Following inhalation exposure to TCE, 10–11% of the absorbed dose is found in expired air as TCE and 2% is eliminated as trichloroethanol.

A linear correlation has been reported between the inhalation exposure of TCE and the urinary levels of trichloroethanol and trichloroacetic acid. In a kinetics study in male volunteers, trichloroacetic acid had the highest plasma concentration and the largest area under the curve of any metabolite (45). Because urinary trichloroacetic acid has a longer half-life than trichloroethanol, it better reflects long-term exposure, whereas urinary trichloroethanol reflects recent exposure (2). The American Conference of Governmental Industrial Hygienists adopted biological exposure indices for TCE based on blood concentrations of free trichloroethanol and trichloroacetic acid and trichloroethanol in urine (46). It should be noted, however, that there is great inter-individual variability in the concentrations of trichloroethanol and trichloroacetic acid in urine and that trichloroacetic acid is not specific to TCE exposure.

Although some studies have shown that protein and DNA adducts may form with chlorinated hydrocarbons (37), their application has not been validated sufficiently to justify their use as biological markers of exposure (47). Some researchers have developed methods to interpret biomarkers by reconstructing human population exposures (48,49). These methods are based on PBPK models (see below) combined with Monte Carlo or Bayesian analysis and estimate TCE environmental concentrations based on known concentrations in blood. This approach involves the interpretation of human biomonitoring data and a possible comparison with health-based exposure guidelines.

Physiologically based pharmacokinetic modelling

Efforts to develop physiologically based pharmacokinetic (PBPK) models have led to an improved assessment of TCE. Several PBPK models have been proposed. They focus on descriptions of both TCE and its major oxidative metabolites in humans (trichloroacetic acid, trichloroethanol and its glucuroconjugate) (see Chiu et al. (32) for additional discussion). Several families of PBPK model are available:

  • Fisher models permit the modelling of liver cancer risks following oral and inhalation exposure of TCE and the formation of metabolites in the liver in humans and mice. None of these models consider renal metabolism (and, therefore, the glutathione pathway in particular), which may play an important role in toxic signs in humans (5054).
  • The Clewell model is more complex than the Fisher models, since it includes sub-models for the main metabolites and for the three target organs demonstrated during toxicity studies in animals (lung for trichloroacetaldehyde, kidney for dichlorovinylcysteine and liver for trichloroacetaldehyde, di- and trichloroacetic acids, trichloroethanol and its conjugate). This model takes into account inhalation and oral exposure, along with hepatic and renal metabolism (55).
  • The Bois model calibrates the existing models of Fischer and Clewell with new toxicokinetic data and includes a Bayesian statistical framework to bring in issues on variability and uncertainty for each parameter (56,57). More recently, other researchers have shown that a combination of Bayesian approaches and PBPK analysis provides better predictions and yields an accurate characterization of the uncertainty in metabolic pathways for which data are sparse (58).
  • Combining the Fischer and Clewell models, and considering the reassessment of the parameters of these models conducted by Bois et al. (56,57) using Bayesian methods, the United States Air Force proposed, in 2004, a harmonized model for use in mice, rats and humans (32,59).

PBPK models have largely been applied in risk assessment to predict dose metrics, toxicity or guideline values. Yoon et al. (60) found that liver-only metabolism may be a reasonable simplification for PBPK modelling of TCE to predict dose metrics. Simmons et al. (61) explored the relationship between measures of internal doses of TCE and neurotoxic outcomes in rats. Another application is based on the time course of TCE in blood and brain of rats and humans to adjust duration for acute guidelines in place of the Haber's law (6264). Hacks et al. (65) used a Bayesian approach to reduce uncertainty in dose metric prediction of TCE and its metabolites (particularly trichloroacetic acid and trichloroethanol). More recently, Evans et al. (66) examined the question of whether the presence of trichloroacetic acid in the liver is responsible for TCE-induced hepatomegaly in mice, and concluded that oxidative metabolites, in addition to trichloroacetic acid, are necessary contributors.

Despite their continuous development, none of these PBPK models currently incorporates all exposure routes or all the possible toxic effects. Similarly, the dose measurement to be used in these models is not clearly explained, i.e. whether it should be area under the curve, cumulative dose or maximum concentration. The development of more complex models is notably linked to advances in scientific knowledge concerning understanding of the mechanism(s) of toxic action of TCE.

Health effects

Effects in experimental animals and in vitro test systems

Non-carcinogenic effects

In animals, the major effect of acute exposure includes a state of excitation followed by CNS depression and drowsiness. This depression is marked by a loss of reflexes and motor coordination, potentially progressing to coma. The LC50 values are 142 g/m3 (1 hour) and 71 g/m3 (4 hours) in rats and 46 g/m3 (4 hours) in mice, indicating a low acute inhalation toxicity. A transient hepatic toxicity has been observed in rodents. A specific pulmonary toxicity (to Clara cells) has been demonstrated in mice and a transient specific nephrotoxicity is demonstrated when the metabolism is saturated in rats (35,44).

Effects on the CNS have also been demonstrated during subchronic and chronic inhalation exposure. In rats, a NOAEL of 2700 mg/m3, based on an increase of latency in visual discrimination tasks, was identified after 18 weeks of inhalation exposure (67). In a 16-week study in rats, brainstem auditory-evoked response potentials were depressed at test concentrations as high as 8640 and 17 280 mg/m3(68). Electroencephalograph changes have also been reported in rats exposed to up to 50 ppm for 6 weeks (69). In rabbits, neuro-ophthalmological reversible modifications were observed during a 12-week period of inhalation exposure at 1890 and 3780 mg/m3(70). Based on the hypothesis that organic solvents can promote noise-induced hearing loss, Vyskocil et al. (71) reviewed the effects of low-level exposure to TCE on the auditory system. In rats, TCE affects the auditory function mainly in the cochlear mid- to high-frequency range, with a LOAEL of 2000 ppm. Supra-additive interaction after exposure to noise and TCE has been reported (71). A recent animal study was conducted to determine whether TCE exposure is neurotoxic to the striatonigral dopamine system that degenerates in Parkinson's disease. The study showed that oral administration of TCE for 6 weeks leads to a complex 1-mitochondrial impairment in the midbrain and a striatonigral fibre degeneration and loss of dopamine neurons (72).

Transient hepatic hypertrophy has also been observed, but the results of studies are equivocal and its toxicological significance is not clear. Inhalation exposures to 2000 ppm TCE show elevated plasma alanine and aspartate aminotransferase activities and liver histopathological abnormalities in mice. At the same dose, TCE significantly upregulates PPARα (39). Sano et al. (73) demonstrated distinct transcriptional profiles and differences in biological pathways between rats and mice, suggesting species differences in liver toxicity.

An increase in kidney weight has been demonstrated in rats, but without any particular associated histological changes (74). Megalonucleocytosis has been observed and a NOAEL has been defined at 115 mg/m3(75). The relative importance of metabolism of TCE by the CYP450 and glutathione conjugation pathways in the acute renal and hepatic toxicity of TCE was studied in isolated cells and microsomes from rat kidney and liver. Increases in cellular glutathione concentrations increased TCE cytotoxicity in kidney cells but not in hepatocytes, consistent with the role of glutathione conjugation in TCE-induced nephrotoxicity. In contrast, depletion of cellular glutathione concentrations moderately reduced TCE-induced cytotoxicity in kidney cells but increased cytotoxicity in hepatocytes, consistent with the different bioactivation pathways in kidney and liver (76). The involvement of CYP450 in TCE-induced hepatotoxicity was also studied in mice and its major role in the hepatotoxicity of TCE confirmed (39). Recently, Khan et al. (77) showed that TCE causes altered carbohydrate metabolism and suppresses the antioxidant defence system in rats. These results are consistent with the hypothesis that TCE induces oxidative stress in kidney and other tissues.

Immune disorders have been observed in rats (78,79). A decline of CD4+ in T lymphocytes was observed after intradermic administration of TCE, but no significant concentration differences in IFN-gamma and IL-4 were found between TCE-treated animals and controls (78). Other animal experiments suggest that immunotoxicity is mediated via haptenization of macromolecules and that haptenized proteins may act as neo-antigens that can induce humoral immune response and T-cell-mediated hepatitis in mice. Further observations suggest that TCE promotes inflammation in the liver, pancreas, lung and kidney, which may lead to SLE-like disease (80). DCA could be involved in the immune disorders and hepatotoxicity induced by TCE (81,82).

Moreover, Tang et al. (83) recently found that TCE can induce non-dose-related hepatitis classified as a delayed-type hypersensitivity at low doses in guinea-pigs exposed via intradermal injection (below the dose causing liver injury) and with different histopathological changes. TCE exposure in mice generates a time-dependant increase in antibodies specific for liver proteins in mice, upregulates the methionine/homocysteine pathway in the liver, and alters the expression of selective hepatic genes associated with immunity (84). Moreover, TCE enhances histamine release from antigen-stimulated mast cells and inflammatory mediator production (79). Other researchers suggest that protein oxidation (carbonylation and nitration) could contribute to TCE-induced autoimmune response because an increase in oxydatively modified proteins is associated with significant increase in cytokines. These first results observed in mice require further study (85).

TCE could be skin irritant: a recent study investigating acute and cumulative TCE topical treatment in BALB/c hairless mice showed skin reaction (erythema and oedema) and histopathological changes (hyperkeratosis and inflammatory cell infiltrates) (86).

Based on in vivo and in vitro studies, the US National Research Council (87) concluded that exposure to TCE disrupts spermatogenesis, reduces male fertility and the fertilization capacity of spermatozoa, and reduces the capacity of oocytes to be fertilized in females. Studies conducted in male Wistar rats exposed by inhalation to 376 ppm of TCE for 12 or 24 weeks (88,89) demonstrated a significant reduction in the number and motility of spermatozoa and in steroid enzyme activity (dehydrogenases), with a reduction of testosterone levels in the sperm. This was associated with a reduction of absolute testicular weight and histopathological changes. The fertility of these male rats was reduced when mating was performed with untreated females. Testicular cholesterol was elevated in exposed rats, suggesting that TCE acts on the biosynthesis of testosterone in the testis. Histological changes have also been shown (in spermatogonia and spermatids, seminal tubes and Leydig cells) but the reversibility of the effects has not been studied. A study in mice exposed to 1000 ppm for 1–6 weeks did not demonstrate any effects on testis or sperm (90) but mating with non-exposed females resulted in a significant decrease in the rate of fertilized oocytes after 2 and 4 weeks. Furthermore, an in vitro assay demonstrated a reduction in the number of spermatozoa per oocyte after treatment with 0.1–10 μg/l chloral hydrate or trichloroethanol. This suggests that the metabolites (in particular chloral hydrate) were responsible for the reproductive toxicity of TCE. More recently, Kan et al. (91) showed epithelial damage (vesiculation in the cytoplasm, disintegration of basolateral cell membranes and sloughing of epithelial cells) in the epididymis of mice exposed to 1000 ppm TCE for 1–4 weeks. Further experiments (92,93) demonstrated that TCE could also cause reproductive toxicity in female rats, with a decrease in spermatozoon penetration and oocyte fertilization and reduced membrane-binding protein in female rats treated with TCE (2 weeks' administration of drinking-water containing 0.45% TCE). These effects also appear to be dependent on metabolic activation by CYP2E1 (without metabolite(s) being involved) and on glutathione conjugation to DCVC (94). The real impact of the biological effects observed on the reproductive function of the animals is not known, nor is the transposability of these effects to human reproductive function. Finally, these observations suggest that enzyme induction and oxidative metabolism may play a role in the reproductive toxicity of TCE. Oxidative metabolites of TCE are formed in the mouse epididymis, resulting in epididymal damage, and at systemically toxic high doses TCE may adversely affect the maturation of sperm and decrease sperm motility (95). Lamb & Hentz consider that protection against systemic toxicity should also protect against adverse effects, including male reproductive toxicity (95).

One study has suggested the possibility of an increased incidence of malformations in rat pups after oral exposure via drinking-water. An increase in the incidence of cardiac and eye malformations was observed in pups from dams exposed to 0.18 and 132 mg/kg body weight per day before and during gestation (3 months before or 2 months before and during pregnancy) or only 132 mg/kg body weight per day in dams exposed only during gestation, without maternal toxicity (33). However, a recent evaluation by Williams & Desesso (96) pointed out the maternal toxicity associated with this birth defect. The mechanism of action is not known but it may involve metabolism by CYP450 2E1 and dichloroacetic and trichloroacetic acids. No malformation has been reported in inhalation studies in rats and other oral animal studies have not demonstrated conclusive results (33,87,97). A recent study, in which mice were exposed to 31 mg/kg body weight per day via drinking-water from gestational day 1 to postnatal day 42, showed that developmental and early life exposure of TCE could modulate the immune function of pups and may be associated with neurodevelopmental disorders (98).

A summary of relevant animal inhalation studies for subchronic and chronic exposures to TCE, indicating derived NOAELs and LOAELs, is shown in Table 8.1.

Table 8.1. A review of animal inhalation studies for subchronic and chronic exposure.

Table 8.1

A review of animal inhalation studies for subchronic and chronic exposure.

Carcinogenic effects

Exposure to TCE was responsible for an increased incidence of liver tumours in male Swiss mice and B6C3F1 mice of both sexes exposed by the inhalation route to 600 ppm for 78 weeks. Pulmonary tumours were also increased in female B6C3F1 and male Swiss mice at 600 ppm, but not among the male B6C3F1 mice (75). Other studies have demonstrated a significant increase in pulmonary adenocarcinomas in female ICR mice exposed to 150 or 400 ppm for 104 weeks (100). In male Sprague-Dawley rats, inhalation exposure to 600 ppm of TCE for 104 weeks led to a dose-dependent increase in Leydig cell tumours in the testis and a marginal increase in renal tumours (adenocarcinomas of the renal tubules). Finally, in mice, rats and hamsters exposed to 100 and 500 ppm TCE for 18 months, the only significant increase in tumour incidence was for malignant lymphomas in female mice (75). The results of experimental studies are presented in Table 8.2.

Table 8.2. Review of inhalation carcinogenic studies in animals.

Table 8.2

Review of inhalation carcinogenic studies in animals.

The metabolism of TCE doubtless plays a very important role in its mechanism of carcinogenic action. Metabolic pathways have largely been described (33,101103). The research to date indicates that TCE-induced carcinogenesis is complex, involving multiple carcinogenic metabolites acting in various ways. Past explanations, such as the hypothesis linking mouse liver tumours to peroxisome proliferation, are not consistent with the whole of the data, and more complex hypotheses have been formulated (38). A plausible mode of action is that TCE induces liver tumours through trichloroacetic acid and DCA modifying the cell signalling systems that control cell division rate and cell death (103105). This hypothesis suggests that humans are likely to be much less responsive than mice and that carcinogenic effects are unlikely to occur at low environmental exposures. The induction of pulmonary tumours in mice may be linked to the fact that Clara cells rapidly metabolize TCE into chloral hydrate, via CYP450 2E1, leading to pulmonary accumulation of this metabolite, which ultimately produces cell changes and compensatory proliferation. But other mechanisms of action may be involved, particularly since chloral hydrate is probably genotoxic and, at high doses, clastogenic. In rats, Clara cells are capable of metabolizing chloral hydrate into trichloroethanol. In humans, the capacity of the lung to transform TCE into chloral hydrate is thought to be negligible and, consequently, the mechanism of pulmonary carcinogenesis demonstrated in mice may be specific to mice.

Finally, renal tumours in male rats may be linked to cytotoxicity and persistent cell regeneration. Conjugation to glutathione and the involvement of beta-lyase in the renal tubules may lead to the formation of nephrotoxic and probably genotoxic reactive metabolites, in particular DCVC and DCVG (101). Studies have demonstrated that TCE induces mutations in the VHL (Von Hippel-Lindau) tumour suppressor gene in the cells of renal carcinomas in patients with this cancer (106,107), but Charbotel et al. (108) did not confirm the association between the number and type of VHL gene mutations and exposure to TCE.

A second mechanism may involve increased secretion of formic acid, leading to a disruption in the detoxification mechanism by methionine. The mechanism of carcinogenic action leading to the development of renal tumours in rats is less clearly determined but its transposability to humans is questionable. There is no tested hypothesis to take into account the mechanistic aspects of the induction of malignant lymphomas in mice and testis tumours in male rats. These considerations, and the effects observed in humans, justify a cautious attitude regarding the extrapolation to humans of results observed in animals.

In conclusion, despite the numerous limitations to confident interpretation of some of the data (e.g. the response of animals differs depending on sex for renal tumours and depending on species for hepatic and pulmonary tumours), animal evidence is deemed sufficient for evaluating the carcinogenic effect of TCE. The European Commission's risk assessment report on TCE concludes that studies provide clear evidence that TCE is carcinogenic in rats and mice through oral and inhalation exposure (1).

Genotoxicity

In Europe, TCE has been classified since 2001 as mutagenic category 3 (risk phrase R68) under Directive 67/548/CEE. Several reviews on the mutagenicity of TCE are available in the literature (2,3,110). TCE seems to be genotoxic in vitro: the Ames test and in vitro mouse lymphoma test have shown a weak positive response with activation, but this characteristic is equivocal in vivo. A recent paper by Hu et al. (111) assesses the in vitro genotoxic effects of TCE and the underlying mechanisms using human HepG2 cells: TCE exposures (0.5–4 mM) caused positive response in comet assay and micronuclei assay. These results suggest that TCE causes DNA strand breaks and chromosome damage in hepatocytes. In another study evaluating the in vivo genotoxicity of TCE (99.5% purity) by inhalation (500–1000–2000 ppm) and DCVC (> 95% purity) by oral gavage (1–10 mg/kg), using the comet assay to assess DNA breakage in the proximal tubules of kidneys, rats were exposed at dose levels in excess of those that produced renal tumours. TCE gave a clearly negative response in the assay at all dose levels. DCVC gave a negative response at the lower dose level. At the higher dose level, there was limited evidence of DNA damage in a small number of animals. The authors suggest that the mechanism for induced renal tumours is non-genotoxic (112).

Finally, while the available data on genotoxicity do not show a consistent pattern, the results indicate that TCE has a weak genotoxic action causing numerical chromosomal aberrations (aneuploidy) in vivo and probably DNA strand breaks in hepatocytes in vitro.

Currently, the various hypotheses suggested do not enable accurate identification of the key events responsible for the development of cancers at different sites (lungs, liver, kidneys, etc.). The ambiguity concerning the role of active metabolites, and the various mechanisms of action and effects, lead to a high level of caution when transposing animal data to humans (differences in sensitivity, quantitative differences in kinetics between species and as a function of exposure levels) (38).

In conclusion, the mechanism of carcinogenic action of TCE can be attributed to numerous mechanisms, involving both non-genotoxic and genotoxic phenomena. It is thus prudent to consider that TCE can induce a risk of cancer in humans based on a non-threshold hypothesis (42).

Unit risk values based on animal data have been estimated using the linearized multistage model on carcinogenicity data from mice and rats and using the most sensitive tumour type for which there is sufficient evidence. Unit risk values were 9.3 × 10−8 per μg/m3 and 1.6 × 10−7 per μg/m3 for pulmonary adenomas in B3C6F1 mice and Swiss mice, respectively, and 4.3 × 10−7 per μg/m3 for Leydig cell tumours in rats. The most protective unit risk (4.3 × 10−7 per μg/m3) was used to derive health-based guideline values for TCE in air in Europe (37).

Effects in humans

Non-carcinogenic effects

In humans, the main target following the inhalation of high concentrations of TCE is the CNS, as is observed in animals. Neurological damage, especially affecting the optic and trigeminal nerves, has been reported following accidental exposure.

The acute neurological effects of TCE may be more related to maximum blood concentrations than to “area under the curve”. In rats, the peak TCE concentration inducing toxicity appears to be higher than in humans, suggesting that humans are more sensitive than animals for these neurological effects (42). The neurological effects have been observed at concentrations from 270 mg/m3 (changes in visual and auditory potentials) to approximately 600–1000 mg/m3 (decreased psychomotor performances) over several hours (37). Cardiac effects (ventricular fibrillation) may also cause death following massive exposure (42). A recent accidental inhalation of TCE during the cleaning of a metal-degreasing machine produced a reversible kidney injury (urinary proteins and enzymes 7 and 74 hours after exposure) in a 54-year-old man. TCE and trichloroacetic acid had peak blood concentrations at 11 and 62 hours after poisoning, respectively (113). Another case report showed acute liver and kidney failure followed by severe brain oedema and death in a 27-year-old man, probably caused by abuse of glue containing TCE (114).

Effects on the CNS have also been demonstrated during chronic inhalation exposures. The majority of studies in humans describe the symptoms following acute exposure, but these are often of inadequate quality (absence of data on exposure or on confounding factors). More discrete neurological effects such as motor incoordination have also been observed for exposures of 87, 60 and 38 mg/m3, respectively (102) but there is no convincing evidence of TCE-induced hearing losses in workers. No studies on ototoxic interaction after combined exposure to noise and TCE have been identified in humans (71). Analysis of a cluster of 30 workers with neurological disease who were chronically exposed to TCE showed that the 3 workers with workstations nearest the TCE source had Parkinson disease. The authors suggest that TCE is a probable risk factor for Parkinsonism (72).

Renal and pulmonary damage in humans following TCE exposure is absent or very slight, but transient effects on the liver have been observed (37). Levels of total cholesterol and high-density lipoprotein cholesterol increased slightly with dose, but without modification of serum enzyme activity. It was suggested that exposure to TCE can influence hepatic function (115). However, all the studies suffer from major methodological limitations, particularly in terms of characterization of exposures. In addition, individuals exposed to TCE were also exposed to other solvents.

Some authors suggest that TCE induces and exacerbates autoimmunity. This is supported by animal experimentations and observations of SLE and other immunological disorders in occupationally exposed human (82). However, idiosyncratic generalized skin disorders complicated by hepatitis have rarely been observed in populations occupationally exposed to TCE in factories where TCE metabolites could extensively accumulate (urinary trichloroacetic acid concentrations from 318 to 1617 mg/l) (116). This is consistent with a recent study by Xu et al. (117) in which TCE induced hypersensitivity dermatitis and liver dysfunction in Chinese workers exposed to 18–683 mg/m3 for an average of 38.2 days (range 5–90 days). Liu et al. (118) found autoantibodies in sera collected from patients who (had) suffered from TCE-induced dermatitis. These antibodies could perhaps be used to understand underlying mechanisms in the immunotoxicity of TCE.

The effect of inhaled TCE on fertility in humans has not been studied. The most recent studies have demonstrated a number of modifications in endocrine function, revealed by measurement of steroid hormones in particular, with changes observed following exposure to 60 mg/m3 TCE (119121), but the toxicological significance of these observations has not been investigated.

Epidemiological studies have been carried out in occupational environments to investigate any link between exposure to degreasing solvents (including TCE) and pregnancy outcomes. Some of these have reported increased risks for cardiac anomalies, with OR ranging from 3.4 (95% CI 1.6–6.9) to 6 (95% CI 1.7–21.3) (122,123). But it is not possible to reach any conclusion with respect to the precise role of TCE. Pregnancy outcomes have been studied in several cohorts in the general population exposed via the oral route (drinking-water) in the United States. Developmental abnormalities (cardiac, neural tubes, cleft palate, eye and ear malformations), perinatal deaths and low birth weights were observed (42,87,124128) but the presence of possible bias or misclassification precludes confident conclusions being made. An interaction between maternal age and TCE exposure in increasing congenital heart defects has been observed by Yauck et al. (129), although the mechanism by which this might occur is unknown. Finally, the National Research Council suggested that epidemiological observations concerning malformations (particularly cardiac ones) and delayed intrauterine growth in humans exposed to TCE are consistent with the animal studies and are supported by mechanistic studies and a relative agreement in the type of malformations.

However, to date, no definitive conclusion has been put forward for humans and it is not possible to extract from these studies either a well-defined dose–response relationship or a LOAEL for assessing the risk of TCE, particularly since the populations are often concomitantly exposed to several toxic substances (halogenated solvents, metals, etc.) (87).

Carcinogenic effects

The results of the Finnish cohort study, in which 2050 men and 1924 women were exposed to TCE and other solvents in the context of their work (130), demonstrated a statistically significant increase in non-Hodgkin's lymphoma and cervical cancer, with a significantly higher risk in the individuals with the highest urinary trichloroacetic acid concentrations (signal-to-interference ratio 4.4; 95% CI 1.4–10.1), and an increase in liver cancers for workers exposed for more than 20 years (RR = 6.1, 95% CI 2.8–17.7). Kidney cancers were not significantly increased. However, the exact exposure duration was not known and the workers were exposed to other solvents (although the estimates were adjusted to the urinary trichloroacetic acid concentrations). In the same cohort, the risk of liver cancer was increased among male printers, lacquerers and varnishers exposed to chlorinated hydrocarbons (RR = 2.65, 95% CI 1.38–5.11). The authors suggest a role of TCE, which is consistent with previous data (131).

The results of the Swedish cohort study, in which 1670 workers (1421 men and 249 women) were exposed to TCE (132), demonstrated that mortality and morbidity from cancer were not significantly higher in these exposed individuals than in the general population. The majority of workers had urinary trichloroacetic acid levels below 50 mg/l, which may correspond – according to the authors – to an exposure of approximately 20 ppm.

A study conducted on a cohort of 14 457 American aircraft maintenance workers exposed to multiple solvents, including TCE, demonstrated a non-significant increase in mortality due to liver cancer, kidney cancer and non-Hodgkin's lymphoma: a statistically significant increase was observed for multiple myelomas (SMR 236; 95% CI 87–514) and non-Hodgkin's lymphoma (SMR 212; 95% CI 102–390) in white women, and for cancers of the bile duct and liver in white men who died after 1980 (SMR 358; 95% CI 116–836). Exposures were classified according to certain indices (as a function of occupational category) and did not therefore permit a quantitative approach. When only individuals exposed to TCE were examined (6929 people), no significant association was found between the additional risk of cancer and TCE measurements (133).

A recent case-control study in Germany analysed the relationship between exposure to organic solvents (including TCE) and malignant lymphoma in 710 patients. A statistically significant association was found between high exposure to chlorinated solvents and malignant lymphoma (OR 2.1; 95% CI 1.1–4.3). When TCE only is considered, this trend persists (borderline statistical significance) (134).

Other studies conducted in general populations exposed to TCE via drinking-water have demonstrated associations between (increased) incidence of leukaemia or non-Hodgkin's lymphoma and TCE exposures, which is consistent with the data from occupational cohorts. Scott & Chiu (135) reviewed recently published scientific literature examining cancer and TCE exposure and suggested that the studies appear to provide further support for the kidney, liver and lymphatic system as targets of TCE toxicity.

All the retrospective cohort studies conducted on TCE have methodological limitations linked either to the absence of quantification of exposures to TCE, to potential co-exposures not taken into account in occupational environments, or to the low number of subjects studied. Nevertheless, some epidemiological studies have measured trichloroacetic acid in urine, which can be directly related to TCE exposure (130). The strongest associations between TCE exposure and human cancer are for the kidney, liver and lympho-haematopoietic system, sites where TCE causes cancer in rats and mice (102). These aspects of biological plausibility and coherence suggest that a cause-and-effect association between TCE exposure and cancer in humans is credible, even if the interpretation of individual studies may be difficult.

However, several recent meta-analyses have been conducted that did not confirm previous findings. A meta-analysis of 14 occupational cohort and four case-control studies of workers exposed to TCE investigated the relationship between TCE exposure and risk of non-Hodgkin's lymphoma. The comparisons carried out by the authors did not indicate exposure–response trends, suggesting insufficient evidence of a causal link between TCE exposure and non-Hodgkin's lymphoma (136,137). Alexander et al. (138) found the same results in analysing occupational studies of TCE exposure and liver/biliary tract cancer. The main conclusions drawn are that exposure to solvents may cause cancer in humans and that TCE is likely to be one of these, but a number of challenging issues need to be considered before concluding clear causal relationships between TCE exposure and cancer (139).

Thus, considering the bias and confounding in epidemiological studies, human evidence of the carcinogenicity of TCE can be considered limited. IARC has classified TCE as probably carcinogenic to humans (Group 2A) based on sufficient evidence in animals and limited evidence in humans (3).

Based on data presented by the USEPA in its health risk assessment of TCE (102), and particularly the cancer potency values, Lewandowski & Rhomberg (140) proposed a method for selecting the most appropriate carcinogenic inhalation unit risk estimate for TCE. The method is based on an in-depth analysis of the key studies used to derive unit risks in both animals and humans (protocol, rigour, statistical power, characterization of exposure, confusion factors, critical effects, etc.). The authors evaluated the validity of the studies (suitability of the protocol and dose–response relationships for a quantitative assessment) and the plausibility of the effects (with the use of Hill's criteria). These considerations led to the choice of the unit risk derived from the Finnish cohort study, based on the increase in the incidence of hepatic tumours. The selected unit risk was 9 × 10−7 per μg/m3.

Sensitive populations

An extensive review of factors that may affect risk of exposure to TCE, with a particular examination of age (children), genetics, sex, altered health state, co-exposure to alcohol and enzyme induction, was published in 2000 (141).

Since the metabolism of TCE is largely implicated in its toxic mechanisms of action, all the known polymorphisms, and particularly those concerning the CYP2E1, glutathione-S-transferase (GST) and N-acetyltransferase enzymes, are liable to modify individual sensitivity to this substance, although it is currently impossible to accurately quantify the scale of this modification or the number of people affected by these polymorphisms. In the example of polymorphism of GST, it appears that some subgroups of the population have a risk of kidney cancer that is four times higher than others (102). However, a recent study of about 134 renal cell cancer cases from Brüning et al. (142) does not confirm the hypothesis of an influence of the deletion polymorphisms of GST on renal cell cancer development due to exposure to TCE (143). With respect to the polymorphism of CYP450 2E1, Pastino et al. (141) reported that a 10- to 50-fold variability in the protein or its activity has been observed in humans.

Merdink et al. (45) showed that individuals with an impaired capacity for glucuronidation may be very sensitive to the CNS-depressant effects of high doses of chloral hydrate, which are commonly attributed to plasma levels of TCE.

Individuals with hepatic and/or renal failure may constitute a more sensitive population owing to reduced metabolism of TCE and/or a reduction in the elimination of its toxic metabolites, whether these disturbances be genetic, environmental (alcohol, medicinal products, etc.) or secondary to a disease. Individuals with a history of cardiac arrhythmias may be more susceptible to high-level TCE exposure (2). The provisional version of the 2001 assessment made by the USEPA also cites diabetics among sensitive populations owing to their specific susceptibility to neuropathies and certain cancers, and the specific effects of TCE on the metabolism of carbohydrates and cell signalling (102).

Finally, according to ATSDR (2), people with a high consumption of alcohol or taking disulfiram may be more sensitive to the neurological effects of TCE, owing to an interaction process.

Health risk evaluation

Critical health outcomes

TCE is a chlorinated solvent. Its main health effects are neurotoxic and carcinogenic. Immunotoxic, hepatic and developmental effects are also reported.

Neurotoxic effects

Effects on the CNS (damage affecting the optic and trigeminal nerves) have been observed in humans and animals exposed to high acute (600–1000 mg/m3) or moderate chronic (38–87 mg/m3) occupational levels. Sufficient evidence exists to conclude that there is an association between TCE and neurotoxic effects.

Immunotoxic effects

The evidence is suggestive for an association between TCE exposure and the exacerbation or induction of autoimmunity. Several mechanistic hypotheses are suggested in rodents but further research is needed before firm conclusions can be reached. Recent studies in humans confirm the possibility of immune disorders in individuals exposed occupationally to high-to-moderate levels of TCE (18–683 mg/m3). However, not enough human studies are available to allow a conclusion to be drawn on causality, especially because the human immune response varies greatly among individuals. It is concluded that there is limited evidence of an association between immunological effects and TCE exposure.

Hepatic effects

Transient hepatic hypertrophy has been observed in rodents, but the results of studies are equivocal. The human epidemiological studies suffer from methodological limitations, particularly in terms of characterization of exposure. In addition, individuals exposed to TCE were often also exposed to other solvents. It is concluded that there is limited evidence of an association between hepatic effects and TCE exposure.

Developmental effects

Developmental effects, notably cardiac and eye malformations, have been reported in rodents but the results are inconsistent (possible maternal toxicity, positive results only in oral studies in rats). Occupational studies in humans suggest a link between the use of degreasing solvents and adverse pregnancy outcomes. Epidemiological studies in the general population suggest malformations, perinatal death and low birth weight, but possible bias and exposure misclassification prevent firm conclusions being drawn. There is thus insufficient evidence for an association between developmental effects and TCE exposure.

Carcinogenic effects

Animal evidence is sufficient to demonstrate carcinogenic effects of TCE by both oral and inhalation routes, and there is sufficient evidence to conclude that TCE is at least weakly genotoxic. Positive associations have been established between occupational exposure and risks for cancer of the liver, kidney and bile duct and non-Hodgkin's lymphoma. Lung and testis tumours observed in rodents have not been reported in humans but cannot be excluded. The presence of possible exposure misclassification or co-exposure in occupational cohort studies somewhat weakens the confidence in the association. Overall, it is concluded that sufficient evidence exists to suggest an association between TCE exposure and cancer (liver and kidney).

Health relevance of indoor exposures

Since there is sufficient evidence that TCE is a genotoxic carcinogen, all exposures indoors are considered relevant and no threshold can be determined.

Inhalation of TCE is the main route of exposure in the general population. Ambient and indoor air concentrations of TCE are generally less than 1 μg/m3 in European and North American countries. Indoor TCE levels of up to 30 μg/m3 (90th percentile) have been reported during the EXPOLIS study (1998–1999). More recent studies in French dwellings and American office buildings showed lower levels (95th percentile 7.4 and 2.6 μg/m3, respectively).

Consumers may be exposed to TCE by using wood stains, varnishes, finishes, lubricants, adhesives, typewriter correction fluid, paint removers and certain cleaners, where TCE is used as a solvent. Contaminated water or soil may also contribute to indoor pollution through TCE.

Conclusions of other reviews

The previous WHO air quality guideline (37) was based on the unit risk estimate of 4.3 × 10−7 (μg/m3)−1 derived from the increase of Leydig cell tumours in rats.

IARC has considered TCE a probable carcinogen since 1995 (Group 2A, limited evidence in humans but sufficient in animals). The EU classified it as carcinogenic category 2, risk phrase R45 (may cause cancer) in 2001 for the same reasons. The IARC evaluation is based on experimental data and on three human cohort studies conducted in Finland, Sweden and the United Kingdom, which demonstrated an increased risk of several cancers including liver, kidney and bile duct cancers and non-Hodgkin's lymphoma (3). An additional risk for cervical cancer was observed in two of the studies.

Guidelines

The existence of both positive and negative results has in the past led risk assessors to different interpretations of TCE toxicity and to divergent estimates of human cancer risk (144,145). For a health risk evaluation, bearing in mind recent data on a mechanism of action that is not species-specific, the evidence for weak genotoxicity, and the consistency between certain cancers observed in animals and in humans (in particular liver cancer), it is prudent to consider that the carcinogenicity in animals, the positive epidemiological studies and the plausibility of a human cancer risk leads to the recommendation of a non-threshold approach with a risk estimate rather than a safe level.

Therefore, carcinogenicity (with the assumption of genotoxicity) is selected as the end-point for setting the guideline value. The unit risk estimate of 4.3 × 10−7 (μg/m3)−1, derived on the basis of increased Leydig cell tumours (testicular tumours) in rats, is proposed as the indoor air quality guideline. This was also the conclusion of WHO in 2000 (37), the EU in 2004 (1) and the French Agency for Environmental and Occupational Health in 2009 (42).

The concentrations of airborne TCE associated with an excess lifetime cancer risk of 1:10 000, 1:100 000 and 1:1 000 000 are respectively 230, 23 and 2.3 μg/m3.

The guidelines section was formulated and agreed by the working group meeting in November 2009.

Summary of main evidence and decision-making in guideline formulation

Critical outcome for guideline definition

Carcinogenicity (liver, kidney, bile duct and non-Hodgkin's lymphoma), with the assumption of genotoxicity.

Source of exposure–effect evidence

Increased Leydig cell tumours (testicular tumours) in rats provided the basis for calculation of unit risk, applying a linearized multistage model (37).

Supporting evidence

Sufficient evidence exists for an association between TCE exposure and cancer (liver and kidney) (3).

Unit risk derived from a cohort study of occupationally exposed adults (139), based on the increase in the incidence of hepatic tumours, was 9 × 10−7 per μg/m3(140).

Results of other reviews

IARC: Group 2A (limited evidence in humans but sufficient in animals) (3).

EU: carcinogenic category 2, risk phrase R45 (may cause cancer) (1).

Guidelines

Unit risk estimate of 4.3 × 10−7 per μg/m3.

The concentrations of airborne TCE associated with an excess lifetime cancer risk of 1:10 000, 1:100 000 and 1:1 000 000 are respectively 230, 23 and 2.3 μg/m3.

References

1.
European Union risk assessment report Trichloroethylene. Brussels: European Commission; 2004.
2.
Agency for Toxic Substances and Disease Registry (ATSDR) Toxicological profile for trichloroethylene. Atlanta, GA: US Department of Health and Human Services; 1997.
3.
Dry cleaning, some chlorinated solvents and other industrial chemicals Summary of data reported and evaluation. Lyon: International Agency for Research on Cancer; 1995. (IARC Monographs on the Evaluation of Carcinogenic Risks to Humans, Vol. 63)
4.
Trichloroethylene. Geneva: World Health Organization; 1985.
5.
Trichloroethylene. Priority substances assessment report for the Canadian Environmental Protection Act. Ottawa: Environment Canada and Health Canada; 1993.
6.
German Chemical Society. Trichloroethene. Stuttgart: Hirzel; 1994. (BUA Report 95)
7.
Lewis RG, Gordon SM. Sampling for organic chemicals in air. In: Keith LH, editor. Principles of environmental sampling. 2nd ed. Washington, DC: American Chemical Society; 1996. pp. 401–470.
8.
Draft guidance for evaluating the vapor intrusion to indoor air pathway from groundwater and soils (subsurface vapor intrusion guidance). Washington, DC: US Environmental Protection Agency; 2002.
9.
Haddad S, Tardif GC, Tardif R. Development of physiologically based toxicokinetic models for improving the human indoor exposure assessment to water contaminants: trichloroethylene and trihalomethanes. Journal of Toxicology and Environmental Health, Part A. 2006;69:2095–2136. [PubMed: 17060096]
10.
Fan C, et al. Risk assessment of exposure to volatile organic compounds in groundwater in Taiwan. Science of the Total Environment. 2009;407:2165–2174. [PubMed: 19167026]
11.
Sources, emission, and exposure for trichloroethylene (TCE) and related compounds. Washington, DC: US Environmental Protection Agency; 2001.
12.
Adgate JL, et al. Outdoor, indoor, and personal exposure to VOCs in children. Environmental Health Perspectives. 2004;112:1386–1392. [PMC free article: PMC1247565] [PubMed: 15471730]
13.
Adgate JL, et al. Personal, indoor, and outdoor VOC exposures in a probability sample of children. Journal of Exposure Analysis and Environmental Epidemiology. 2004;14:S4–S13. [PubMed: 15118740]
14.
Clayton CA, et al. National Human Exposure Assessment Survey (NHEXAS): distributions and associations of lead, arsenic and volatile organic compounds in EPA Region 5. Journal of Exposure Analysis and Environmental Epidemiology. 1999;9:381–392. [PubMed: 10554141]
15.
Payne-Sturges DC, et al. Personal exposure meets risk assessment: a comparison of measured and modeled exposures and risks in an urban community. Environmental Health Perspectives. 2004;112:589–598. [PMC free article: PMC1241926] [PubMed: 15064166]
16.
Sax SN, et al. Differences in source emission rates of volatile organic compounds in inner-city residences of New York City and Los Angeles. Journal of Exposure Analysis and Environmental Epidemiology. 2004;14(Suppl. 1):S95–S109. [PubMed: 15118751]
17.
Sexton K, et al. Comparison of personal, indoor, and outdoor exposures to hazardous air pollutants in three urban communities. Environmental Science and Technology. 2004;38:423–430. [PubMed: 14750716]
18.
Van Winkle MR, Scheff PA. Volatile organic compounds, polycyclic aromatic hydrocarbons and elements in the air of ten urban homes. Indoor Air. 2001;11:49–64. [PubMed: 11235231]
19.
Weisel CP, et al. Relationship between indoor, outdoor and personal air (RIOPA). Houston, TX: Health Effects Institute and National Urban Air Toxics Research Center; 2005. (Report No. 130, Part 1)
20.
Jantunen MJ, et al. Air pollution exposure in European cities: the EXPOLIS Study. Kuopio: National Public Health Institute; 1999.
21.
Kirchner S, et al. National dwellings survey: report on air quality in French dwellings, Final Report. Paris: Indoor Air Quality Observatory; 2006.
22.
Ohura T, et al. Organic air pollutants inside and outside residences in Shimizu, Japan: levels, sources and risks. Science of the Total Environment. 2006;366:485–499. [PubMed: 16298419]
23.
Building Assessment Survey and Evaluation (BASE) Study. Washington, DC: US Environmental Protection Agency; 2008. [2 July 2010]. http://www​.epa.gov/iaq​/base/voc_master_list.html.
24.
Chao CY, Chan GY. Quantification of indoor VOCs in twenty mechanically ventilated buildings in Hong Kong. Atmospheric Environment. 2001;35:5895–5913.
25.
Eklund BM, et al. Spatial and temporal variability in VOC levels within a commercial retail building. Indoor Air. 2008;18:365–374. [PubMed: 18636973]
26.
Guo H, et al. Risk assessment of exposure to volatile organic compounds in different indoor environments. Environmental Research. 2004;94:57–66. [PubMed: 14643287]
27.
Loh MM, et al. Measured concentrations of VOCs in several non-residential microenvironments in the United States. Environmental Science & Technology. 2006;40:6903–6911. [PubMed: 17153993]
28.
Von Grote J, et al. Reduction of occupational exposure to perchloroethylene and trichloroethylene in metal degreasing over the last 30 years: influences of technology innovation and legislation. Journal of Exposure Analysis and Environmental Epidemiology. 2003;13:325–340. [PubMed: 12973361]
29.
You XI, et al. Determinants of airborne concentrations of volatile organic compounds in rural areas of Western Canada. Journal of Exposure Science and Environmental Epidemiology. 2008;18:117–128. [PubMed: 17327851]
30.
Chan CY, et al. Volatile organic compounds in roadside microenvironments of metropolitan Hong Kong. Atmospheric Environment. 2002;36:2039–2047.
31.
Lash LH, et al. Metabolism of trichloroethylene. Environmental Health Perspectives. 2000;108(Suppl. 2):177–200. [PMC free article: PMC1637769] [PubMed: 10807551]
32.
Chiu WA, et al. Issues in the pharmacokinetics of trichloroethylene and its metabolites. Environmental Health Perspectives. 2006;114:1450–1456. [PMC free article: PMC1570093] [PubMed: 16966104]
33.
Guidelines for Canadian drinking water quality: supporting documentation. Trichloroethlyene. Ottawa: Ontario, Health Canada; 2005.
34.
Trichloroethylene. Geneva: International Programme on Chemical Safety; 1985. (Environmental Health Criteria No. 50)
35.
Fiches de données toxicologiques et environnementales des substances chimiques: trichloroéthylène. Verneuil-en-Halatte: Institut National de l'Environnement Industriel et des Risques; 2005.
36.
Hazardous Substances Data Bank (HSDB) [online database]. Bethesda, MD: National Library of Medicine; 2010. [19 May 2010]. http://toxnet​.nlm.nih​.gov/cgibin/sis/htmlgen?HSDB.
37.
Air quality guidelines for Europe. 2nd ed. Copenhagen: WHO Regional Office for Europe; 2000. Trichloroethylene. (WHO Regional Publications, European Series, No. 91)
38.
Caldwell JC, Keshava N. Key issues in the modes of action and effects of trichloroethylene metabolites for liver and kidney tumorigenesis. Environmental Health Perspectives. 2006;114:1457–1463. [PMC free article: PMC1570066] [PubMed: 16966105]
39.
Ramdhan DH, et al. Molecular mechanism of trichloroethylene-induced hepatotoxicity mediated by CYP2E1. Toxicology and Applied Pharmacology. 2008;231:300–307. [PubMed: 18565563]
40.
Kim S, et al. Pharmacokinetic analysis of trichloroethylene metabolism in male B6C3F1 mice: formation and disposition of trichloroacetic acid, dichloroacetic acid, S-(1,2-dichlorovinyl)glutathione and S-(1,2-dichlorovinyl)-L-cysteine. Toxicology and Applied Pharmacology. 2009;238:90–99. [PMC free article: PMC2737827] [PubMed: 19409406]
41.
Lipscomb JC, et al. In vitro to in vivo extrapolation for trichloroethylene metabolism in humans. Toxicology and Applied Pharmacology. 1998;152:376–387. [PubMed: 9853006]
42.
Indoor air quality guidelines for trichloroethylene. Maisons-Alfort: French Agency for Environmental and Occupational Health Safety; 2009.
43.
Goeptar AR, et al. Metabolism and kinetics of trichloroethlyene in relation to toxicity and carcinogenicity. Relevance of the mercapturic pathway. Chemical Research in Toxicology. 1995;8:3–21. [PubMed: 7703363]
44.
Fiche toxicologique: trichloroéthylène. Paris: Institut National de la Recherche et de la Sécurité; 2008.
45.
Merdink JL, et al. Kinetics of chloral hydrate and its metabolites in male human volunteers. Toxicology. 2008;245:130–140. [PubMed: 18243465]
46.
TLVs and BEIs based on the documentation of the threshold limit values for chemical substances and physical agents and biological exposure indices. Cincinnatti, OH: ACGIH; 2006.
47.
Waksman JC, Phillips SD. Biologic markers of exposure to chlorinated solvents. Clinics in Occupational and Environmental Medicine. 2004;4:413–421. [PubMed: 15325313]
48.
Sohn MD, McKone TE, Blancato JN. Reconstructing population exposures from dose biomarkers: inhalation of trichloroethylene (TCE) as a case study. Journal of Exposure Analysis and Environmental Epidemiology. 2004;14:204–213. [PubMed: 15141149]
49.
Liao KH, Tan YM, Clewell HJ 3rd. Development of a screening approach to interpret human biomonitoring data on volatile organic compounds: reverse dosimetry on biomonitoring data for trichloroethylene. Risk Analysis. 2007;27:1223–1236. [PubMed: 18076492]
50.
Allen BC, Fisher JW. Pharmacokinetic modeling of trichloroethylene and trichloroacetic acid in humans. Risk Analysis. 1993;13:71–86. [PubMed: 8451462]
51.
Abbas R, Fisher JW. A physiologically based pharmacokinetic model for trichloroethylene and its metabolites, chloral hydrate, trichloroacetate, dichloroacetate, trichloroethanol, and trichloroethanol glucuronide in B6C3F1 mice. Toxicology and Applied Pharmacology. 1997;147:15–30. [PubMed: 9356303]
52.
Fisher JW, Mahle D, Abbas R. A human physiologically based pharmacokinetic model for trichloroethylene and its metabolites, trichloroacetic acid and free trichloroethanol. Toxicology and Applied Pharmacology. 1998;152:339–359. [PubMed: 9853003]
53.
Fisher JW. Physiologically based pharmacokinetic models for trichloroethylene and its oxidative metabolites. Environmental Health Perspectives. 2000;108(Suppl. 2):265–273. [PMC free article: PMC1637763] [PubMed: 10807557]
54.
Greenberg MS, Burton GA, Fisher JW. Physiologically based pharmacokinetic modeling of inhaled trichloroethylene and its oxidative metabolites in B6C3F1 mice. Toxicology and Applied Pharmacology. 1999;154:264–278. [PubMed: 9931286]
55.
Clewell HJ 3rd, et al. Development of a physiologically based pharmacokinetic model of trichloroethylene and its metabolites for use in risk assessment. Environmental Health Perspectives. 2000;108(Suppl. 2):283–305. [PMC free article: PMC1637761] [PubMed: 10807559]
56.
Bois FY. Statistical analysis of Fisher et al. PBPK model of trichloroethylene kinetics. Environmental Health Perspectives. 2000;108(Suppl. 2):275–282. [PMC free article: PMC1637766] [PubMed: 10807558]
57.
Bois FY. Statistical analysis of Clewell et al. PBPK model of trichloroethylene kinetics. Environmental Health Perspectives. 2000;108(Suppl. 2):307–316. [PMC free article: PMC1637757] [PubMed: 10807560]
58.
Chiu WA, Okino MS, Evans MV. Characterizing uncertainty and population variability in the toxicokinetics of trichloroethylene and metabolites in mice, rats, and humans using an updated database, physiologically based pharmacokinetic (PBPK model, and Bayesian approach. Toxicology and Applied Pharmacology. 2009;241:36–60. [PubMed: 19660485]
59.
Report of the peer consultation of harmonized PBPK model for trichloroethylene. Cincinnati, OH: Toxicology Excellence for Risk Assessment; 2004.
60.
Yoon M, Madden MC, Barton HA. Extrahepatic metabolism by CYP2E1 in PBPK modeling of lipophilic volatile organic chemicals: impacts on metabolic parameter estimation and prediction of dose metrics. Journal of Toxicology and Environmental Health, Part A. 2007;70:1527–1541. [PubMed: 17710613]
61.
Simmons JE, et al. A physiologically based pharmacokinetic model for trichloroethylene in the male long-evans rat. Toxicological Sciences. 2002;69:3–15. [PubMed: 12215655]
62.
Bruckner JV, Keys DA, Fisher JW. The Acute Exposure Guideline Level (AEGL) program: applications of physiologically based pharmacokinetic modeling. Journal of Toxicology and Environmental Health, Part A. 2004;67:621–634. [PubMed: 15192858]
63.
Boyes WK, et al. Duration adjustment of acute exposure guideline level values for trichloroethylene using a physiologically-based pharmacokinetic model. Risk Analysis. 2005;25:677–686. [PubMed: 16022699]
64.
Simmons JE, Evans MV, Boyes WK. Moving from external exposure concentration to internal dose: duration extrapolation based on physiologically based pharmacokinetic derived estimates of internal dose. Journal of Toxicology and Environmental Health, Part A. 2005;68:927–950. [PubMed: 16020185]
65.
Hacks CE, et al. Bayesian population analysis of a harmonized physiologically bases pharmacokinetic model of trichloroethylene and its metabolites. Regulatory Toxicology and Pharmacology. 2006;46:63–83. [PubMed: 16889879]
66.
Evans MV, et al. Development of an updated PBPK model for trichloroethylene and metabolites in mice, and its application to discern the role of oxidative metabolism in induced hepatomegaly. Toxicology and Applied Pharmacology. 2009;236:329–340. [PubMed: 19249323]
67.
Kulig BM. The effects of chronic trichloroethylene exposure on neurobehavioral functioning in the rat. Neurotoxicology and Teratology. 1987;9:171–178. [PubMed: 3657753]
68.
Rebert CS, et al. Sensory-evoked potentials in rats chronically exposed to trichloroethylene: predominant auditory dysfunction. Neurotoxicology and Teratology. 1991;13:83–90. [PubMed: 2046630]
69.
Arito H, Takahashi M, Ishikawa T. Effect of subchronic inhalation exposure to low-level trichloroethylene on heart rate and wakefulness-sleep in freely moving rats. Sangyo Igaku. 1994;36:1–8. [PubMed: 8126932]
70.
Blain L, Lachapelle P, Molotchnikoff S. Electroretinal responses are modified by chronic exposure to trichloroethylene. Neurotoxicology. 1994;15:627–631. [PubMed: 7854598]
71.
Vyskocil A, et al. Ototoxicity of trichloroethylene in concentrations relevant for the working environment. Human and Experimental Toxicology. 2008;27:195–200. [PubMed: 18650250]
72.
Gash DM, et al. Trichloroethylene: parkinsonism and complex 1 mitochondrial neurotoxicity. Annals of Neurology. 2008;63:184–192. [PubMed: 18157908]
73.
Sano Y, et al. Trichloroethylene liver toxicity in mouse and rat: microarray analysis reveals species differences in gene expression. Archives of Toxicology. 2009;83:835–849. [PubMed: 19448997]
74.
Prendergast JA, et al. Effects on experimental animals of long-term inhalation of trichloroethylene, carbon tetrachloride, 1,1,1-trichloroethane, dichlorodifluoromethane, and 1,1-dichloroethylene. Toxicology and Applied Pharmacology. 1967;10:270–289. [PubMed: 4962252]
75.
Maltoni C, et al. Long-term carcinogenicity bioassays on trichloroethylene administered by inhalation to Sprague-Dawley rats and Swiss mice and B3C6F1 mice. Annals of the New York Academy of Sciences. 1988;534:316–342. [PubMed: 3389663]
76.
Lash LH, et al. Modulation of hepatic and renal metabolism and toxicity of trichloroethylene and perchloroethylene by alterations in status of cytochrome P450 and glutathione. Toxicology. 2007;235:11–26. [PMC free article: PMC1976278] [PubMed: 17433522]
77.
Khan S, et al. Effect of trichloroethylene (TCE) toxicity on the enzyme carbohydrate metabolism, brush border membrane and oxidative stress in kidney and other rat tissues. Food and Chemical Toxicology. 2009;47:1562–1568. [PubMed: 19361549]
78.
Chen XY, et al. Immune responses to trichloroethylene and skin gene expression profiles in Sprague Dawley rats. Biomedical and Environmental Sciences. 2006;19:346–352. [PubMed: 17190186]
79.
Seo M, et al. Augmentation of antigen-stimulated allergic responses by a small amount of trichloroethylene ingestion from drinking water. Regulatory Toxicology and Pharmacology. 2008;52:140–146. [PubMed: 18721841]
80.
Cai P, et al. Chronic exposure to trichloroethene causes early onset of SLE-like disease in female MRL +/+ mice. Toxicology and Applied Pharmacology. 2008;228:68–75. [PMC free article: PMC2442272] [PubMed: 18234256]
81.
Cai P, et al. Immuno- and hepato-toxicity of dichloroacetic acid in MRL +/+ and B6C3F1 mice. Journal of Immunotoxicology. 2007;4:107–115. [PubMed: 18958719]
82.
Cai P, et al. Differential immune responses to albumin adducts of reactive intermediates of trichloroethene in MRL+/+ mice. Toxicology and Applied Pharmacology. 2007;220:278–283. [PMC free article: PMC1959509] [PubMed: 17376499]
83.
Tang X, et al. Characterization of liver injury associated with hypersensitive skin reactions induced by trichloroethylene in the guinea pig maximization test. Journal of Occupational Health. 2008;50:114–121. [PubMed: 18403861]
84.
Gilbert KM, et al. Delineating liver events in trichloroethylene-induced autoimmune hepatitis. Chemical Research in Toxicology. 2009;22:626–632. [PubMed: 19254012]
85.
Wang G, et al. Increased nitration and carbonylation of proteins in MRL+/+ mice exposed to trichloroethene: potential role of protein oxidation in autoimmunity. Toxicology and Applied Pharmacology. 2009;237:188–195. [PMC free article: PMC2734328] [PubMed: 19332086]
86.
Shen T, et al. Trichloroethylene induced cutaneous irritation in BALB/c hairless mice: histopathological changes and oxidative damage. Toxicology. 2008;248:113–120. [PubMed: 18472203]
87.
National Research Council. Assessing the human health risks of trichloroethylene: key scientific issues. Washington, DC: National Academies Press; 2006.
88.
Kumar P, Prasad AK, Dutta KK. Steroidogenic alterations in testes and sera of rats exposed to trichloroethylene (TCE) by inhalation. Human and Experimental Toxicology. 2000;19:117–121. [PubMed: 10773841]
89.
Kumar P, et al. Trichloroethylene induced testicular toxicity in rats exposed by inhalation. Human and Experimental Toxicology. 2001;20:585–589. [PubMed: 11926613]
90.
Xu H, et al. Exposure to trichloroethylene and its metabolites causes impairment of sperm fertilizing ability in mice. Toxicological Sciences. 2004;82:590–597. [PubMed: 15375293]
91.
Kan FW, Forkert PG, Wade MG. Trichloroethylene exposure elicits damage in epididymal epithelium and spermatozoa in mice. Histology and Histopathology. 2007;22:977–988. [PubMed: 17523075]
92.
Berger T, Horner CM. In vivo exposure of female rats to toxicants may affect oocyte quality. Reproductive Toxicology. 2003;17:273–281. Erratum in Reproductive Toxicology, 2004, 18:447. [PubMed: 12759095]
93.
Wu KL, Berger T. Trichloroethylene metabolism in the rat ovary reduces oocyte fertilizability. Chemico-Biological Interactions. 2007;170:20–30. [PMC free article: PMC2085368] [PubMed: 17673192]
94.
Wu KL, Berger T. Reduction in rat oocyte fertilizability mediated by S-(1,2-dichlorovinyle)-L-cysteine: a trichloroethylene metabolite produced by the glutathione conjugation pathway. Bulletin of Environmental Contamination and Toxicology. 2008;81:490–493. [PubMed: 18679558]
95.
Lamb JC, Hentz KL. Toxicological review of male reproductive effects and trichloroethylene exposure: assessing the relevance to human male reproductive health. Reproductive Toxicology. 2006;22:557–563. [PubMed: 16938429]
96.
Williams AL, DeSesso JM. Trichloroethylene and ocular malformations: analysis of extant literature. International Journal of Toxicology. 2008;27:81–95. [PubMed: 18293215]
97.
Carney EW, et al. Developmental toxicity studies in Crl:CD (SD) rats following inhalation exposure to trichloroethylene and perchloroethylene. Birth Defects Research, Part B, Developmental and Reproductive Toxicology. 2006;77:405–412. [PubMed: 17066414]
98.
Blossom SJ, et al. Developmental exposure to trichloroethylene promotes CD4+ T cell differentiation and hyperactivity in association with oxidative stress and neurobehavioral deficits in MRL+/+ mice. Toxicology and Applied Pharmacology. 2008;231:344–353. [PubMed: 18579175]
99.
Kjellstrand P, et al. Trichloroethylene: further studies of the effects on body and organ weights and plasma butylcholinesterase activity in mice. Acta Pharmacologica et Toxicologica. 1983;53:375–384. [PubMed: 6659967]
100.
Fukuda K, Takemoto K, Tsuruta H. Inhalation carcinogenicity of trichloroethylene in mice and rats. Industrial Health. 1983;21:243–254. [PubMed: 6654707]
101.
Lash LH, Parker JC, Scott CS. Modes of action of trichloroethylene for kidney tumorigenesis. Environmental Health Perspectives. 2000;108(Suppl. 2):225–240. [PMC free article: PMC1637767] [PubMed: 10807554]
102.
Trichloroethylene health risk assessment: synthesis and characterization (draft). Washington, DC: US Environmental Protection Agency; 2001. (EPA/600/P-01/002A)
103.
Clewell HJ, Andersen ME. Applying mode-of-action and pharmacokinetic considerations in contemporary cancer risk assessments: an example with trichloroethylene. Critical Reviews in Toxicology. 2004;34:385–445. [PubMed: 15560567]
104.
Bull RJ. Mode of action of liver tumor induction by trichloroethylene and its metabolites. Environmental Health Perspectives. 2000;108(Suppl. 2):241–259. [PMC free article: PMC1637759] [PubMed: 10807555]
105.
Bull RJ, et al. Contribution of dichloroacetate and trichloroacetate to liver tumor induction in mice by trichloroethylene. Toxicology and Applied Pharmacology. 2002;182:55–65. [PubMed: 12127263]
106.
Report on carcinogens. 11th ed. Washington, DC: National Toxicology Program; 2005.
107.
Trichloroethylene in drinking-water. Background document for development of WHO Guidelines for Drinking-water Quality. Geneva: World Health Organization; 2005. (document WHO/SDE/WSH/05.08/22)
108.
Charbotel B, et al. Trichloroethylene exposure and somatic mutations of the VHL gene in patients with Renal Cell Carcinoma. Journal of Occupational Medicine and Toxicology. 2007;12:13. [PMC free article: PMC2211482] [PubMed: 17997830]
109.
Henschler D, et al. Carcinogenicity study of trichloroethylene by long-term inhalation in three animal species. Archives of Toxicology. 1980;43:237–248. [PubMed: 7387385]
110.
Trichloroethylene: assessment of human carcinogenic hazard. Brussels: European Centre for Ecotoxicology and Toxicology of Chemicals; 1994. (Technical Report No. 60)
111.
Hu C, et al. Possible involvement of oxidative stress in trichloroethylene-induced genotoxicity in human HepG2 cells. Mutation Research. 2008;652:88–94. [PubMed: 18289923]
112.
Clay P. Assessment of the genotoxicity of trichloroethylene and its metabolite, S-(1,2-dichlorovinyl)-L-cysteine (DCVC), in the comet assay in rat kidney. Mutagenesis. 2008;23:27–33. [PubMed: 18003627]
113.
Carrieri M, et al. Acute, nonfatal intoxication with trichloroethylene. Archives in Toxicology. 2007;81:529–532. [PubMed: 17285313]
114.
Takaki A, et al. A 27-year-old man who died of acute liver failure probably due to trichloroethylene abuse. Journal of Gastroenterology. 2008;43:239–242. [PubMed: 18373167]
115.
Nagaya T, et al. Subclinical and reversible hepatic effects of occupational exposure to trichloroethylene. International Archives of Occupational and Environmental Health. 1993;64:561–563. [PubMed: 8314614]
116.
Kamijima M, et al. Trichloroethylene causes generalized hypersensitivity skin disorders complicated by hepatitis. Journal of Occupational Health. 2008;50:328–338. [PubMed: 18540116]
117.
Xu X, et al. Severe hypersensitivity dermatitis and liver dysfunction induced by occupational exposure to trichloroethylene. Industrial Health. 2009;47:107–112. [PubMed: 19367038]
118.
Liu J, et al. Identification of antigenic proteins associated with trichloroethylene-induced autoimmune disease by serological proteome analysis. Toxicology and Applied Pharmacology. 2009;240:393–400. [PubMed: 19647757]
119.
Chia SE, et al. Semen parameters in workers exposed to tcrichloroethylene. Reproductive Toxicology. 1996;10:295–299. [PubMed: 8829252]
120.
Chia SE, Goh VH, Ong CN. Endocrine profiles of male workers with exposure to trichloroethylene. American Journal of Industrial Medicine. 1997;32:217–222. [PubMed: 9219650]
121.
Goh VH, Chia SE, Ong CN. Effects of chronic exposure to low doses of trichloroethylene on steroid hormone and insulin levels in normal men. Environmental Health Perspectives. 1998;106:41–44. [PMC free article: PMC1532938] [PubMed: 9417767]
122.
Wilson PD, et al. Attributable fraction for cardiac malformations. American Journal of Epidemiology. 1998;148:414–423. [PubMed: 9737553]
123.
Ferencz C, et al. Genetic and environmental risk factors of major cardiovascular malformations. The Baltimore–Washington Infant Study 1981 – 1989. New York, NY: Blackwell/Futura; 1997.
124.
Lagakos SW, Wessen BJ, Zelen M. An analysis of contaminated well water and health effects in Woburn, Massachusetts. Journal of the American Statistical Association. 1986;81:583–596.
125.
Swan SH, et al. Congenital cardiac anomalies in relation to water contamination, Santa Clara County, California, 1981–1983. American Journal of Epidemiology. 1989;129:885–893. [PubMed: 2784935]
126.
Bove FJ, et al. Public drinking water contamination and birth outcomes. American Journal of Epidemiology. 1995;141:850–862. [PubMed: 7717362]
127.
Rodenbeck SE, Sanderson LM, Rene A. Maternal exposure to trichloroethylene in drinking water and birth-weight outcomes. Archives of Environmental Health. 2000;55:188–194. [PubMed: 10908102]
128.
Health statistics review: cancer and birth outcome analysis, Endicott Area, Town of Union, Broome County, New York. New York, NY: Center for Environmental Health; 2005.
129.
Yauck JS, et al. Proximity of residence to trichloroethylene-emitting sites and increased risk of offspring congenital heart defects among older women. Birth Defects Research and Clinical Molecular Teratology. 2004;70:808–814. [PubMed: 15390315]
130.
Anttila A, et al. Cancer incidence among Finnish workers exposed to halogenated hydrocarbons. Journal of Environmental and Occupational Medicine. 1995;37:797–806. [PubMed: 7552463]
131.
Lindbohm ML, et al. Risks of liver cancer and exposure to organic solvents and gasoline vapors among Finish workers. International Journal of Cancer. 2009;124:2954–2959. [PubMed: 19319983]
132.
Axelson O, et al. Updated and expanded Swedish cohort study on trichloroethylene and cancer risk. Journal of Occupational Medicine. 1994;36:556–562. [PubMed: 8027881]
133.
Spirtas R, et al. Retrospective cohort mortality study of workers at an aircraft maintenance facility. I. Epidemiological results. British Journal of Industrial Medicine. 1991;48:515–530. [PMC free article: PMC1035412] [PubMed: 1878308]
134.
Seidler A, et al. Solvent exposure and malignant lymphoma: a population-based case-control study in Germany. Journal of Occupational Medicine and Toxicology. 2007;2:2–12. [PMC free article: PMC1851965] [PubMed: 17407545]
135.
Scott CS, Chiu WA. Trichloroethylene cancer epidemiology: a consideration of select issues. Environmental Health Perspectives. 2006;114:1471–1478. [PMC free article: PMC1570052] [PubMed: 16966107]
136.
Mandel JH, et al. Occupational trichloroethylene exposure and non-Hodgkin's lymphoma: a meta-analysis and review. Occupational and Environmental Medicine. 2006;63:597–607. [PMC free article: PMC2078160] [PubMed: 16644896]
137.
Alexander DD, et al. A meta-analysis of occupational trichloroethylene exposure and multiple myeloma or leukaemia. Occupational Medicine (London). 2006;56:485–493. [PubMed: 16905622]
138.
Alexander DD, et al. A meta-analysis of occupational trichloroethylene exposure and liver cancer. International Archives of Occupational and Environmental Health. 2007;81:127–143. [PubMed: 17492303]
139.
Wartenberg D, Reyner D, Siegel C. Trichloroethylene and cancer: epidemiologic evidence. Environmental Health Perspectives. 2000;10(Suppl. 2):161–176. [PMC free article: PMC1637753] [PubMed: 10807550]
140.
Lewandowski TA, Rhomberg LR. A proposed methodology for selecting a trichloroethylene inhalation unit risk value for use in risk assessment. Regulatory Toxicology and Pharmacology. 2005;41:39–54. [PubMed: 15649826]
141.
Pastino G, Yap W, Carroquino M. Human variability and susceptibility to trichloroethylene. Environmental Health Perspectives. 2000;108(Suppl. 2):201–214. [PMC free article: PMC1637770] [PubMed: 10807552]
142.
Brüning T, et al. Renal cell cancer risk and occupational exposure to trichloroethylene: results of a consecutive case-control study in Arnsberg, Germany. American Journal of Industrial Medicine. 2003;43:274–285. [PubMed: 12594774]
143.
Wiesenhütter B, et al. Re-assessment of the influence of polymorphisms of phase-II metabolic enzymes on renal cell cancer risk of trichloroethylene-exposed workers. International Archives of Occupational and Environmental Health. 2007;81:247–251. [PubMed: 17479278]
144.
Ruden C. Interpretations of primary carcinogenicity data in 29 trichloroethylene risk assessments. Toxicology. 2001;169:209–225. [PubMed: 11718961]
145.
Ruden C. The use and evaluation of primary data in 29 trichloroethylene carcinogen risk assessments. Regulatory Toxicology and Pharmacology. 2001;34:3–16. [PubMed: 11502152]
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