Page 9: Guidelines for Canadian Drinking Water Quality: Guideline Technical Document – Trichloroethylene

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8.0 Health Effects

8.1 Effects in Humans

Central nervous system effects were the primary effects noted from acute inhalation exposure to TCE in humans, with symptoms including sleepiness, fatigue, headache, confusion and feelings of euphoria (ATSDR, 1997). Simultaneous exposure to TCE and ethanol results in a marked inhibition of the metabolism of TCE, which leads to an accumulation of TCE in blood and increases the extent of central nervous system depression (Muller et al., 1975). Effects on the liver, kidneys, gastrointestinal system and skin have also been noted (ATSDR, 1997). In its wide use as an inhalant anaesthetic drug in humans, concentrated solutions of TCE have proved quite irritating to the gastrointestinal tract and have caused nausea and vomiting (DeFalque, 1961).

Information from medium- to long-term TCE exposures via inhalation and dermal routes has been reviewed (ATSDR, 1997). These studies indicated that the central nervous system is the most sensitive organ for toxicity, with the liver and kidneys the next most sensitive sites for the chronic toxicity of TCE exposure. Case reports of intermediate and chronic occupational exposures included effects such as dizziness, headache, sleepiness, nausea, confusion, blurred vision, facial numbness and weakness. The liver effects noted included liver enlargement and increases in serum levels of liver enzymes, and the kidney effects included increased N-acetyl-β-D-glucosaminidase. Cardiovascular, immunological, reproductive and carcinogenic effects were also observed (ATSDR, 1997).

The demonstration of TCE-induced genetic toxicity in humans has been largely inconclusive. Four studies of sister chromatid exchange (SCE) in peripheral lymphocyte cultures from exposed workers showed no or only minor effects on SCE frequencies (Gu et al., 1981a,b; Nagaya et al., 1989; Brandom et al., 1990; Seiji et al., 1990). Although the studies by Gu et al. (1981a,b) suggested that TCE or a metabolite may have caused chromosomal aberrations or SCE in chronically exposed humans, exposure to additional compounds, including TCE contaminants, cannot be ruled out. Konietzko et al. (1978) found a higher incidence of hypodiploid cells and a greater frequency of chromosome breaks in exposed workers compared with an unmatched control group; the authors did not consider this increase to be biologically significant, and no statistical evaluation of the data was provided. Rasmussen et al. (1988) found a highly significant increase in the frequency of structural aberrations and hyperdiploid cells in cultured lymphocytes from TCE degreasers. However, even though the control group used in that study consisted of physicians and was therefore not equivalent to the exposed group, the study did not account for the different lifestyles of the two groups and confounding factors such as smoking, as well as possible simultaneous exposure to a number of other substances, possibly including genotoxic polycyclic aromatic hydrocarbons.

Most epidemiological studies have found no association between adverse reproductive effects in humans and exposure to TCE in contaminated drinking water (IPCS, 1985; ATSDR, 1997). Although an epidemiological study of 2000 male and female workers exposed to TCE via inhalation found no increase in malformations in babies born following exposure (IPCS, 1985), an association was found between the occurrence of congenital heart disease in children and a drinking water supply contaminated with TCE and other similar chemicals (IPCS, 1985). These earlier studies were confounded by, among other factors, potential exposure to many other contaminants or compounds that produce similar metabolites, the lack of characterization of the exposure levels and the exposed populations, and failure to characterize the nature of the "congenital heart disease," which may not necessarily be equivalent to cardiac anomalies. Therefore, their use in inferring a causal association between TCE and congenital cardiac anomalies remains very limited. More recent epidemiological studies of women exposed to degreasing solvents, including TCE, have reported elevated risks for cardiac anomalies in their offspring (Goldberg et al., 1990; Ferencz et al., 1997; Wilson et al., 1998). Large, statistically significant excesses were observed for specific cardiac defects: left-sided obstructive defects (odds ratio [OR] = 6.0, 95% confidence interval [CI] = 1.7-21.3) and hypoplastic left heart (OR = 3.4, 95% CI = 1.6-6.9), with an attributable risk^{Footnote 2} of 4.6% (Wilson et al., 1998). Neural tube defects have also been noted with either occupational or drinking water exposure to solvents, including TCE (Holmberg and Nurminen, 1980; Holmberg et al., 1982; Bove et al., 1995). Overall, these epidemiological studies are plagued by lack of clarity on the background co-exposure. For example, in the Wilson et al. (1998) study, the investigators asked subjects about their exposure to "solvents/de-greasing compounds," but not specifically about their exposure to TCE. Although it is generally acknowledged that subjects at air force bases are exposed to jet fuels as well as other solvents on a daily basis (Stewart et al., 1991), it is unlikely that the individuals know the exact compounds contained in the degreasing compounds or solvents. This suggests that, based on currently available human studies, TCE cannot be specifically implicated; however, these studies can be used as supporting evidence, complementary to developmental-reproductive effects reported in animal studies. In a study in which semen parameters of workers exposed to TCE were evaluated (Chia et al., 1996), sperm density showed a significant difference between low- and high-exposure subjects. In a recent study involving a small number of subjects, TCE and its metabolites were identified in seminal fluids of workers exposed to TCE (Forkert et al., 2003), suggesting that TCE may play a role in the observed effects on sperm parameters.

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

The evidence for TCE-induced cancers in humans has been reviewed in depth by IARC (1995). Three cohort studies were considered to be relevant to TCE evaluation. Two of these studies, in Sweden and Finland (Axelson et al., 1994; Anttila et al., 1995), involved people who had been monitored for exposure to TCE by measurement of TCA in urine. The third study, in the United States (Spirtas et al., 1991), covered workers exposed to TCE during maintenance of military aircraft and missiles, some of whom were also exposed to other solvents. In none of the available cohort studies was it possible to control for potential confounding factors, such as smoking (IARC, 1995). Most importantly, an elevated risk for liver and biliary tract cancer was observed, in addition to a modestly elevated risk for non-Hodgkin's lymphoma seen in cohort studies. A marginally increased risk for non-Hodgkin's lymphoma was suggested to exist in areas where groundwater is contaminated with TCE (IARC, 1995). The occurrence of renal cancer was not elevated in the cohort studies, although a study of German workers exposed to TCE yielded five cases of renal cancer compared with none in a control comparison group (IARC, 1995).

After meta-analysis of the four occupational studies (Garabrant et al., 1988; Spirtas et al., 1991; Axelson et al., 1994; Anttila et al., 1995), the following standardized mortality ratios (SMRs) resulted: liver cancer, 1.32; prostate cancer, 1.09; kidney cancer, 1.09; bladder cancer, 1.15; and non-Hodgkin's lymphoma, 1.25. However, the small number of cases (except for prostate cancer), even though they were aggregated across four studies, limits the interpretation of these findings. Other limitations include narrowly defined exposure groups, lack of data on potential confounders, such as smoking, diet and exposure to other solvents, and no direct measure of personal exposure.

The authors of a retrospective cohort study conducted on 169 workers in a cardboard factory in Germany who were exposed to TCE for at least 1 year between 1956 and 1975 claim a causal link between cancer and TCE exposure (Henschler et al., 1995a,b). By the close of the study in 1992, 50 members of the study group had died, 16 from malignant neoplasms. In 2/16 cases, kidney cancer was the cause of death (SMR = 3.28, vs. local population). Five workers were diagnosed with kidney cancer: four with renal cell cancer and one with a urothelial cancer of the renal pelvis (standardized incidence ratio [SIR] = 7.77, 95% CI = 2.50-18.59). After the close of the observation period, two additional kidney tumours (one renal and one urothelial) were diagnosed in the study group. By the end of the study, 52 members of the control group, which consisted of 190 unexposed workers from the same plant, had died -- 16 from malignant neoplasms, but none from kidney cancer. No case of kidney cancer was diagnosed in the control group. For the seven cases of kidney cancer, the average exposure duration was 15.2 years (range 3-19.4 years).

The GST gene family encodes multi-functional enzymes that catalyse several reactions between GST and elecrophilic as well as hydrophobic compounds (Raunio et al., 1995). Certain defective GST genes are known to be associated with an increased risk of different kinds of cancer. A recent case-control study (Bruning et al., 1997b) investigated the role of GST polymorphisms on the incidence of renal cell cancer in two occupational groups exposed to high levels of TCE. The data indicate a higher risk for development of renal cell cancer if TCE-exposed persons carry either the GSTT1 or GSTM1 gene. The authors concluded that this genetic polymorphism may indicate predisposition for TCE-induced renal cell cancer. These results tend to support the view of the mode of action of TCE-induced kidney cancer as involving metabolites derived from the GSH-dependent pathway, at least in humans, and are supported by the study of Henschler et al. (1995a), which reaffirms the relevance of increased incidences of renal cell tumours in a cohort of cardboard workers exposed to TCE.

The epidemiological studies of TCE and PCE as they relate to risk of renal cell cancer were critically reviewed by McLaughlin and Blot (1997). The authors state that there was little evidence of an increased risk of renal cell cancer with exposure to TCE or PCE. The few studies with elevations in risk suffered from important methodological shortcomings. Although it was virtually impossible, using epidemiological data, to conclusively rule out a small increase in risk of renal cell cancer, the totality of the epidemiological evidence clearly did not support a causal association with TCE or PCE (McLaughlin and Blot, 1997). Although McLaughlin and Blot (1997) criticized the Henschler et al. (1995a) study, it is impossible to ignore the findings of Henschler et al. (1995a), particularly in light of the authors' response to the published critique (Henschler et al., 1995b).

Over 80 published papers and letters on the cancer epidemiology of people exposed to TCE were reviewed byWartenberg et al.. (2000). Evidence of excess cancer incidence among occupational cohorts with the most rigorous exposure assessment is found for kidney cancer (relative risk [RR] = 1.7, 95% CI = 1.1-2.7), liver cancer (RR = 1.9, 95% CI = 1.0-3.4) and non-Hodgkin's lymphoma (RR = 1.5, 95% CI = 0.9-2.3), as well as for cervical cancer, Hodgkin's disease and multiple myeloma. However, since few studies isolate TCE exposure, results are likely confounded by exposure to other solvents and risk factors. More recently, a positive association between renal cancer and prolonged occupational exposure to high levels of TCE has been reaffirmed (Bruning et al., 2003) in a case-control study in Germany involving 134 renal cell cancer patients and 410 controls, comprising workers from industries with and without TCE exposure. When the results were adjusted for age, gender and smoking, a significant excess risk was determined for the longest-held job in industries with TCE exposure (OR = 1.80, 95% CI = 1.01-13.32). Any exposure to degreasing agents was found to be a risk factor for renal cell cancer (OR = 5.57, 95% CI = 2.33-13.32), while self-reported narcotic symptoms, an indication of peak exposures, were associated with an excess risk for renal cell cancer (OR = 3.71, 95% CI = 1.80-7.54). However, the levels of occupational exposure in that study were very high and unlikely to be reached from environmental exposure. The prolonged exposure to high levels likely affect the metabolism of TCE, with the net production of active metabolites underlying the development of renal cell cancer in occupationally exposed industrial workers.

A recent novel feature of the cancer database for TCE has been the molecular information on the von Hippel Landau (VHL) tumour suppressor gene. Mutations in the VHL tumour suppressor gene have been associated with increased risk of renal cell carcinoma. Recent studies provide evidence that TCE exposure may be associated with VHL mutations among renal cell carcinoma patients (Bruning et al., 1997a; Brauch et al., 1999). Bruning et al. (1997a) examined VHL mutation by single-stranded conformation polymorphism (SSCP) in 23 renal cell carcinoma patients with documented high occupational TCE exposure. All (100%) TCE-exposed renal cell carcinoma patients had VHL mutations, which was higher than the background frequency (33-55%) among unexposed renal cell carcinoma patients. Brauch et al. (1999), in a follow-up study that determined VHL mutations by SSCP and direct sequencing of mutations in renal tissue from 44 TCE-exposed renal cell carcinoma patients, found that 75% of TCE-exposed patients had VHL mutations and 39% had a C to T mutation at nucleotide 454. All the C to T transitions in the control renal cell carcinoma patients were relatively rare (6% of the total incidence). In the Brauch et al. (1999) study, the VHL mutations were detected in patients with medium and high, but not low, TCE exposure, although only three patients were classified as having low exposure. These data indicate a highly significant association (p = 0.0006) between TCE exposure and multiplicity of VHL mutations.

Overall, although several studies have indicated a positive association between solvent exposure and human cancer, further study is still necessary to better specify the specific agents that confer this risk and to estimate the magnitude of that risk (Wartenberg et al., 2000).

8.2 Effects on Experimental Animals and In Vitro

Many studies of a wide range of toxic endpoints using repeated oral exposures to TCE have been reviewed (NTP, 1985, 1986, 1990; Barton et al., 1996; Kaneko et al., 1997). Due to the poor solubility of TCE in water, few studies used water as a vehicle (Tucker et al., 1982), although some drinking water or water gavage studies have used emulsifying agents. Many of the studies are therefore confounded by the use of corn oil as a vehicle, which has been found to alter the pharmacokinetics of TCE and to affect lipid metabolism and other pharmacodynamic processes. The best-documented systemic effects are neurotoxicity, hepatotoxicity, nephrotoxicity and pulmonary toxicity in adult animals. Reproductive and developmental effects have also been extensively studied.

8.2.1 Acute Toxicity

Neurological, lung, kidney and heart effects have been reported in animals acutely exposed to TCE (ATSDR, 1993, 1997). Tests involving acute exposure of rats and mice have shown TCE to have low toxicity from inhalation exposure and moderate toxicity from oral exposure (RTECS, 1993; ATSDR, 1997). The 14-day acute oral LD₅₀ values for TCE were determined to be 2400 mg/kg bw in mice (Tucker et al., 1982) and 4920 mg/kg bw in rats (Smyth et al., 1969; IPCS, 1985; ATSDR, 1993, 1997). The 4-hour inhalation LC₅₀ was calculated to be 12 500 ppm in rats (Siegel et al., 1971) and 8450 ppm in mice (Fan, 1988). A review of studies of dermal exposure of TCE in rabbits indicates that skin irritation occurs after 24 hours at 0.5 mL and degenerative skin changes occur within 15 minutes at 1 mL in guinea pigs (Fan, 1988). Instillation of 0.1 mL to rabbit eyes caused conjunctivitis and keratitis, with complete recovery within 2 weeks.

8.2.2 Short-term Exposure

In a 13-week oral study, Fischer 344/N rats and B6C3F1 mice (10 per sex per dose) were administered TCE in corn oil by gavage at doses of up to 1000 mg/kg bw per day in female rats and up to 2000 mg/kg bw per day in male rats, or up to 6000 mg/kg bw per day in mice of both sexes, for 5 days per week (NTP, 1990). Body weights were decreased in male rats at 2000 mg/kg bw per day. Pulmonary vasculitis involving small veins was reported in female rats at 1000 mg/kg bw per day. Mild to moderate cytomegaly and karyomegaly of the renal tubular epithelial cells occurred in rats at 1000 mg/kg bw per day (females) or 2000 mg/kg bw per day (males). The no-observed-adverse-effect level (NOAEL) in rats was reported as 1000 mg/kg bw per day (males) and 500 mg/kg bw per day (females). Among the mice, there were decreases in survival in both sexes and body weight gain in males at 750 mg/kg bw per day and above. Doses of 3000 mg/kg bw per day and above were associated with centrilobular necrosis and multifocal calcification in the liver, as well as mild to moderate cytomegaly and karyomegaly of the renal tubular epithelial cells in both sexes. A NOAEL was set at 375 mg/kg bw per day for mice.

Exposure to TCE through drinking water has been evaluated in subchronic studies (Sanders et al., 1982; Tucker et al., 1982). CD-1 and ICR outbred albino mice (140 per sex per dose) were administered TCE in a 1% solution of Emulphor in drinking water at dose levels of 0, 0.1, 1.0, 2.5 or 5.0 mg/L (equivalent to 0, 18.4, 216.7, 393 or 660 mg/kg bw per day) for 4 or 6 months. Females at 5.0 mg/L and males at and above 2.5 mg/L consumed less water than the controls. A decrease in body weight gain in both sexes and an increase (p < 0.05) in kidney weight in males occurred at 5.0 mg/L. In addition, at 5.0 mg/L, there were elevated urinary protein and ketone levels in both sexes, decreases in leukocyte and red blood cell counts in males, altered coagulation times in both sexes and shortened prothrombin times in females. At 2.5 mg/L, there was enlargement of the liver and an increase in urinary protein and ketone levels in males. Inhibition of humoral immunity, cell-mediated immunity and bone marrow stem cell colonization was seen among females at 2.5 mg/L and greater. The lowest-observed-adverse-effect level (LOAEL) was considered to be 2.5 mg/L, based on decreased water consumption, enlargement of the liver, increases in urinary protein and ketone levels in males (an indication of renal effects) and changes in immunological parameters in females. A NOAEL of 1.0 mg/L (equivalent to 216.7 mg/kg bw per day) was determined as a result of these studies. Several previous oral studies in animals had not documented evidence of renal toxicity in mice or rats exposed to TCE (Stott et al., 1982).

Several studies have evaluated the toxicity of TCE to rodents following short-term inhalation exposure. In a 14-week inhalation study, rats were exposed to 0, 49, 175 or 330 ppmv TCE for 4 hours per day, 5 days per week, for 14 weeks. Another group was exposed to 55 ppmv TCE for 8 hours per day, 5 days per week, for 14 weeks. There were significant increases (p < 0.01) in the absolute and relative liver weights in treated animals compared with controls, although liver and kidney function tests of treated animals remained within normal limits (Kimmerle and Eben, 1973). In a study in which mice, rats and gerbils (unspecified strains) were exposed to TCE continuously by inhalation at 150 ppmv for 30 days, there was a significant increase (p < 0.05) in the liver weights of all three species (Kjellstrand et al.., 1981). Renal effects of inhaled TCE have also been reported (Kjellstrand et al., 1981, 1983a,b). Male and female gerbils exposed to 150 ppmv atmospheres of TCE continuously for 30 days had increased (p < 0.05) kidney weight. NMRI mice exposed to 37, 75, 150 or 300 ppmv TCE continuously for 30 days had significantly increased (p < 0.05) kidney weight at 75 ppmv in males and above 150 ppmv in females. No kidney effects were evident in the remaining strains of mice (Kjellstrand et al., l983a).

8.2.3 Long-term Exposure and Carcinogenicity

Administration of high doses of TCE by gavage for long durations in rats and mice has been associated with nephropathy, with characteristic degenerative changes in the renal tubular epithelium (NCI, 1976), while toxic nephrosis, characterized by cytomegaly of the renal tubular epithelium, has been reported in cancer bioassays in mice and rats (NTP, 1983, 1988, 1990). The toxicity of TCE was investigated in F344 rats and B6C3F1 mice (50 per sex per dose) given 0, 500 or 1000 mg/kg bw per day (rats) and 0 or 1000 mg/kg bw per day (mice) in corn oil, 5 days per week for 103 weeks. Survival was reduced in male rats and mice but not in females (NTP, 1983). Toxic nephrosis, characterized as cytomegaly of renal tubular epithelium, occurred in rats at 500 mg/kg bw per day and above and in mice at 1000 mg/kg bw per day. LOAELs of 500 mg/kg bw per day in rats and 1000 mg/kg bw per day in mice were defined for long-term effects. A NOAEL was not determined (NTP, 1990).

Carcinogenicity studies of TCE by the oral route in rodents have demonstrated treatment-related liver tumours in mice in both sexes and kidney tumours in rats of both sexes (NCI, 1976; NTP, 1983, 1988, 1990). Oral exposure to TCE has also been shown to increase malignant lymphomas in female mice (U.S. EPA, 2001a). An increase in the incidence of testicular interstitial cell tumours was also reported in male rats. However, due to inadequacies of the study, a conclusive interpretation of the interstitial cell tumour incidence data could not be reached (NTP, 1988). Carcinogenicity studies of TCE by the inhalation route have shown treatment-related tumours in the lungs of female and male mice (Fukuda et al., 1983; Maltoni et al., 1986), testes of rats (Maltoni et al., 1986), the lymphoid system (lymphomas) in female mice (Henschler et al.., 1980), the kidney in male rats and the liver in mice of both sexes (Maltoni et al.., 1986). However, the early oral studies were confounded by the use of impure test material (TCE), which was stabilized with other compounds, such as epichlorohydrin, that are themselves known to be carcinogenic.

In a carcinogenicity assay exposing rodents to TCE by gavage (NTP, 1983), there was a significant increase in the incidences of hepatocellular carcinomas (p < 0.05) in male mice (13/49 relative to 8/48 in controls) and hepatocellular adenomas (p < 0.05) in female mice (8/49 compared with 2/48 in controls) (Table 1). There were no treatment-related liver tumours in rats.The male rats at 1000 mg/kg bw per day that survived until the end of the study exhibited a higher (p = 0.028) incidence of renal tubular cell adenocarcinomas (3/16 compared with 0/33 among controls; Table 1). These kidney tumours were considered biologically significant, given the rarity of kidney tumours in that rat strain.

In another carcinogenicity study (NTP, 1988) exposing four different rat strains (ACI, August, Marshall and Osborne-Mendel) to TCE by gavage, male Osborne-Mendel rats exhibited a statistically significant (p < 0.05) increase in the incidence of renal cell adenomas and adenocarcinomas (Table 2). The incidence of testicular interstitial cell tumours was also increased in the male Marshall rats (Table 2). However, closer audits of this study indicated that the documentation of many aspects of the study was inadequate to support proper interpretation of the reported tumour incidence data, although, given the rarity of kidney tumours in rats, this finding was still considered significant. No other treatment-related tumours were reported in these rat strains.

In a more recent carcinogenicity study (NTP, 1990) exposing B6C3F1 mice and F344/N rats to TCE by gavage, there was a significant (p < 0.05) increase in the incidences of combined hepatocellular carcinoma and adenomas (p < 0.05) in female mice (Table 3). No treatment-related kidney tumours were observed in mice. Although the study authors considered the results equivocal due to reduced survival in the treated groups, the kidney tumour incidences in rats were statistically significant (p < 0.05) when adjusted for reduced survival and were considered toxicologically significant due to the rarity of kidney tumours in the rats (Table 3).

In a long-term carcinogenicity study by the inhalation route (Maltoni et al., 1986), the increased incidence of renal tubular adenocarcinomas in male rats was statistically significant (p < 0.05) when adjusted for survival (U.S. EPA, 2001a) (Table 4). The authors noted a lack of statistical significance, but indicated that the findings were biologically significant due to the rarity of renal tubular adenocarcinomas in control animals and the rarity of kidney tumours in historical controls (0/460) (Maltoni et al., 1986).

In summary, animal carcinogenicity studies conducted using pure TCE showed that chronic exposure to this compound by the oral route resulted in malignant liver tumours in mice of both sexes and kidney tumours in male rats, while inhalation exposure led to lymphomas in female mice, malignant liver and lung tumours in mice of both sexes and malignant kidney tumours in male rats.

8.2.4 Mutagenicity/Genotoxicity

A range of assays, covering a wide spectrum of genetic endpoints, has been performed to assess possible genotoxic effects produced by TCE or its metabolites. DNA- or chromosome-damaging effects have been evaluated in bacteria, fungi, yeast, plants, insects, rodents and humans. The genetic endpoints measured by these assays include forward and reverse mutation, SCE, unscheduled DNA synthesis, gene conversion, chromosomal aberrations, micronuclei formation and mitotic recombination. Induction of DNA repair and covalent binding to DNA have also been examined.

The evidence for TCE genotoxicity is often conflicting, in part because of the presence of impurities or mutagenic stabilizers in the test material. In fact, the information from many of the early studies may not be adequate for complete evaluation of the genotoxic potential of TCE, as few of the studies identified the grade and purity of the test TCE. In addition, some TCE samples used contained a mutagenic stabilizer, and other assays used pure samples without stabilizers, which may have decomposed to chemicals with mutagenic activity, further confounding the interpretation of the significance of the findings.

Genotoxicity studies conducted until the mid-1990s often reported conflicting results, so the evidence for TCE or its metabolites being potent mutagens is quite limited. TCE is weakly active both in vitro and in vivo, inducing recombination responses, including SCE, and aneuploidies, including micronuclei; however, it appears to be unable to induce gene mutations or structural chromosomal aberrations (Crebelli and Carere, 1989; Fahrig et al., 1995). TCE was also observed to induce increased DNA synthesis and mitosis in mouse liver in vivo (Dees and Travis, 1993). Despite the apparent lack of "typical" genetic toxicity, TCE could be involved in the expression of carcinogen-induced mutations due to its potential to induce recombination and aneuploidy (Fahrig et al., 1995). In general, TCE , TCA and DCA have all been shown to cause DNA strand breaks in rodent liver cells in vivo and in culture, at high concentrations, as either the parent molecule or its metabolites (Bull, 2000). However results of some studies appear to contradict these findings (Styles et al., 1991; Chang et al., 1992), and it is still unclear whether DNA strand breaks are produced by TCE itself or by its metabolites.

Many genotoxicity studies have been conducted for the major metabolites of TCE. In a recent review, Moore and Harrington-Brock (2000) concluded that TCE and its metabolites CH, DCA and TCA require very high doses to be genotoxic, but that there was not enough information to draw any conclusions for TCOH and the conjugates DCVC and DCVG. Definitive conclusions as to whether TCE will induce tumours in humans via a mutagenic mode of action cannot, therefore, be drawn from the available information.

In summary, while the genotoxicity data are not fully conclusive, there appears to be evidence to show that TCE has a weak, likely indirect, genotoxic effect at high doses. Therefore, the mutagenic potential for this compound cannot be disregarded.

8.2.5 Reproductive and Developmental Toxicity

The reproductive, embryo-fetotoxic and teratogenic effects of TCE have been studied in several species (Smith et al., 1989; Dawson et al., 1990, 1993; Johnson et al., 1998a,b). In an inhalation reproductive toxicity study, Long-Evans rats were exposed to TCE at 1800 ppmv for 6 hours per day, 5 days per week, for 12 weeks before mating; for 6 hours per day, 7 days per week, during pregnancy through gestation day 21; or for 6 hours per day, 5 days per week, for 2 weeks before mating and for 6 hours per day, 7 days per week, during pregnancy through gestation day 21. Incomplete ossification of the sternum, indicative of delay in maturation, occurred in animals exposed during pregnancy, while a significant decrease in postnatal weight gain occurred in offspring of the premating exposed group. No maternal toxicity, teratogenicity or other effects on reproductive parameters were observed (Dorfmueller et al., 1979).

In a two-generation reproductive toxicity study, male and female Fischer 344 rats were fed diets containing microencapsulated TCE at doses of approximately 0, 75, 150 or 300 mg/kg bw per day from 7 days before mating right through to the birth of the F₂ generation. Although left testicular and epididymal weights decreased in the F₀ and F₁ generations, no associated histopathological changes were observed. The weight changes were attributed to general toxicity, rather than reproductive toxicity (NTP, 1986). In a similar two-generation reproductive toxicity study in CD-1 mice given TCE up to 750 mg/kg bw daily, sperm motility was reduced by 45% in F₀ males and 18% in F₁ males, but there were no treatment-related effects on mating, fertility or reproductive performance in the F₀ or F₁ animals (NTP, 1985).

A number of teratogenicity studies have been conducted using TCE by both oral and inhalation routes. Swiss Webster mice exposed by inhalation to 300 ppm TCE for 7 hours per day on gestation days 6-15 did not have any observable treatment-related maternal toxicity or terata (Leong et al., 1975). When Swiss Webster mice and Sprague-Dawley rats were exposed to TCE by inhalation at a concentration of 1600 mg/m³ (300 ppmv), 7 hours per day on gestation days 6-15, a significant decrease (p < 0.05) in maternal weight gain and some evidence of haemorrhages in the cerebral ventricles were observed, but no teratogenic or reproductive effects were seen (Schwetz et al., 1975). In contrast, a significant decrease in fetal weight and some increase in fetal resorptions were reported in rats (strain not specified) exposed to 100 ppmv TCE for 4 hours per day during gestation days 8-21 (Healy et al., 1982).

In a study of the effect of exposure to TCE on developmental/reproductive function, female Sprague-Dawley rats were exposed to TCE in drinking water at 0, 1.5 or 1100 ppm (equal to 0, 0.18 or 132 mg/kg bw per day) in one of three dose regimens: for 3 months before pregnancy; for 2 months before and 21 days during pregnancy; or for 21 days during pregnancy only (Dawson et al., 1993). No maternal toxicity was observed at any dose level or regimen. An increase in incidence of fetal heart defects (3% controls, 8.2% and 9.2%) was observed in treated animals at both dose levels (0.18 or 132 mg/kg bw per day) in dams exposed before and during pregnancy and only at the high (132 mg/kg bw per day) dose (10.4% vs. 3% in controls) in animals exposed only during pregnancy. The LOAEL was set at 0.18 mg/kg bw per day, based on the increased incidence of heart defects in fetuses born to dams that were exposed prior to and during gestation. However, the study was limited in that it expressed the incidence of malformation only as a proportion of the total number of fetuses in the dose group and did not attempt to establish the incidence of heart defects on a per litter basis. Notwithstanding that shortcoming, the study lends support to similar findings of increased congenital defects in epidemiological studies (Goldberg et al., 1990; Bove et al., 1995), despite lack of a clear dose-response relationship.

A subsequent study (Fisher et al., 2001) conducted with Sprague-Dawley rats treated with TCE, TCA and DCA at dose levels as high as 400 mg/kg bw per day failed to reproduce the heart malformations reported in Dawson et al. (1993). However, there were differences in design between the two studies, which may partially account for the incongruence of the results. First, the Fisher et al. (2001) study used soybean oil vehicle, while the Dawson et al. (1993) study used water as a vehicle. Second, the Fisher et al. (2001) study administered a very large dose of TCE (400 mg/kg bw per day) in soybean oil in boluses from gestation days 5 to 16 only, whereas the Dawson et al. (1993) study administered TCE in drinking water at relatively lower doses (maximum 1100 ppm, or 129 mg/kg bw per day) ad libitum either during the entire gestation period (gestation days 1-21) or prior to and throughout pregnancy; both the form of test agent and the timing of the dosage may partially account for the variations between the two studies. Third, the Fisher et al. (2001) study had a very high background incidence of heart malformations (on a per litter basis) among the soybean oil control fetuses (52%), a rate much higher than the incidence of heart malformations in the parallel water controls (37%), whereas the Dawson et al. (1993) study reported a much lower incidence of heart malformations (25% on a per fetus basis) in the water control fetuses; the high background incidence of heart malformations associated with the TCE vehicle controls in the Fisher et al. (2001) study might have masked the effects in the TCE treatment groups. Finally, it is also possible that slight strain differences in the Sprague-Dawley rats and differences in the purity of the test agents used may account for the incongruent findings in the two studies. Curiously, the Fisher et al. (2001) study failed to reproduce heart malformations in animals treated with high doses of TCA or DCA, which had been previously shown to cause heart malformations in Sprague-Dawley rats (Johnson et al., 1998a,b) and Long Evan rats (Smith et al., 1989, 1992; Epstein et al., 1992).

A recent developmental toxicity study by Johnson et al. (2003) used a study design and experimental protocol similar to those in the Dawson et al. (1993) study and was able to corroborate the treatment-related heart malformations reported in Dawson et al. (1993). In that study (Johnson et al., 2003), pregnant Sprague-Dawley rats were exposed to TCE throughout pregnancy. There was a significant increase in the percentage of abnormal hearts in the treated groups. The number of litters with abnormal hearts ranged from 0 to 66.7%, while 16.4% of control litters had abnormal hearts (Table 5). Although this study appears to suggest the presence of a dose-response, with the effects beginning to manifest at a dose of 250 µg/L (0.048 mg/kg bw per day) and a NOAEL at 2.5 µg/L (0.00045 mg/kg bw per day), the dose-response is not as clear as might first appear on closer examination of the data.

While the study authors' conclusion that their data support the cardiac teratogenicity of TCE seem quite reasonable, their assertion that the threshold is below 250 µg/L seems less sure when the dose-response is closely scrutinized. While the authors do point out that the doses, even the no-effect dose, are well in excess of those in epidemiological studies, there is still a need for more data, perhaps with larger dose groups and a wider range of dose levels. However, this endpoint, which results from very short term (acute) exposure, deserves close scrutiny and is chosen as the critical endpoint on the basis of the currently available data.

8.2.6 Mode of Action of TCE

The similarity between carcinogenic effects induced by the parent compound and metabolites supports the conclusion that TCE metabolites are mostly responsible for the liver and kidney tumours observed in TCE bioassays. This is particularly true for renal cell carcinoma, with additional supporting evidence of human GST isozyme dependence and DNA adducts formed from genotoxic DCVC metabolites. TCE-induced human renal carcinomas potentially have a mode of action of VHL tumour suppressor gene mutation followed by induction of neoplasia (Bruning et al., 1997a). Indeed, multiple mutations of the VHL tumour suppressor genes, primarily C to T changes, including nucleotide 454, were found in renal carcinoma patients with high prolonged TCE exposure (Bruning et al., 1997b; Brauch et al., 1999). These findings augment the characterization of exposure to TCE at high levels as highly likely to produce kidney cancer in humans.

The complexity of TCE metabolism and clearance complicates the identification of a metabolite that could be identified as responsible for TCE-induced effects. More than one mode of action may explain TCE-induced carcinogenicity, and several hypotheses have been put forward. In all likelihood, a number of events would be significant to tumour development in the rodent under bioassay conditions. Uncertainty exists, however, as to which events may be more relevant to human exposure to TCE at environmental levels.

It has been considered that mouse liver carcinogenesis arises in parallel with peroxisome proliferation (PP) in the liver by TCE metabolites. Although PP has been correlated with carcinogenesis, the actual mechanism of carcinogenesis as it relates to PP is unknown (Bull, 2000). PP is more substantial in mice than in rats (Bogen and Gold, 1997). The prevailing view of TCE-induced mouse liver carcinogenesis has been that these tumours arise in parallel with PP in the liver by TCE metabolites (Elcombe, 1985; Elcombe et al., 1985; Goldsworthy and Popp, 1987; Melnick et al., 1987; DeAngelo et al., 1989; Cattley et al., 1998). However, the role of PP has been questioned as a mechanism of action for human liver carcinogenesis. As PP has not been observed in humans, agents that produced this result in the rodent would be unlikely to present a liver carcinogenic hazard to humans.

Modification of cell signal pathways by TCA and DCA, resulting in alterations in cell replication, selection and apoptosis (programmed cell death), is likely an important contributor to the hepatocarcinogenicity of TCE and its metabolites (Bull, 2000). The ability of TCA to activate the peroxisome proliferator activated receptor (PPAR) and the subsequent cascade of responses, including effects on gene transcription, is an example of cell signalling. DCA exposure has additionally been shown to influence other cell signalling pathways, and observed perturbations provide insight on mode-of-action hypotheses regarding induction of DCA tumours.

The potential for PP to play a role in TCE-induced kidney toxicity has been assessed and is considered unlikely (Lash et al., 2000). While TCE has been reported to cause PP in rat and mouse kidney, with mice showing a greater response, TCE has not been shown to induce kidney cancer in mice. In addition, studies indicate that renal peroxisomes are generally less responsive to peroxisome proliferators than hepatic peroxisomes (Lash et al., 2000).

Alpha-2u globulin is a major component of urinary protein unique to male rats, and its accumulation was previously considered to contribute to TCE-induced kidney tumours. More recent information indicates that TCE does not cause α_2u globulin accumulation (Goldsworthy et al., 1988). In addition, TCE has been identified as causing kidney damage in both male and female rats (Barton and Clewell, 2000). As such, α_2u globulin accumulation does not appear to be a mode of action of TCE-induced kidney toxicity, as was previously thought.

The cysteine and GSH intermediates formed during the metabolism of TCE -- DCVC and DCVG -- have been shown to be capable of inducing point mutations in Salmonella genotoxicity assays. Furthermore, DCVC induces the expression of proto-oncogenes, including c-jun, c-fos and c-myc, in mouse liver tumours (Tao et al., 2000a,b). The proto-oncogene c-myc is believed to be involved in the control of cell proliferation and apoptosis, which also points towards epigenetic mechanisms for the induction of liver tumours in mice. The cysteine intermediate DCVC has also been shown to induce DNA double-strand breaks and unscheduled DNA synthesis in LLC-PK₁ cells (Lash et al., 2000). There is also evidence that DCVC and DCVG can induce primary DNA damage in mammalian cells (OEHHA, 1999). Other evidence supports the cytotoxic mode of action. Most rats chronically exposed to TCE in the National Cancer Institute and National Toxicology Program bioassays developed toxic nephrosis, and more than 90% of rats (and mice) developed cytomegaly, which was most evident in male rats. Associated with these findings, kidney tumours were increased only in male rats. The TCE conjugates 1,2-DCVC and S-(2,2-dichlorovinyl)-L-cysteine (2,2-DCVC) and the corresponding mercapturic acids -- N-acetyl-S-(1,2-dichlorovinyl)-L-cysteine (1,2-DCVNac) and N-acetyl-S-(2,2-dichlorovinyl)-L-cysteine (2,2-DCVNac) -- are rodent, and possibly human, nephrotoxicants. These compounds can produce proximal tubular necrosis and other lesions in rat kidney after conversion to reactive mutagenic intermediates by cytosolic cysteine conjugate β-lyase (Goeptar et al., 1995).

It is thought that TCE-induced kidney tumours may occur as a result of cellular necrosis and activation of repair processes that lead to cellular proliferation. Study into this mode of action has also focused on DCVG and DCVC. These metabolites, through the β-lyase enzyme or other enzymatic processes, lead to the production of reactive species, which may be responsible for nephrotoxicity (Lash et al., 2000; Vaidya et al., 2003). The reactive species can lead to mitochondrial dysfunction, protein or DNA alkylation and oxidative stress. These effects lead to additional cytotoxic effects as well as repair and proliferative responses along a continuum that may ultimately result in tumorigenesis (Lash et al., 2000; Vaidya et al., 2003). The in vivo formation of DCVG and DCVC in animals and humans indicates that this mode of action may be relevant to assessing the mode of action in humans. While cytotoxicity may play an important role in TCE-induced kidney cancer in rodents, it is uncertain what role it plays in human cancers induced by TCE at exposure levels below those expected to cause frank kidney toxicity.

It has also been hypothesized that formic acid plays a role in kidney toxicity (Green et al., 1998). Increased excretion of formic acid occurs with exposure to TCE and may be related to folate deficiency. Kidney toxicity has been reported in humans and rabbits with exposure to foric acid. However, data indicating that formic acid induces kidney tumours are lacking (Bogen and Gold, 1997).

The accumulation of the TCE metabolite CH is thought to be the cause of TCE lung carcinogenicity, as CH exposure results in lung lesions identical to TCE-induced tumours (Green et al., 1997; Green, 2000). The accumulation of CH in the Clara cells of the lung is thought to lead to lung tumours by causing cell damage and compensatory cell replication, which in turn leads to tumour formation (Green et al., 1997; Green, 2000). It is thought that the mechanism by which CH results in tumour formation in animals may not be pertinent to humans, as there is little CYP2E1 activity in human lungs (Green et al., 1997; Green, 2000). Lung tumours were induced in female mice following exposure to TCE (Odum et al., 1992). A specific lesion, characterized by vacuolization of Clara cells, was seen only in mice, and mice exposed to 100 ppm chloral in air had similar lesions. Only mild effects were seen with inhaled TCOH, and none with intraperitoneally administered TCA. These results suggest that acute lung toxicity of TCE may be due to accumulation of chloral in Clara cells in mice. Since chloral is also genotoxic, the toxicity observed with intermittent exposures is likely to exacerbate any genotoxic effect through compensatory cell proliferation in rodents.

In conclusion, the mode of action for tumour induction by TCE may be attributed to non-genotoxic processes related to cytotoxicity, PP and altered cell signalling; genotoxic processes, such as the production of genotoxic metabolites, including chloral and DCVC; or the production of reactive oxygen species related to peroxisomal induction in the liver. The potential role of several mutagenic or carcinogenic metabolites of TCE cannot be ignored, particularly given the supporting evidence of human DNA adducts formed from genotoxic DCVC metabolites and the evidence of VHL tumour suppressor gene mutation in TCE-exposed kidney cancer patients (Bruning et al., 1997a).

Information on the mode of action for non-cancer effects of TCE is more limited, and support for hypotheses is largely based on observations of common activities with other agents. The major endocrine system effects associated with TCE exposure include the development of testicular (Leydig cell) tumours in rats (Maltoni et al.., 1988; NTP, 1988). TCE and its metabolites TCA and TCOH have been found to partition in the male reproductive organs of rats following inhalation exposure (Zenick et al.., 1984). The same compounds have been identified in seminal fluids of humans occupationally exposed to TCE (Forkert et al., 2003).

Generally, agents that affect steroid hormone levels, such as testosterone, estradiol and luteinizing hormone, will also induce Leydig cell tumours in the rat (Cook et al., 1999). Peroxisome proliferating chemicals have been shown to induce Leydig cell tumours via a modulation of growth factor expression by estradiol (Cook et al., 1999). Peroxisome proliferating chemicals induce hepatic aromatase activity, which can increase serum and testis estradiol levels. The increased interstitial fluid estradiol levels can modulate growth factors, including transforming growth factor-α (TGFα), and stimulate Leydig cell proliferation (Cook et al., 1999). Since steroid hormones are regulated through the hypothalamic-pituitary-testis axis in both rats and humans, agents that induce Leydig cell tumours in rats by disruption of this axis may pose a hazard to humans (Cook et al., 1999). The occurrence of Leydig cell tumours in rats exposed to TCE may therefore act as a signal for disturbance of the endocrine system and be indicative of potential endocrine disturbances in humans. The effects of endocrine disruption in human populations exposed to TCE are an area for more research.

Studies of the mode of action hypotheses for observed developmental effects seen with TCE, TCA and DCA exposure and data specific to TCE exposure are also scant. Developmental effects that have been associated with TCE or TCE metabolite exposure include eye defects (microphthalmia and anophthalmia) in rats and cardiac defects in rats and humans. Microphthalmia has been reported in human offspring with maternal alcohol and retinoic acid exposures. Both retinoic acid and ethanol have, in common with TCE, peroxisome receptor activity. It is possible that PPARα activation may be important to the development of eye anomalies following TCE exposure, although no data currently support this hypothesis (Narotsky and Kavlock, 1995; Narotsky et al., 1995).

The mode of action for TCE-induced cardiac teratogenicity is being evaluated as to whether the gene expression critical for normal heart development is affected during cardiogenesis. Treatment with TCE (equivalent to 110 ppm) produced a dose-dependent inhibition of mesenchymal cell transformation (a critical event in development of the heart) in progenitors of the valves and septa in the heart in vitro (Boyer et al., 2000). Although debate continues regarding the experimental evidence linking observed cardiac anomalies in the developmental assays, TCE appears to affect events important to the development of the heart, events that are consistent with an induction of cardiac anomalies (Boyer et al., 2000).

The TCE metabolites TCA and DCA both produce cardiac anomalies in rats (Smith et al., 1989, 1992; Epstein et al., 1993; Johnson et al., 1998a,b). DCA also concentrates in rat myocardial mitochondria (Kerbey et al., 1976), freely crosses the placenta (Smith et al., 1992) and has known toxicity to tissues dependent on glycolysis as an energy source (Stacpoole et al., 1979; Katz et al., 1981; Yount et al., 1982; Cicmanec et al., 1991). More research into TCE and its metabolites is needed to more fully elucidate possible modes of action for the effects observed in standard developmental protocols.

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