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4 Review of Toxicologic Studies This chapter summarizes findings of animal studies of trichloroethylene (TCE) and tetrachloro- ethyle (perchloroethylene, PCE) toxicity and relevant end points. The review was based in part on previ- ously published comprehensive reviews on the two chemicals of interest, but numerous published studies were reviewed individually in greater detail. Studies were examined according to criteria that reflected robustness of study design related to the hypothesis being tested and that included such characteristics as number of animals tested, measurement methods used, appropriateness of statistical methods, and con- cordance of conclusions with data presented. Studies substantially lacking in some of or all those and other measures of study quality and studies whose outcomes were not able to be repeated in later studies or in other laboratories were given less weight in the evaluation. Salient findings on principal health end points are summarized by chemical and organ system. The administered doses or the doses associated with the no-observed-adverse-effect levels (NOAELs) or the lowest-observed-adverse-effect levels (LOAELs) are reported when possible. At the conclusion of this toxicologic review, a hazard evaluation of TCE and PCE exposure at Camp Lejeune was conducted for selected health end points. A hazard evaluation is conducted to provide information on the intrinsic toxic potential of an exposure and is not meant to provide a quantitative risk assessment. As noted in Chapter 2, the committee identified nine volatile organic compounds (VOCs) of con- cern. To manage the vast amount of information on each, we provide different degrees of review accord- ing to the findings from the exposure assessment regarding the frequency and concentrations of the con- taminants in the affected drinking-water systems. This chapter presents detailed toxicologic evaluations of the two chemicals of greatest concern, TCE and PCE. Information on the metabolism of TCE and PCE and factors that influence their toxicity was presented in Chapter 3 and is drawn upon in this chapter. Chapter 7 provides an integrated discussion of the toxicologic evidence in context with the epidemiologic evidence on TCE and PCE. For completeness of the literature review, Appendix D provides brief reviews of the toxicologic data on the seven other chemicals. TRICHLOROETHYLENE Data on the toxicity of TCE were summarized in a report by the National Research Council (NRC 2006). In some cases, more recent literature reviews on particular subjects were available (e.g., Lamb and Hentz 2006; Watson et al. 2006), and they were relied on for defining the body of literature available up to the time of publication. In addition, a literature search of Medline was done to determine whether any relevant new publications were available. Conclusions drawn for the present report were based on a re- view of the body of available peer-reviewed literature. Because TCE and PCE have some of the same me- tabolites and effects, salient finding of studies of PCE are discussed in relevant sections of the TCE re- view. More detailed review of the PCE literature is provided later in the chapter. To facilitate a comparison of the toxicologic data with the epidemiologic data in Chapter 7, the toxicologic data are pre- 90
Review of Toxicologic Studies 91 sented below according to organ system and in some sections divided to consider toxic effects separately from carcinogenic effects. Hepatic Effects Toxicity TCE, even in high doses, produces only a modest degree of injury of hepatocytes in laboratory animals. Klaassen and Plaa (1966) compared the acute hepatotoxicity of TCE with that of other common halogenated aliphatic hydrocarbons (halocarbons) in male mice dosed by intraperitoneal injection. The dose of TCE required to produce an increase in serum alanine-aminotransferase activity, 1.6 mL/kg, was almost as high as the dose that was lethal in 50% of test animals, 2.2 mL/kg. Oxidative stress was as- sessed by measuring thiobarbituric-acid-reactive substances in the livers of male Fischer rats that received one intraperitoneal injection of TCE at 0, 100, 500, or 1,000 mg/kg (Toraason et al. 1999). Thiobarbituric- acid-reactive substances were increased in the 500- and 1,000-mg/kg groups. Hepatic concentrations of 8- hydroxy-2â²-deoxyguanosine adducts, induced in DNA by oxygen-based radicals, were also increased at 500 mg/kg and presumably at 1,000 mg/kg. It should be recognized that the 500- and 1,000-mg/kg doses produced stage II and stage III-IV anesthesia, respectively. Channel et al. (1998) gave male B6C3F1 mice TCE at 0, 400, 800, or 1,200 mg/kg in corn oil by gavage 5 days/week for 8 weeks. Transient increases in cell and peroxisome proliferation, centered around day 10, were observed only at the highest dose. There were no differences from controls in the incidences of hepatocellular apoptosis or necrosis. Thiobarbi- turic-acid-reactive substances were significantly increased in the groups treated with TCE at 800 and 1,200 mg/kg on days 6-14. 8-Hydroxy-2â²-deoxyguanosine adducts in liver DNA were significantly in- creased throughout much of the study with TCE at 1,200 mg/kg. Buben and OâFlaherty (1985) saw a modest increase in serum alanine aminotransferase and decrease in hepatic glucose-6-phosphatase activity in mice given TCE at 500 mg/kg or greater in corn oil by gavage five times a week for 6 weeks. Mice re- ceiving as little as 100 mg/kg per day had an increase in relative liver weight. It is clear that TCE, even when given repeatedly to mice and rats at narcotic doses, has little ability to damage hepatocytes. Adverse effects of TCE on the liver are usually attributed to metabolites of the cytochrome P- 450-mediated oxidative pathway (Bull 2000). Buben and OâFlaherty (1985) reported that plots of their mouse subchronic-hepatotoxicity data against urinary-metabolite excretion values indicated that TCEâs effects are directly related to the extent of its metabolism. As described in Chapter 3, TCE is oxidized by cytochrome P-450s (notably CYP2E1 at low to moderate TCE doses) to chloral, which is converted to chloral hydrate. That intermediate has a short half-life; it is rapidly oxidized to trichloroacetic acid, which is reduced to trichloroethanol (Lash et al. 2000a). Relatively small amounts of dichloroacetic acid may arise from trichloroacetic acid or other metabolites. Induction of CYP2E1 in rats with pyridine increases the toxicity of TCE to isolated rat hepatocytes (Lash et al. 2007). High concentrations of trichloroacetic acid and dichloroacetic acid are not toxic to hepatocytes freshly isolated from B6C3F1 mice (Bruschi and Bull 1993); the researchers proposed that trichloroacetic acid and dichloroacetic acid cause peroxisome proliferation and the ensuing generation of reactive moieties that deplete glutathione and can cause oxida- tive injury. Dichloroacetic acid does not induce peroxisome proliferation in male B6C3F1 mice in the same dose range at which it produces hepatic tumors (DeAngelo et al. 1999). Laughter et al. (2004) found that high oral doses of TCE increased liver weight, peroxisome proliferation, and hepatocellular prolifera- tion in male mice. Those effects appeared to be due primarily to trichloroacetic acidâs activating a nuclear protein known as the peroxisome-proliferator-activated receptor alpha (PPARÎ±). PPARÎ±-dependent changes seen in gene expression may contribute to the carcinogenicity of TCE in mouse liver. TCE-induced hepatic injury is not a common finding in humans and was rarely reported in pa- tients when TCE was used as an anesthetic (Lock and Reed 2006). Clearfield (1970) described hepatocel- lular degeneration in two men who intentionally inhaled extremely high vapor concentrations of TCE for their intoxicating effects. In contrast, James (1963) saw only small foci of fatty accumulation in the liver
92 Contaminated Water Supplies at Camp LejeuneâAssessing Potential Health Effects (steatosis) of a man who died after 10 years of TCE abuse. Bruning et al. (1997) found renal injury but no evidence of hepatotoxicity in a man rendered unconscious for 5 days by drinking about 70 mL of TCE in a suicide attempt. Pembleton (1974) reported a transient postoperative rise in serum aspartate aminotrans- ferase activity in four of 100 patients anesthetized with TCE for surgical procedures. A study of 289 Brit- ish workers who experienced central nervous system (CNS) symptoms from TCE inhalation and dermal exposure in the workplace revealed no clear diagnoses of hepatotoxicity (McCarthy and Jones 1983). Such findings over the last 50 years indicate that acute or repeated high-dose exposures to TCE will pro- duce a modest degree of hepatic injury in some people but not in most people (ATSDR 1997a). Cancer The carcinogenic effects of TCE and its metabolites have been assessed in a number of lifetime studies of several strains of mice and rats (NCI 1976; Fukuda et al. 1983; Henschler et al. 1984; Maltoni et al. 1986; NTP 1988, 1990a). Results of studies of TCE induction of hepatic tumors in rodents are summarized here on the basis of the extensive National Research Council review (NRC 2006). It has been well established that TCE, when administered chronically in very high doses by ga- vage, can produce an increased incidence of hepatocellular cancer in B6C3F1 mice. In the original bioas- say (NCI 1976), technical grade TCE (containing epichlorohydrin and 1,2-epoxybutane as stabilizers) had this effect. Concern that these stabilizers are well-established mutagens and contributed to TCEâs appar- ent carcinogenicity led scientists to utilize TCE without these stabilizers in future bioassays. Henschler et al. (1984) saw no increase in liver tumors in either sex of Swiss ICR/HA mice, rats, or Syrian hamsters that inhaled highly-purified TCE (stabilized with 0.0015% triethanolamine) for 18 months. Exposure of male and female B6C3F1 mice to epichlorohydrin-free TCE by corn oil gavage at 1,000 mg/kg/day for 2 years caused increases in hepatocellular carcinoma. No such increase in liver tumor incidence was mani- fest in F344/N rats (NTP 1990a). Another study of four additional strains of rats of both sexes ingesting epichlorohydrin-free TCE at 125-1,000 mg/kg also showed no increase in liver tumors (NTP 1988). Thus, it has been demonstrated that TCE itself, when administered chronically in very high oral doses, results in an increased incidence of liver cancer limited to male and female B6C3F1 mice. The major oxidative metabolites of TCEâtrichloroacetic acid, dichloroacetic acid, and chloral hydrateâhave also been extensively studied in rodents (Herren-Freund et al. 1987; Bull et al. 1990; DeAngelo et al. 1991, 1996, 1997, 1999; Daniel et al. 1992, 1993; Pereira 1996; George et al. 2000; NTP 2002a,b,c; Leakey et al 2003). Trichloroacetic acid is a species-specific carcinogen that induces perox- isome proliferation and hepatocellular carcinomas when administered in drinking water to male and fe- male B6C3F1 mice (B6C3F1 mice are particularly susceptible) (Herren-Fruend et al. 1987; Bull et al. 1990; DeAngelo et al. 1991). The blood concentration of trichloroacetic acid required to induce hepatic tumors in mice is in the millimolar range. Effects have been observed with drinking-water concentrations of trichloroacetic acid of 0.05-5 g/L. TCA did not induce hepatic tumors in male F344 rats under similar treatment conditions (Daniel et al. 1993; DeAngelo et al. 1997). B6C3F1 mice produce a large amount of trichloroacetic acid after exposure to TCE relative to unresponsive mouse strains (see Chapter 3). Tri- chloroacetic acid increases the rate of hepatocellular proliferation, production of reactive oxygen species, hepatocellular hyperplasia, and hepatomegaly (see Chapter 3). Marked species differences in susceptibil- ity to peroxisome proliferation associated with liver cancer after increased fatty-acid beta oxidation and modulation of hepatocellular replication related to activation of the PPARÎ± nuclear receptor by TCE and its metabolites have been investigated and reviewed in detail (Klaunig et al. 2003; Cattley 2004; Laughter et al. 2004). Rats exhibit saturation of TCE oxidative metabolism that results in amounts of trichloroacetic acid that are probably insufficient to induce hepatic peroxisome proliferation. It is thought that humans, like rats, have lower rates of oxidative metabolism and higher rates of conjugation than do mice. Trichloroacetic acid produces hepatic tumors only in B6C3F1 mice, but dichloroacetic acid in- duces them in mice and in F344 rats at exposures up to 5 g/L in drinking water for 104 weeks (Herren- Freund et al. 1987; Bull et al. 1990; Daniel et al. 1992; DeAngelo et al. 1996, 1999; Pereira 1996; NRC
Review of Toxicologic Studies 93 2006). Dichloroacetic acid is a major metabolite of TCE in B6C3F1 mice but a minor metabolite in Spra- gue-Dawley rats (Larson and Bull 1992). Marked liver enlargement and cytomegaly in dichloroacetic acid-treated mice also indicate that induction of hepatic tumors depends on stimulation of increased cell division secondary to hepatoxoic damage (Bull et al. 1990). Inhibition of dichloroacetic acid metabolism by the parent compound at less than 1 to 500 ÂµM (Kato-Weinstein et al. 1998) is thought to contribute to the variation in mouse hepatic tumors observed at this dose range (Bull et al. 2002). Choral hydrate induces hepatic tumors in male B6C3F1 mice but not in female mice or F344 male rats (George et al. 2000; NTP 2002a,b; Leakey et al. 2003). Female B6C3F1 mice given choral hydrate in water by oral gavage for 104 weeks at up to 100 mg/kg per day had no increase in hepatic tumors (NTP 2002a), whereas exposure at the same doses in two groups of male mice fed ad libitum (NTP 2002a,b) or fed a calorie-controlled diet (Leakey et al. 2003) had increased incidences of hepatocellular adenoma or carcinoma (combined). Dietary control of caloric intake in the latter study was thought to improve sur- vival and to decrease interassay variation. Choral hydrate is metabolically converted to trichloroacetic acid or dichloroacetic acid, and this contributes to its weak carcinogenicity. Overall, choral hydrate is an ineffective hepatic carcinogen that induces tumors only in male mice. An epidemiologic study was conducted of short-term clinical exposure to choral hydrate used as a hypnosedative and possible cancer risk in humans (Haselkorn et al. 2006). An increasing risk of prostatic cancer with chloral hydrate was found, but the trend wat not statistically significant. Thus, the authors concluded that there was no persuasive evidence of a causal relationship between choral hydrate exposure and cancer in humans, but they were unable to rule out a causal relationship because statistical power was low. Trichloroacetic acid elicits hepatic tumors in mice with a phenotype typical of peroxisome prolifera- tors, whereas dichloroacetic acid produces hepatic tumors with a distinctly different phenotype and also increases tumor growth (Bull 2000; Thai et al. 2003). The relevance of TCE- and PCE-induced hepatic tumors to humans has been the subject of a great deal of research. Oral and inhalation carcinogenicity bioassays of TCE in rodents have shown that adenocarcinomas are strain- and species-specific (that is, are limited to the B6C3F1 mouse). Haseman et al. (1998) reported a spontaneous hepatic-tumor incidence of 42.2% in male control B6C3F1 mice used in National Toxicology Program (NTP) studies. The NTP recently held a series of workshops to determine whether another mouse strain and a rat strain should be adopted. In light of the high background hepatic- tumor incidence, it was recommended that the NTP explore the use of multiple mouse strains (King- Herbert and Thayer 2006). It has been clearly established that the toxicokinetics (target-organ dosimetry) of TCE and PCE of the mouse and the human are different (see Chapter 3). Mice absorbed substantially more TCE and PCE because of their greater respiratory and alveolar ventilation rate, cardiac output and pulmonary blood flow rate, and blood:air partition coefficient. Mice also metabolically activate substantially more of their ab- sorbed doses to bioactive substances (Lipscomb et al. 1998). On an equivalent inhalation exposure to PCE, rats exhibited markedly higher blood and urinary concentrations of trichloroacetic acid and di- chloroacetic acid than humans (Volkel et al. 1998). The ratsâ blood also contained much higher concen- trations of protein adducts (Pahler et al. 1999). Physiologically based toxicokinetic models similarly pre- dict that mice will produce higher target-organ (liver) doses of trichloroacetic acid than humans after exposure to PCE (Clewell et al. 2005) and TCE (Clewell and Andersen 2004). The primary mode of action of trichloroacetic acid, and to a smaller extent dichloroacetic acid, is activation of PPARÎ±. Stimulation of PPARÎ± can enhance DNA replication, resulting in expansion of some clones of hepatocytes and suppression of apoptosis, so initiated and precancerous cells will be spared. Male wild-type mice dosed orally with TCE exhibit hepatocyte proliferation and changes in ex- pression of genes involved in cell growth (Laughter et al. 2004). PPARÎ±-null mice are refractory to those effects, which are associated with carcinogenesis. Mice expressing human PPARÎ± fail to show increases in markers of cell proliferation and are resistant to liver cancer if treated with PPARÎ± agonists (Morimura et al. 2006; Yang et al. 2008). The concentration of PPARÎ± in human cells is about 10% of that in the liv- ers of rodents (Palmer et al. 1998; Klaunig et al. 2003; Lai 2004). The interpretation of mouse hepatic- tumor induction in 2-year bioassays relative to the inducing compoundâs mode of action, including induc-
94 Contaminated Water Supplies at Camp LejeuneâAssessing Potential Health Effects tion of peroxisome proliferation, has been assessed in a human-relevance framework (Cohen et al. 2003, 2004; Meek et al. 2003; Holsapple et al. 2006; Meek 2008). The relevance of B6C3F1 mouse hepatic tu- mors to humans is also weakened by the observations that the background incidence of hepatic tumors in unexposed B6C3F1 mice is about 60% and that large numbers of chlorinated compounds induce such tu- mors in mice (Gold and Slone 1995). The human is likely to be much less responsive toxicodynamically than the mouse to the cellular effects of trichloroacetic acid and dichloroacetic acid. Many toxicologists have judged that the mode of action for hepatic carcinogenesis observed in mice after administration of peroxisome-proliferation-inducing drugs and other chemicals (such as TCE and PCE) makes it unlikely that such chemicals pose a hepatic-cancer risk in humans (Cattley et al. 1998; NTP 2000; Clewell and Andersen 2004; NRC 2006; Klaunig et al. 2007). It was concluded by the Na- tional Research Council that the PPARÎ± mode of action for liver cancer in mice is not relevant to humans (NRC 2006). However, others have raised questions about the interpretation of PPARÎ± actions and whether it is the only relevant mode of action for such chemicals (Keshava and Caldwell 2006), and this continues to be a subject of active debate (Peters et al. 2005; Klaunig et al. 2007; NRC 2008). Toxicodynamic mechanisms of hepatic carcinogenicity other than peroxisome proliferation have been explored. Both trichloroacetic acid and dichloroacetic acid apparently contribute to hepatic tumori- genesis in mice (Bull et al. 2002; Caldwell and Keshava 2006). High, repeated doses of those TCE and PCE metabolites initially stimulate and then depress the growth of normal liver cells (Bull 2000). That may confer a growth advantage on initiated cells. Dichloroacetic acid at high concentrations also appears to act by increasing the clonal expansion and decreasing apoptosis of such precancerous cells. Moderate amounts of dichloroacetic acid are apparently produced from trichloroacetic acid and trichloroethanol in mice, but only trace amounts were found in one of three studies of TCE-exposed humans (see Chapter 3). It is important to recognize that stimulation or inhibition of cell growth through PPARÎ± activation ceases when the metabolites are eliminated (Miller et al. 2000). Thus, such alteration of cell signaling is not a genotoxic mechanism of action. Very high concentrations of dichloroacetic acid and chloral hydrate have a weak genotoxic action in vitro. Bull (2000) and Moore and Harrington Brock (2000), however, con- clude that it is unlikely that those metabolites would cause tumors in any organ through genotoxocity or mutagenicity at exposure concentrations relevant to humans. Renal Effects Toxicity TCE has limited capacity to produce renal injury in rodents that are subjected to high oral expo- sures for extended periods. Jonker et al. (1996), for example, gave female Wistar rats TCE at 500 mg/kg by corn-oil gavage for 32 consecutive days. Urinalyses showed only slight increases in N-acetyl-Î²- glucosaminidase and alkaline phosphatase activities. A comparable exposure to PCE produced somewhat larger increases. Kidney weights were modestly increased by both chemicals. Microscopic examination revealed multifocal areas of vacuolation and karyomegaly in the animalsâ renal tubules. Male Eker rats received TCE at 50, 100, 250, 500, or 1,000 mg/kg by corn-oil gavage 5 times a week for 13 weeks (Mally et al. 2006). There were no changes in Î³-glutamyltransferase activity or other urinary indexes of renal cytotoxicity. There was tubular-cell proliferation at 250 mg/kg or greater and karyomegaly at 500 mg/kg or greater. Overt nephropathy was restricted to the 1,000-mg/kg group. Nephropathy has been a common finding in rats and mice in chronic, high-dose cancer bioassays of TCE (NCI 1976; NTP 1986a, 1988, 1990a). Nephrosis and cytomegaly were more severe in the rats than in the mice, and male rats were generally affected more severely than females. Cytomegaly was manifested as frank enlargement of the cytoplasm and the nucleus of scattered tubular cells in the inner cortex and outer stripe of the medulla. Karyomegaly was later observed in the proximal tubular epithelial cells of the pars recta. The affected tubules were dilated, and the cells were flattened and elongated and contained enlarged, hyperchromatic
Review of Toxicologic Studies 95 nuclei with irregular shapes. A low incidence of renal tumors was seen consistently in several strains of male rats in the bioassays. TCE has also been found to have some adverse renal effects when inhaled acutely or repeatedly at high concentrations for long periods. Proximal tubular damage was reported in male F344 rats exposed for 6 h to TCE vapor at 1,000 or 2,000 ppm (Chakrabarti and Tuchweber 1988). Mensing et al. (2002) subjected male F344 rats to TCE at 500 ppm for 6 h/day 5 days/week for 6 months. Glomerulonephritis was seen on histopathologic examination, but urinary biomarkers of glomerular damage were not found. Increases in urinary N-acetyl-Î²-glucosaminidase and low-molecular-weight proteins reflected mild proximal tubular damage. Adverse effects of TCE on the kidneys are due largely to metabolites formed via the glutathione conjugation pathway (Lash et al. 2000b). As described in Chapter 3, conjugation of TCE with glutathione to form S-(1,2-dichlorovinyl)glutathione (DCVG) occurs primarily in the liver. DCVG is secreted into bile and blood. That in the bile is converted to S-(1,2-dichlorovinyl)-L-cysteine (DCVC), which is reab- sorbed into the bloodstream. As noted in Chapter 3, humans have a lower capacity than rats to metabolize TCE by the glutathione pathway. Lash et al. (1999) were able to detect DCVG in the blood of humans who had inhaled TCE at 50 or 100 ppm for 4 h, but Bloemen et al. (2001) could not find DCVG or DCVC in the urine of similarly exposed subjects. DCVG in the blood is taken up by the kidneys and me- tabolized to DCVC by Î³-glutamyltransferase and a dipeptidase. Lash et al. (2001b) observed the follow- ing decreasing order of toxic potency in freshly isolated rat cortical cells: DCVC > DCVG >> TCE. DCVC can be detoxified by acetylation and activated further by two pathways: (1) cleavage by renal cy- tosolic and mitochondrial Î²-lyases to dichlorothioketene, which in turn can lose a chloride ion to yield chlorothioketene or tautomerize to form chlorothionacyl chloride (the latter two moieties are very reactive and acylate proteins and DNA), and (2) oxidation by renal cytochrome P-450s or flavin-containing mono- xygenases to the epoxide, DCVC sulfoxide (DCVCS). Lash et al. (1994) reported that DCVCS was a more potent nephrotoxin than DCVC in vitro and in vivo in rats. Apoptosis was observed after as little as 1 h of incubation of cultured human renal proximal tubular cells with DCVC and DCVCS (Lash et al. 2003, 2005). Cellular proliferation accompanied by increased expression of proteins associated with cel- lular growth, differentiation, stress, and apoptosis was also an early response to low doses. Necrosis, however, was a late, high-dose phenomenon in this cell system. Exposure of human renal proximal tubu- lar cells to DCVC at lower concentrations for 10 days also resulted in expression of genes associated with cell proliferation, apoptosis, and stress (Lash et al. 2005) and repair and DCVC metabolism (Lash et al. 2006). Proximal tubular-cell damage, as discussed above, appears to be a prerequisite for renal-cell can- cer. Bruning et al. (1996) observed urinary protein-excretion patterns indicative of tubular damage in all of a group of 17 workers exposed for years to peak TCE vapor concentrations that caused CNS depres- sion. They later reported small increases in urinary excretion of glutathione S-transferase Î± and Î±1- microglobulin in a group of 39 cardboard workers without renal-cell cancer who had been heavily ex- posed to TCE for about 16 years (Bruning et al. 1999). Both indexes are markers of proximal tubular in- jury. Higher Î±1-microglobulin excretion was reported in renal-cell cancer patients with TCE exposure than in renal-cell cancer patients without TCE exposure in an updated study (Bolt et al. 2004). Green et al. (2004) described similar findings in 70 electronics workers who inhaled TCE at an average concentra- tion of 32 ppm for about 4 years. A battery of tests for nephrotoxicity was assessed after 4 days of expo- sure. Urinary albumin and N-acetyl-Î²-glucosaminidase were higher than in controls, although there was no correlation with the magnitude or duration of TCE exposure. There was also a suggested increase in urinary glutathione S-transferase Î± activity that correlated with the intensity but not with the years of ex- posure. Finally, Bruning et al. (1998) evaluated renal damage in a man who ingested about 70 mL of TCE in a suicide attempt. He was rendered unconscious for 5 days. His urinary glucose and protein concentra- tions were normal, but Î±1- and Î²2-microglobulin, N-acetyl-Î²-glucosaminidase, and several low-molecular- weight protein concentrations were increased. Such modest, reversible signs of renal injury demonstrate that TCE, even in extreme exposure conditions, has quite small nephrotoxic potential in humans.
96 Contaminated Water Supplies at Camp LejeuneâAssessing Potential Health Effects Cancer TCE was given in corn oil to F344/N rats and B6C3F1 mice of both sexes by oral gavage at doses up to 1,000 mg/kg in rats and 6,000 mg/kg in mice in a 13-week study and up to 1,000 mg/kg in both spe- cies and sexes in a 103-week study (NTP 1990a). Two-year oral-gavage studies in four additional rat strains were also conducted (NTP 1988). Nonneoplastic renal lesions were found in all animals dosed for 2 years. In all strains of rats tested, cytomegaly and karyomegaly of tubular cells in the renal corticome- dullary region were observed. Frank toxic nephropathy was observed with higher frequency beginning at 52 weeks of exposure. A statistically significant increase in renal-tumor incidence was observed only in male F344/N rats exposed to TCE at 1,000 mg/kg for 2 years (this was the LOAEL). TCE has been shown to cause toxicity in proximal renal tubules in vivo; results of in vitro studies have also indicated toxicity of TCE and its metabolite DCVC in primary cultures of rat tubular cells (Cummings et al. 2000). Nephrotoxicity was reported in Long-Evans rats after 6 months of inhalation exposure to TCE at 500 ppm (Mensing et al. 2002). The urinary-protein profile reported is consistent with impairment of tu- bular reabsorption of filtered protein. Inhalation studies were conducted in both sexes of Sprague-Dawley rats with TCE at 100, 300, and 600 ppm for 2 years and in Swiss mice at 100 and 600 ppm for 78 weeks (Maltoni et al. 1988a). Renal adenocarcinomas were reported in male rats at 600 ppm (the LOAEL), but no renal effects were observed in mice. Cytokaryomegaly or megalonucleocytosis was observed at the end of 2 years of exposure in male rats (77% of the 600-ppm group and 17% of the 300-ppm group) with no indication of pathologic conditions earlier. Inconclusive evidence of induction of Î±2Âµ-globulin by TCE, formic acid formation, or peroxisome proliferation as a mechanism or mode of action of TCE as a renal carcinogen was found (Goldsworthy et al. 1988; Green et al. 2003). Results of animal studies indicate that kidney cancer occurs at high doses (for example, 1,000 mg/kg and 600 ppm) in male rats and is preceded by nephrotoxicity affecting the proximal tubule. An analysis by the U.S. Environmental Protection Agency with pooling across strains indicated a modest tu- mor effect in female rats (EPA 2001). Renal-cell cancers observed in German workers who were highly exposed to TCE have generally been assumed to be due to an initiating genotoxic effect of DCVC or DCVC coupled with the promoting effects of recurring cytotoxicity and compensatory hyperplasia (Brun- ing and Bolt 2000). The complete TCE glutathione conjugation pathway and assumed penultimate nephrotoxic metabolites are described in Chapter 3. It has been proposed that exposures below nephro- toxic concentrations or some threshold of exposure probably pose no risk of cancer in that nephrotoxicity is deemed to be a prerequisite for development of kidney cancer (Bruning and Bolt 2000; Harth et al. 2005). TCE oxidative metabolizing enzymes (such as CYP2E1 and CYP3A5 isoforms) have polymorphic forms. Known human population diversity in bioactivation and detoxification capabilities is an additional consideration in determining the exposure concentration below which nephrotoxicity is unlikely. For TCE inhalation exposure in the occupational setting, the suggested practical threshold below which nephrotox- icity is unlikely to occur is 250 ppm as an 8-h time-weighted average (Harth et al. 2005). In humans, inactivation of the von Hippel-Landau (VHL) tumor-suppressor gene is responsible for the hereditary VHL cancer syndrome. Affected people are predisposed to a variety of tumors; more than 80% of sporadic renal-cell carcinomas are associated with inactivation of this gene. Brauch et al. (2004) noted that renal-cell cancer patients unexposed to TCE did not have the somatic VHL gene muta- tional characteristics of TCE-exposed renal-cell cancer patients. According to Moore and Harrington- Brock (2000), TCE itself has little if any mutagenic potential, and it is unlikely that any TCE-induced tu- mors would be mediated by its major oxidative metabolites. TCE recently also yielded negative results when tested in a Salmonella typhimurium strain (Ames test) that contained DNA coding for cytochrome P-450 reductase, cytochrome b5, and cytochrome P-450 2E1 (Emmert et al. 2006). TCE glutathione- conjugated metabolites DCVG and DCVC have, however, been shown to have genotoxic effects in in vitro test systems. A recent study provides insight into a TCE renal-carcinogenesis threshold proposal. A strain of rats (Eker) uniquely susceptible to renal carcinogens was exposed to TCE at an administered dose of 100,
Review of Toxicologic Studies 97 250, 500, and 1,000 mg/kg by gavage 5 days/week for 13 weeks (Mally et al. 2006). The Eker rat is a unique animal model for renal-cell carcinoma, carrying a germ-line alteration of the Tsc-2 tumor- suppressor gene. Results showed a significant increase in cell proliferation in renal tubular cells but no increased preneoplastic renal lesions or tumor incidence. In vitro studies were conducted on primary Eker rat renal epithelial cells by exposing them to the TCE metabolite DCVC dissolved in water at 10-50 ÂµM for 8, 24, and 72 h. Concentrations of DCVC that reduced rat renal-cell survival to 50% also resulted in cell transformation. No carcinogen-specific mutations were identified in the VHL or Tsc-2 tumor- suppressor genes in the transformed cells. Renal-cell carcinomas in the Eker rat have substantial similari- ties to human renal-cell carcinomas. It is not entirely clear that this or any contemporary experimental animal model adequately mirrors humans with regard to the effects of TCE-induced mutations in the VHL gene, but the authors firmly suggest that TCE-mediated renal carcinogenicity may occur only secondarily to nephrotoxicity and sustained regenerative cell proliferation. The latter findings, coupled with the aforementioned data of Lash et al. (2005, 2006), suggest that renal-cell cancer may result from prolonged, high-dose cytotoxicity and sustained cell proliferation but that TCEâs metabolites may lack initiating ac- tivity. Both DCVC and its mercapturic acid metabolite N-acetyl-S-(1,2-dichlorovinyl)-L-cysteine have been found in urine of humans exposed to TCE, and illustrates that the glutathione conjugation pathway is active (Bernauer et al. 1996). Exposure of volunteers to TCE at 50 or 100 ppm showed that DCVG con- centrations were 3.4 times higher in males than in females (Lash et al. 1999). Genes associated with stress, apoptosis, cell proliferation, repair, and DCVC metabolism were up-regulated almost double in cultured human renal tubular cells exposed to subcytotoxic doses of DCVC for 10 days (Lock et al. 2006). Male rats display higher reduced glutathione conjugation, Î³-glutamyl transpeptidase, and cysteine conjugate Î²-lyase activity than female rats. Taken together, results in the cited studies indicate that male humans and male rats both possess significant glutathione conjugation capacity and can produce the criti- cal TCE metabolite DCVC; renal carcinoma has been observed in male rats and male workers when both have been exposed to high TCE concentrations for prolonged periods of time. These observations show data congruence, indicating that the conjugation pathway plays a central role in induction of renal carci- noma in males of both species. As discussed in Chapter 3, rats have greater capacity to metabolically ac- tivate TCE by this pathway than humans. Evaluation of potential risks to human health related to contaminants in water supplies is a central concern of this project. Given the foregoing, it is sensible to begin to apply recent toxicologic information to contemporary maximum environmental values. In summary, exposure to high TCE concentrations ap- pears to lead to saturation of the oxidative metabolic pathway with an attendant pronounced increase in metabolism via the glutathione-dependent pathway and likely increased production of penultimate toxic metabolites, such as DCVC sulfoxide, chlorothioketene, and thionoacylchloride from DCVC (Dobrev et al. 2002). As previously described, substantially larger quantities of these toxic moieties are produced from TCE by rat kidney than by human kidney. In addition, cultured rat cortical cells have been shown to be more susceptible to DCVC-induced necrosis than cultured human proximal tubular cells (Lash et al. 2001a). Human kidney cells have the capacity to metabolically activate and to respond adversely to low concentrations of DCVC, but not to the extent exhibited by male rat kidneys. Pulmonary Effects Toxicity The pulmonary-toxicity potential of TCE has been studied extensively in mice and rats; there ap- pear to be no reports of TCE-induced lung injury in humans. Forkert et al. (1985) were among the first scientists to describe lung toxicity in mice. Intraperitoneal injection of very high doses of TCE (2,000 and 2,500 mg/kg) into male CD mice rapidly caused damage of bronchiolar Clara cells and alveolar type II cells, anesthesia, and a marked reduction in pulmonary cytochrome P-450 content. Female CD-1 mice
98 Contaminated Water Supplies at Camp LejeuneâAssessing Potential Health Effects inhaling TCE at 20-2,000 ppm 6 h/day for up to 5 days exhibited dose-dependent vacuolation of Clara cells (Odum et al. 1992). Pyknosis of the bronchiolar epithelium also occurred at the higher concentra- tions. No morphologic changes were seen in the lungs of rats that were exposed to TCE vapor at 500 or 1,000 ppm. Isolated mouse Clara cells metabolized TCE to chloral, trichloroacetate, and trichloroethanol, but no trichloroethanol glucuronide was detected. It was proposed that the inability of these cells to con- jugate trichloroethanol with glucuronic acid led to accumulation of chloral to cytotoxic concentrations (Odum et al. 1992; Green 2000). Forkert et al. (2005) found that oxidation of TCE to chloral was cata- lyzed in murine lung microsomes by cytchrome P-450s 2E1, 2F2, and 2B1. Forkert et al. (2006) later demonstrated that bioactivation of TCE by CYP2E1 and CYP2F2 occurred in Clara cells. Dichloroacetyl lysine adducts were localized in Clara cells in the TCE-treated CD-1 mice, and CYP2E1 and CYP2F2 are highly concentrated there (Forkert 1995). It is generally accepted that the cytotoxicity and possibly the weak mutagenicity of chloral and diacetyl chloride contribute to the development of lung tumors in mice. The mouse appears to be uniquely sensitive to TCE-induced pulmonary toxicity and cancer. Mice, but not rats, developed lung tumors in the inhalation bioassays conducted by Fukuda et al. (1983) and Maltoni et al. (1988a). Clara cells are numerous and present throughout the airways of mice. They are found much less frequently in rats and are rare in humans (Green 2000). Mouse Clara cells contain con- siderable amounts of smooth endoplasmic reticulum, a membrane network in which cytochrome P-450s are bound. Human Clara cells are largely devoid of this organelle. Accordingly, metabolic activation of TCE to chloral is high in mouse, much lower in rat, and undetectable in human microsomes (Green et al. 1997b). Green et al. measured high CYP2E1 concentrations in mouse lung microsomes; concentrations of CYP2E1 were lower in rats and undetectable in humans. Mace et al. (1998), however, were able to detect very low concentrations of CYP2E1 mRNA and protein in human peripheral lung tissue. Forkert et al. (2005) found that male CD-1 mouse lung microsomes efficiently metabolize TCE to chloral hydrate, whereas the reaction was observedâat low ratesâin samples from only three of eight human donors. Those findings suggest that TCE poses only a minimal risk of pulmonary toxicity in humans. Cancer TCE inhalation exposure caused an increased incidence of pulmonary tumors in ICR, Swiss, and B6C3F1 mice but not in rats or hamsters. When female ICR mice were exposed to TCE at 150 and 450 ppm 7 h/day 5 days/week for 104 weeks followed by an observation period of 3 weeks, lung-tumor inci- dence increased by a factor of 3 (Fukuda et al. 1983); epichlorohydrin was used as a TCE stabilizer in this experiment. Female Sprague-Dawley rats exposed at the same concentrations for the same period had no increase in lung tumors. Male Sprague-Dawley rats had no increase in lung tumors but did have an in- crease in testicular and renal tumors after exposure to TCE at 600 ppm for 104 weeks but not at 100 or 300 ppm (Maltoni et al. 1986). Excess lung tumors were observed in Swiss mice and B6C3F1 mice ex- posed to TCE at up to 600 ppm for 78 weeks (Maltoni et al. 1988a). Five gavage studies were also re- viewed for induction of lung tumors in several strains of rats and mice; no excess lung tumors were found (NRC 2006). These results, the information presented in the preceding section on pulmonary toxicity, and the lack of reports of pulmonary injury and cancer in workers suggest that the risk of lung cancer in TCE- exposed human populations is minimal. Genotoxicity TCE is a weak genotoxicant in a number of test systems (Bruning and Bolt 2000; Moore and Har- rington-Brock 2000; NRC 2006). Genotoxicity generally includes mutational end points, cytogeneticity, and primary DNA damage, whereas mutagenicity refers to the ability to induce heritable mutations. TCE oxidative metabolites trichloroacetic acid, dichloroacetic acid, and chloral hydrate generally have shown weak or no reactivity in mutagenicity tests; the weight of evidence in both in vitro and in vivo test sys-
Review of Toxicologic Studies 99 tems indicates that mutations are probably not key events in induction of cancer by these compounds (Moore and Harrington-Brock 2000). TCE was negative in a Salmonella typhimurium test strain that had cytochrome P-450 2E1 metabolizing capacity (Emmert et al. 2006). Neonatal B6C3F1 mice were given chloral hydrate, trichloroacetic acid, and TCE by intraperito- neal injection at the ages of 8 and 15 days; their livers were examined for 8-hydroxy-2â²-deoxyguanosine 24 and 48 h and 7 days after the final dose (Von Tungeln et al. 2002). Mice treated with trichloroacetic acid or chloral hydrate showed significantly higher DNA-8-hydroxy-2â²-deoxyguanosine adduct formation related to lipid peroxidation or oxidative stress; the authors concluded that male neonatal B6C3F1 mice are not sensitive to induction of liver cancer by these compounds. Significant increases in DNA migration in the Comet assay and micronuclei formation were re- ported in human HepG2 cells after treatment with TCE at 0.5-4 mM (Hu et al. 2008). Increases in both 8- hydroxy-2â²-deoxyguanosine-DNA adducts and thiobarbituric acid-reactive substances were observed; depletion of glutathione increased susceptibility to TCE-induced effects, whereas cotreatment with N- acetylcysteine prevented the effects. That indicated that oxidative stress probably played a role in TCE- induced genotoxic damage in those cells. Hypomethylated DNA was found in both dichloroacetic acid- promoted and trichloroacetic acid-promoted mouse hepatic tumors in an initiation-promotion experiment (Tao et al. 2004). Gene expression controlling cell growth, tissue remodeling, and xenobiotic metabolism was altered in in dichloroacetic acid-induced mouse hepatic tumors (Thai et al. 2003). Overall evidence indicates that TCE and the oxidative metabolites trichloroacetic acid, dichloroacetic acid, and chloral hy- drate are unlikely to act primarily by a mutational or genotoxic mechanism as hepatic carcinogens. The TCE glutathione conjugate DCVC has been shown to have genotoxic effects, including in- creased reverse mutations in S. typhimurium tester strains, unscheduled DNA synthesis, and formation of DNA adducts in vitro (Bruning and Bolt 2000; Moore and Harrington-Brock 2000). Genotoxicity meas- ures in rodent kidneys and primary cultures of human renal cells showed significant dose-dependent in- creases in results of the Comet assay (DNA single-strand breaks and alkali-labile sites) and in micronuclei frequency with subtoxic concentrations of TCE (Robbiano et al. 2004). Among the six rodent renal car- cinogens tested, TCE was among the ones that exhibited the lowest potency for these end points; nonethe- less, the results indicated that TCE is genotoxic in renal cells isolated from rats and humans. In another experiment, rats were exposed to TCE by inhalation or to DCVC by oral gavage. Proximal tubules iso- lated from kidneys of treated rats were assessed for DNA damage with the Comet assay (Clay 2008). Positive controls were included to demonstrate the sensitivity of the assay. Test results with TCE indi- cated a negative response in this assay. DCVC showed slight effects in a few animals 2 h after treatment and at the highest dose tested (10 mg/kg), but the effects were not strong enough to be considered posi- tive. On the basis of those findings and other published data, the authors concluded that renal tumors seen in bioassays are nongenotoxic in origin. Reproductive Effects Toxicity Studies in Males Several studies of the reproductive effects of TCE have been conducted, and many of these were reviewed by the National Research Council (NRC 2006). Zenick et al. (1984) found reduced copulatory behavior in male rats after an oral dose of 1,000 mg/kg per day 5 days/week for 6 weeks but indicated that the changes may have been related to the narcotic effects of TCE. Mice exposed to TCE by inhalation 4 h/day for 5 days (Land et al. 1981) showed an increased percentage of abnormal sperm at 2,000 ppm, the highest dose tested (about 3,000 mg/kg per day) and no increase at 200 ppm (about 300 mg/kg per day). Kumar et al. (2000a,b) exposed male Wistar rats by inhalation to 376 ppm for 12 or 24 weeks (4 h/day 5 days/week) and reported decreased epididymal sperm count and motility, reduced testosterone concentra-
100 Contaminated Water Supplies at Camp LejeuneâAssessing Potential Health Effects tions, and lower fertility when the treated rats were mated with untreated females. There were also sig- nificant reductions in body weight, testicular weight, total cauda epididymal sperm count, and percentage of motile sperm; the effects were greater after 24 weeks than after 12 weeks of exposure. By 24 weeks, the testes were atrophied and had smaller seminiferous tubules. Sertoli cells were present, but tubules contained no spermatocytes, and spermatids and Leydig cells were hypoplastic (Kumar et al. 2001). Xu et al. (2004) exposed male mice by inhalation to TCE at 1,000 ppm 6 h/day 5 days/week for 1-6 weeks and found no effects except for a significant reduction in the fertilizing ability of sperm from the TCE- exposed males when they were combined in vitro with eggs from superovulated control females or when the males were mated with superovulated control females. A study in male rabbits (Veeramachaneni et al. 2001) reported that a mixture of several agents, including TCE, caused alterations in mating desire and ability, sperm quality, and Leydig-cell function. The effects were assessed subjectively, and it is difficult to determine the contribution of TCE to the changes seen. Forkert et al. (2002) demonstrated that CYP2E1 is involved in the metabolism of TCE to chloral in Leydig cells and epididymides. Greater sensitivity of the mouse epididymis to high TCE vapor expo- sures correlated with greater chloral formation and higher concentrations of CYP2E1 in the epididymis than in the testis. Forkert et al. (2003) later found CYP2E1 in human epididymal epithelium and Leydig cells. Seminal-fluid samples from eight TCE-exposed mechanics who had diagnoses of clinical infertility contained TCE and some of its oxidative metabolites. More recently, Kan et al. (2007) evaluated epidi- dymal damage by TCE at the light-microscopic and electron-microscopic levels in mice after inhalation at 1,000 ppm for 1 day or for 1, 2, 3, or 4 weeks. The study showed epithelial sloughing and degeneration with separation of the seminal tubules from the basement membrane after exposure for 1 week or more. Epididymal damage became more severe with increasing duration of exposure. DuTeaux et al. (2003) found CYP2E1 and dichloroacetyl adducts in the epididymis and afferent ducts, which were indicative of the formation of reactive cytotoxic metabolites in the cells that were damaged. The absence of mitochon- drial Î²-lyase and the lack of formation of protein adducts in the epididymis and afferent ducts of rats dosed with DCVC suggest that TCE metabolites formed via the glutathione conjugation pathway do not participate in male reproductive toxicity. DuTeaux et al. (2004a,b) investigated the bioactivation of TCE and adduct formation in the testis and epididymis. In male rats ingesting TCE at estimated doses of 1.6- 2.0 and 3.4-3.7 mg/kg per day in drinking water for 14 days, there was a dose-dependent reduction in ca- pacity for in vitro fertilization of ova from untreated females. That effect occurred in the absence of any apparent alteration in the sperm other than a dose-dependent increase in oxidized proteins. The increase in lipid peroxidation implicates CYP2E1-mediated formation of reactive metabolites as a mechanism of tox- icity. Studies in Females Manson et al. (1984) exposed female rats orally by gavage to TCE at 10, 100, or 1,000 mg/kg per day for 2 weeks before mating, 1 week during mating, and throughout gestation. Although high concen- trations of TCE were measured in fat, adrenal glands, and ovaries, and uterine tissue contained high con- centrations of trichloroacetic acid, female fertility was not affected. However, 17% of females in the high- dose group died, and weight gain was significantly reduced. Neonatal survival was also significantly re- duced at the high dose, particularly in female offspring. Cosby and Dukelow (1992) conducted a study of oral exposure of pregnant mice to TCE at 24 or 240 mg/kg per day during gestation and in vitro fertilization studies with TCE, trichloroacetic acid, di- chloroacetic acid, and trichloroethanol. No effects were noted in the in vivo study; in the in vitro studies, there was a dose-related decrease in the percentage of fertilized embryos with trichloroacetic acid, di- chloroacetic acid, and trichloroethanol but not with TCE. Female rats were exposed to several male reproductive toxicants, including TCE, at 0.45% in drinking water for 2 weeks (Berger and Horner 2003). Oocytes collected after induced ovulation were incubated with sperm from unexposed males. The percentage of oocytes fertilized, the number of pene-
Review of Toxicologic Studies 101 trating sperm per oocyte, and the ability of oocytes to bind sperm plasma membrane proteins were all sig- nificantly reduced. Studies in Mating Pairs The NTP (1986b,c) conducted fertility-assessment-by-continuous-breeding studies of TCE die- tary exposure in mice and rats. The feed for both studies contained microencapsulated TCE at 0.15%, 0.30%, and 0.60%. In mice, the body weights of male F1 pups and the combined body weights of male and female F1 pups were significantly reduced in the 0.60% group. Sperm motility was reduced in the F0 parental males at the highest dose, but no other reproductive effects were seen. There were changes in testis and epididymis weight, increased liver weight, and increased combined kidney and adrenal weight. F0 females showed no reproductive effects but had increased liver weight. Treatment-related lesions were seen in the livers and kidneys of both males and females. Increased perinatal mortality was seen in the F1 pups at the highest dose level (NTP 1986b). In rats, there was a statistically significant trend toward re- duced numbers of litters per pair, and crossover mating was reduced if either of the parents was treated. General signs of toxicity included reduced body-weight gain, altered testis and epididymis weight, and increased relative liver weight and kidney and adrenal weight at all doses (NTP 1986c). Cancer The majority of chronic carcinogenicity bioassays of TCE in rodents have failed to reveal an in- creased incidence of testicular tumors. Maltoni et al. (1988a) did, however, report a dose-related increase in Leydig-cell tumors in male Sprague-Dawley rats exposed to TCE vapor at 100, 300, or 600 ppm for 104 weeks. The biologic significance of findings in that investigation has been discounted because of methodologic and statistical deficiencies (ATSDR 1997b). The NTP (1986a, 1988) reported the findings of a 2-year bioassay in which four strains of rats were gavaged with TCE at 0, 500, or 1,000 mg/kg 5 days/week. Only Marshall rats exhibited a dose-related increase in Leydig-cell tumors. Leydig-cell ade- noma is the most frequently encountered testicular tumor in mice and rats (Cook et al. 1999). The inci- dence varies from 1-5 % in control Sprague-Dawley rats to nearly 100% in F344 rats. Almost all those neoplasms are benign and occur in older rats. Most human testicular tumors are of germ-cell or Sertoli- cell origin and occur in young or middle-aged men. Leydig-cell tumors are rare in men (Cook et al. 1999). Thus, spontaneous or TCE-induced Leydig-cell tumors in rats are of questionable relevance to humans. In summary, the 2006 National Research Council report concluded that TCE is toxic to sper- matogenesis and the fertilizing ability of sperm. A detailed review by Lamb and Hentz (2006) concluded that male reproductive effects were generally seen at high concentrations that cause systemic toxicity and are more frequent in mice than in rats. The LOAEL for male reproductive effects after inhalation expo- sure is 376 ppm for 12 weeks (4 h/day 5 days/week) in rats, and there is general toxicity at that exposure. A NOAEL of 200 ppm for 5 days (4 h/day) was reported in mice in the Land et al. (1981) study, but no data for determining general toxicity were available. The LOAEL in rats for oral exposure is 1.6 mg/kg per day for 14 days in drinking water, but the relevance to humans of effects on in vitro fertilizing capac- ity is unclear. At 1,000 mg/kg, there were effects on copulatory behavior but with concomitant narcosis. No oral NOAEL was identified. The oral NOAEL for female fertility in mice was 240 mg/kg per day in in vitro fertilization stud- ies and in rats was 1,000 mg/kg per day with exposure before and during mating and during gestation. The LOAEL for impaired fertility in studies in which both males and females were exposed was about 145 mg/kg per day in rats and 875 mg/kg per day in mice. There was an indication of systemic toxicity at those doses. The NOAEL was about 70 mg/kg per day in rats and 405 mg/kg per day in mice. Additional studies of the reproductive toxicity of TCE are needed to permit better identification of LOAELs and NOAELs in both male and female rats and mice. In addition, more work on the mecha- nisms of action and potency of the various metabolites is needed.
102 Contaminated Water Supplies at Camp LejeuneâAssessing Potential Health Effects Developmental Effects Pregnancy Outcomes Several studies of TCE and metabolic products in rodents and avian species were reviewed in the 2006 National Research Council report. Early rodent studies using inhalation exposure (Schwetz et al. 1975; Dorfmueller et al. 1979) indicated little or no developmental toxicity as a result of exposure, whereas later studies by Dawson et al. (1990, 1993) and Johnson et al. (1998a,b, 2003) reported an in- crease in cardiovascular malformations at concentrations as low as 0.25 ppm. However, the latter studies used direct delivery of TCE to the gravid uterus or in drinking water and a novel examination process for examining the heart and great vessels. Fisher et al. (2001), using the same examination process as the Dawson and Johnson groups and in collaboration with them, reported no increase in cardiac or vascular defects. Warren et al. (2006) examined fetuses from the Fisher et al. (2001) study and found no ocular defects after TCE exposure. More recently, Carney et al. (2006), using a standard test protocol (inhalation exposure to TCE at 0, 50, 150, or 600 ppm for 6 h/day on gestation days 6-20), reported no effect of TCE on development in rats at up to 600 ppm, a concentration that produced minimal maternal toxicity. Collier et al. (2003) showed changes in gene expression during cardiac development after TCE exposure, and Klaunig et al. (1989) reported that TCE inhibited in vitro gap-junction-mediated intercellu- lar communication. Coberly et al. (1992) used the chimera assay and showed no effects of TCE in mouse preimplantation embryos. There is evidence from one laboratory that direct administration of the metabo- lites trichloroacetic acid and dichloroacetic acid to pregnant rats increased congenital cardiac defects in their offspring (Smith et al. 1989, 1992; Epstein et al. 1992); effects were observed at doses of 330 mg/kg per day and greater over multiple days and at single doses of 1,900-3,500 mg/kg. Several in vitro rodent and avian studies have shown effects of TCE on embryonic development, and these models have been used to investigate potential mechanisms for TCE and metabolite effects. For example, Saillenfait et al. (1995) reported concentration-dependent decreases in growth and differentia- tion indexes and increases in morphologic abnormalities in rat whole-embryo cultures, and Boyer et al. (2000), Hoffman et al. (2004), and Drake et al. (2006a,b) reported on TCE effects in a chick model. Changes in eye, pharyngeal arches, and cardiovascular development could be seen at high exposure con- centrations (such as 80-250 ppm). In most cases, the TCE metabolites trichloroacetic acid and di- chloroacetic acid were also studied and found to be at least as effective as TCE. Drake et al. (2006a,b) studied the effects of timing of TCE yolk-sac injection on chick heart development and found a greater effect if exposure occurred during endocardial cushion formation (Hamburger Hamilton [HH] stages 13- 20) than if exposure occurred at earlier stages of development (HH stages 3+-17). Those authors also re- ported hypercellularity and increased proliferation in the outflow tract and atrioventricular canal of the heart. However, Mishima et al. (2006), using chick whole-embryo organ culture and TCE at low concen- trations (10-80 ppm) in medium, reported reduction in mesenchymal cells in endocardial cushions. Ou et al. (2003), using an in vitro bovine organ culture, showed that TCE reduced heat-shock protein interac- tions with endothelial nitric oxide synthase, causing the synthase to shift to superoxide-anion generation, and inhibited vascular endothelial-cell proliferation stimulated by endothelial growth factor. Those effects on endothelial function are important in the development of cardiac defects. Although the in vitro studies are important in understanding the mechanism of TCE effects on development, their relevance for hazard characterization is unknown. The recent studies by Carney et al. (2006) address some of the recommendations of the 2006 Na- tional Research Council report that additional studies are needed to evaluate a LOAEL. The Carney study clearly shows no effects on heart or other organ development in the rat at exposure concentrations up to a minimal maternally toxic concentration. Several studies have been published to address mode of action but have not made clear which species is most appropriate for human modeling. Otherwise, the more re- cent data reviewed here do not change the conclusions of the 2006 National Research Council report on the prenatal toxicity of TCE. An in-depth review of the animal and human data on cardiovascular defects by Watson et al. (2006) concluded that there is no indication of a causal link between TCE and cardiovas-
Review of Toxicologic Studies 103 cular defects at environmentally relevant concentrations. On the basis of that review and the Carney et al. (2006) study results, the conclusion is appropriate. In summary, the database on the prenatal developmental effects of TCE is robust and indicates a lack of pregnancy outcomes up to concentrations that are minimally toxic in adults. The in vitro and whole-embryo studies are intriguing, but effects reported in them are probably due to the degree of expo- sure. On the basis of the Carney et al. (2006) study, the LOAEL of inhalation exposure during prenatal development in rats is greater than 600 ppm, and the NOAEL is also 600 ppm. The LOAEL for maternal or adult toxicity is 600 ppm, and the NOAEL is 150 ppm. Growth and Development A few studies have examined the neurologic effects of TCE after developmental exposure. For example, rat pups from dams exposed during gestation and lactation to TCE in drinking water at 312 mg/L (about 30 mg/kg per day) showed a reduction in 2-deoxyglucose uptake in the brain, indicating a reduction in glucose uptake or brain metabolism (Noland-Gerbec et al. 1986). Taylor et al (1985) showed an increase in activity of 60- and 90-day-old rats whose dams were exposed to TCE at 312 mg/L and above during gestation and lactation. In a followup study, Isaacson and Taylor (1989) reported that TCE at similar doses in rats reduced the amount of myelin in the dorsal hippocampus and proposed that the change might account for the behavioral effects of TCE. Another study by Isaacson et al. (1990) involved dosing young rats beginning at weaning with TCE in drinking water (312 mg/L) for 4 weeks, then with distilled water for 4 weeks. A second group was treated with TCE in drinking water for 4 weeks, distilled water for 2 weeks, then TCE for 2 weeks (as adults). Animals in the second group, but not the first group, showed reduced latency and improved learning in a Morris water maze. Both groups showed reduced hippocampal myelin. All those studies used small numbers of animals, and the dose was unclear, but they suggest neurologic effects of developmental exposure to TCE (see further discussion in the next section). A study by Peden-Adams et al. (2006) reported immunotoxicity after developmental exposure of mice to TCE at 0, 1,400, or 14,000 ppb in drinking water from gestational day 0 through the age of 3 weeks or 8 weeks. There was a decreased plaque-forming-cell response in males at both ages and doses and a decreased plaque-forming-cell response in females exposed to TCE at 1,400 ppb at the age of 8 weeks and at 14,000 ppb at both ages. Reduced numbers of splenic B220 cells were seen in 3-week-old pups exposed at 14,000 ppb. There was an increase in all thymic T-cell types (CD4+, CD8+, CD4+/CD8+, and CD4â/CD8â) at 8 weeks and increased delayed-type hypersensitivity in females at both concentrations and in males at only the high dose. This was the first study to report developmental immunologic effects at lower concentrations than in adults. The authors indicate the need to replicate and expand the examina- tion of critical windows for exposure (see section âImmunologic Effectsâ below). In summary, except for the studies described above, there are no studies on growth and develop- ment of animals after developmental exposure to TCE either prenatally or postnatally. The above studies indicate neurologic and immunologic effects of TCE exposure during development. However, they have limitations in design and interpretation. Further study of TCE is required to determine the types of effects, the lowest effect levels, and critical windows of development. Neurologic Effects TCE, like many other VOCs, inhibits functions of the CNS and possibly the peripheral nervous system. Acute effects in humans range from slight dizziness, fatigue, and headache to incoordination, an- esthesia, and death. TCE was commonly used for decades in vapor concentrations of about 2,000 ppm as a surgical, dental, and obstetrical anesthetic (Pembleton 1974). Such use was discontinued in the late 1970s. Chloral hydrate, an obligate intermediate of TCEâs oxidative metabolic pathway, remains one of the most widely used sedatives for dental, emergency medical, and imaging procedures for young chil-
104 Contaminated Water Supplies at Camp LejeuneâAssessing Potential Health Effects dren (Keengwe et al. 1999). The magnitude of CNS depression induced by chloral hydrate and TCE de- pends on the administered dose and on the target organ (brain) dose. CNS inhibitory effects diminish and disappear as TCE is metabolized and it and its metabolites are eliminated from the body. It should be rec- ognized that trichloroethanol, an end metabolite of the oxidation pathway, depresses the CNS. TCEâs nar- cotic effects are generally considered to be reversible. Irreversible (neurotoxic) effects, however, have been reported in human populations exposed for years to concentrations of TCE and other organic sol- vents high enough to produce clinically significant CNS symptoms (Evans and Balster 1991; ATSDR 1997b; Bruckner et al. 2008). There is concern that exposures to lower concentrations may also pose a risk of residual neurotoxic effects (EPA 2001; NRC 2006). TCE and other VOCs are intentionally inhaled for their euphoric and intoxicating effects. TCE and other solvents may be abused for years and result in malnutrition, cachexia, and residual damage of the brain and other organs; chronic neurologic and neuropsychologic sequelae have long been recognized. Rosenberg et al. (2002), for example, reported that a group of solvent abusers did significantly worse on tests of working memory and executive cognitive function than did alcoholics and cocaine addicts. A much higher percentage of solvent users had structural abnormalities in subcortical regions of the brain, as visualized by magnetic resonance imaging. They also exhibited moderate to severe diffuse abnormali- ties of cerebral white matter, a condition termed white-matter dementia. Inhalant abuse is the extreme form of TCE exposure, in that participants repeatedly subject themselves to vapor concentrations high enough to produce narcosis. Occupational exposures to TCE often involve inhalation of relatively high concentrations for years. Usual exposure concentrations are much lower than those experienced by solvent abusers but sub- stantially higher than encountered environmentally. Several studies of human subjects have been con- ducted to establish thresholds below which inhalation of TCE in the workplace will not impair motor or cognitive functions (ATSDR 1997b). Those studies have yielded surprisingly similar quantitative find- ings. Vernon and Ferguson (1969) exposed eight men to TCE at 0, 100, 300, and 1,000 ppm for 2 h. The highest concentration adversely affected performance on three of six standardized visual-motor tests; no significant decrements were found in response to the lower exposures. Stewart et al. (1970) measured a number of indexes of motor function in humans who inhaled TCE at 100 or 200 ppm for up to 7 h on 5 consecutive days. No decrements in performance were found, but some subjects described mild fatigue and sleepiness during their 4th and 5th days of inhaling TCE at 200 ppm. There were no significant dif- ferences in standardized achievement-test scores and self-reporting scales between controls and subjects who inhaled 100 ppm 6 h/day on 5 consecutive days (Triebig et al. 1977). Results of such studies served as the primary basis for the current occupational threshold limit value of 10 ppm and the short-term expo- sure limit of 25 ppm for TCE (ACGIH 2008). Those values were adopted in recognition that exercise en- hances VOCsâ systemic uptake and CNS effects. There have been a number of reports of different neurophysiologic and neuropsychologic effects of TCE in workers after short-term and long-term exposure (ATSDR 1997b; NRC 2006). Acute expo- sures to vapor at 500 ppm and higher result in dose-dependent signs of intoxication. Those effects are usually reversible, although there have been occasional cases of residual nerve dysfunction in persons overcome by a single high exposure (Feldman et al. 1985; Leandri et al. 1995). The patient described by Leandri et al. exhibited trigeminal nerve damage up to 4 months after exposure. Effects of repeated long- term exposure include memory loss, mood swings, impairment of cognitive function, and olfactory and trigeminal neuropathy. In most instances, TCE concentrations were not known, and many of the study subjects were exposed to solvent mixtures. A few investigations measured vapor concentrations in the workplace. Workers chronically exposed at 38-172 ppm described symptoms of dizziness, headache, nau- sea, and sleepiness, but trigeminal nerve dysfunction was not apparent (El Ghawabi et al. 1973). Albee et al. (2006) recently found no change in trigeminal nerve evoked potentials in rats inhaling TCE at up to 2,400 ppm over 13 weeks. Ruijten et al. (1991) found a change in one of two indexes of trigeminal nerve impairment in 31 printing workers exposed to TCE at 35-80 ppm for an average of 16 years. No impair- ment of motor or autonomic nerves was found.
Review of Toxicologic Studies 105 Feldman et al. (1992) measured prolonged latency in the blink reflex, which is indicative of trigeminal nerve impairment, in two metal degreasers heavily exposed to TCE for 7 and 16 years. Ruijten et al. (1991) found slight reductions in sensory nerve conduction velocity consistent with subclinical im- pairment of the peripheral nervous system. Rasmussen et al. (1993) found no disturbance of the trigemi- nal nerve but observed altered function of the olfactory nerve in 99 metal degreasers exposed to âhighâ concentrations of solvents (primarily TCE) for an average of 11 years; they also described dose- dependent increases in motor dyscoordination in the degreasers. A substantial number of neurotoxicity studies of TCE of acute and intermediate duration have been conducted in rats. CNS-depressant effects in the animals appear to be similar to those in humans and generally occur at higher exposure concentrations (ATSDR 1997b). That may be attributable in part to the availability of less sensitive measures of CNS depression in rodents. Bushnell and Oshiro (2000) found that inhalation of TCE at 2,000 or 2,400 ppm for 9 days reduced performance of rats on a sustained- attention task. Performance progressively improved (tolerance developed) during the protocol. Oshiro et al. (2004) then reported that inhalation of TCE at 1,600 or 2,400 ppm 6 h/day on 20 consecutive days did not impair later learning of a sustained-attention task. Inhalation at up to 1,500 ppm 16 h/day 5 days/week for 18 weeks increased latency in a visual-discrimination task but had no influence on spontaneous activ- ity, grip strength, coordinated movement, or peripheral-nerve conduction time (Kulig 1987). Latency in a visual-discrimination task improved progressively in the 500-ppm and 1,500-ppm groups. Auditory deficits in the midfrequency tone range have been observed in several strains of rats in response to inhalation of high concentrations of TCE (NRC 2006). Crofton and Zhao (1993), for example, described the onset of hearing loss after the fifth daily 6-h exposure at 4,000 ppm. It persisted for up to 14 weeks after exposure. The LOAEL in the study was 2,400 ppm. Histopathologic examination of rats that inhaled 4,000 ppm 6 h/day for 5 days revealed a loss of spiral ganglion cells in the middle turn of the cochlea and an inconsistent loss of hair cells (Fletcher et al. 1998). Recently, Albee et al. (2006) found focal loss of hair cells in the upper basal turn of the cochlea of rats that inhaled TCE at 2,500 ppm but not 800 ppm for 6 h/day 5 days/week for up to 13 weeks. Occupational exposures to such solvents as toluene and styrene have resulted in evidence of some hearing loss (Hodgkinson and Prasher 2006). That outcome has apparently not been assessed in groups exposed to TCE vapor at high concentrations. Kilburn (1999) reported an effect on the vestibulo-oculomotor system (balance) in a study of 150 jet-engine repairmen subjected to metal dusts and solvents, including TCE. There have been some accounts of neurologic effects in animals caused by relatively low doses of TCE. Changes in visual evoked potentials were described in rabbits exposed repeatedly to TCE at 350 ppm over 12 weeks (Blain et al. 1992). Reduced exploratory and social behavior was seen in rats after weeks of daily 6- to 7-h exposures to TCE vapor concentrations as low as 100 ppm. Silverman and Wil- liams (1975) did not use objective measurement techniques in their early study but merely observed the animals. Rats inhaling TCE at 50, 100, or 300 ppm for 8 h/day 5 days/week for 6 weeks exhibited altered sleep patterns; the effects were not dose-dependent (Arito et al. 1994). Decreased wakefulness during and after exposure was observed in the 50- and 100-ppm groups, respectively. The biologic or toxicologic significance of that effect is not apparent, but the Agency for Toxic Substances and Disease Registry (ATSDR) and the U.S. Environmental Protection Agency (EPA) each chose to use 50 ppm as the LOAEL with which to determine human exposure guidelines. ATSDR used an interspecies uncertainty factor of 3; EPA did not account for interspecies kinetic differences in its calculations. As described in Chapter 3, systemic uptake of inhaled VOCs is significantly greater in rodents than in humans. Physiologically based pharmacokinetic modeling has shown that higher blood concentrations are attained in rats during the ini- tial hours of an 8-h exposure to TCE at 10 and 100 ppm (Bruckner et al. 2004). Some investigations of potential cognitive effects of relatively low concentrations of TCE in ro- dents showed few adverse effects. Grandjean (1963) observed that inhalation of TCE at 800 ppm reduced swimming time in rats but produced no change in shuttle box or maze performance. Bushnell (1997) as- sessed the influence of a series of vapor concentrations on ratsâ response times to obtain a food reward; the NOAEL was 800 ppm. Albee et al. (1997) did not find alteration of flash-evoked potentials in rats inhaling 250 ppm. Waseem et al. (2001) stated that daily inhalation by rats of TCE at 376 ppm over 180
106 Contaminated Water Supplies at Camp LejeuneâAssessing Potential Health Effects days or consumption of TCE at 350, 700, or 1,400 ppm in water did not alter acquisition of a conditioned shock-avoidance response (cognition) but did enhance spontaneous motor activity. Similar findings in rats were described by Grandjean (1960) after acute 11- to 14-h inhalation exposures at 200 and 800 ppm. Those activity increases reflect the initial stimulant phase of action of CNS depressants. A few studies of TCE exposure in drinking water at about 30 mg/kg per day during pregnancy and lactation reported increased activity, reduced 2-deoxyglucose uptake in brain, and reduced hippocam- pal myelin (Taylor et al. 1985; Noland-Gerbec et al. 1986; Isaacson and Taylor 1989). An additional study (Isaacson et al. 1990) reported learning deficits and reduced hippocampal myelin in rats exposed as weanlings and adults. All those studies were from the same group, involved small numbers of animals, and require confirmation (see section âGrowth and Developmentâ above). Cancer Standard practice in 2-year bioassays is to perform gross and often microscopic pathologic inves- tigations of all organ systems in animals, including animals that die early. In general, an animal model is deemed relevant to establish the relative importance of the types of cancer, if any, that exposure to a given chemical at specific doses over a lifetime would be likely to elicit. In that context, animal TCE cancer bioassays cited previously did not show causality for brain cancer or other neurologic cancers. Immunologic Effects TCE has been reported to produce several forms of immunotoxicity, including the ability to act as a skin sensitizer, to exacerbate respiratory hypersensitivity (allergic asthma), to produce immunosuppres- sion, and to influence autoimmune diseases. Autoimmunity has been by far the most studied, and will be given the most attention here. Allergic Sensitization There have been many reports that workers exposed to TCE often show a severe irritating contact dermatitis manifested by a rash on the extremities, face, neck, or trunk with or without fever (Kamijima et al. 2007). It is sometimes referred to as severe generalized dermatitis, but it is unclear whether it has an immunologic etiology. Recently, a study conducted in workers at an electronic-element and metal-plating production plant in Guangdong Province, China, suggested an association with TCE-induced severe gen- eralized dermatitis and the HLA-B*1301 allele (Li et al. 2007). HLA alleles, known to be involved in governing immune recognition, are often reported to be associated with immune diseases. The evidence that TCE causes allergic contact dermatitis (skin allergy) comes primarily from a study by Tang et al. (2002) that used a modified guinea pig maximization test. TCE molecules themselves are too small to be antigenic and would need to bind covalently with skin proteins to elicit an immune response. There is no evidence that TCE can directly induce asthma, but data suggest that it can modulate asthma. Acute intraperitoneal administration of TCE to rats at 0.1mL/kg enhanced the production of sev- eral regulatory cytokines, including interleukin-4 (IL-4), and induced histamine release from basophils in animals previously immunized with a protein allergen (Seo et al. 2008). IL-4 and histamine are involved in the development of allergic asthma. The authors showed similar effects by treating cells in vitro with TCE from animals immunized with a protein allergen. Thus, unlike the Tang et al. (2002) study, which suggested that TCE directly causes allergic contact dermatitis, the studies by Seo et al. (2008) suggest that TCE may act as an adjuvant in enhancing allergic respiratory disease. Other studies have shown that VOCs may modulate immune cell types to favor induction of allergic responses in young children (Leh- mann et al. 2001). It is worth noting that a number of indoor and outdoor air pollutants are believed to
Review of Toxicologic Studies 107 exacerbate asthma, particularly in children (Selgrade et al. 2006). Further studies are needed to clarify those observations and determine whether TCE can induce or modulate allergic diseases. Immunosuppression That TCE can cause immunosuppression was first suggested on the basis of experimental-animal studies. Sanders et al. (1982) showed that mice exposed to TCE in drinking water for 4 or 6 months had deficiencies in their ability to mount normal immune responses. At a concentration as low as 100 mg/L (about 22 mg/kg per day), cell-mediated immunity and bone-marrow stem-cell colonization were inhib- ited. Wright et al. (1991) were able to confirm many of those findings in mice and rats treated with TCE by intraperitoneal injection. Peden-Adams et al. (2006) recently reported that mice exposed prenatally and postnatally to TCE are immunosuppressed at concentrations as low as 1.4 ppm in drinking water from gestation day 0 through the age of 3 or 8 weeks. Developmental immunotoxicity was manifested by sup- pression of antibody responses and decreases in B-cell numbers with a concomitant increase in delayed hypersensitivity responses and T-cell numbers. The shift in immune function would favor the develop- ment of infections from extracellular bacteria, such as streptococci and klebsiellae. The authors indicated that their data were preliminary and needed to be replicated (see section âGrowth and Developmentâ). Kaneko et al. (2000) reported that TCE suppressed immune functions in MRL-lpr/lpr mice after inhalation exposure to vapor at 1,000 or 2,000 ppm for 4 h/day for up to 8 weeks. The MRL-lpr/lpr mouse is genetically predisposed to develop systemic lupus erythematosus. Epidemiologic studies of TCE exposure and immunosuppression have been few. Byers et al. (1988) found increased concentrations of CD4+ and CD8+ T cells in a population with chronic domestic exposure to solvent-contaminated drinking water. Iavicoli et al. (2005) investigated the association be- tween serum concentrations of IL-2, IL-4, and interferon gamma (IFN-Î³) in workers exposed to TCE (mean urinary trichloroacetic acid concentration, 13.3 mg/g of creatinine). Serum concentrations of IL-2 and IFN-Î³ were increased, and that of IL-4 was reduced. Without additional immune tests, interpretation of variations in serum cytokines is currently not possible. Taken together, studies seem to be consistent in supporting the ability of TCE to suppress the immune system, at least in experimental animals. It should also be noted that the immunosuppressive effects seen in experimental animals generally occur at doses at which hepatic toxicity can be observed. Autoimmunity The MRL+/+ mouse model has been used historically to study TCE-induced autoimmunity. It is one of several mouse strains that have a mutation that results in the spontaneous development of systemic lupus erythematosus (SLE). The MLR+/+ strain was derived from the MRL-lpr/lpr mouse. The latter has a Fas mutation, a key protein responsible for cellular apoptosis, which influences the development of lu- pus early in life (50% mortality by the age of 6 months). The MLR+/+ mice lack the Fas mutation and develop the disease much slower (50% mortality within 17 months). Activated CD4+ T cells and regula- tory cytokines (such as IFN-Î³) play a key role in the development of SLE in MLR+/+ mice. Khan et al. (1995) showed that TCE accelerates the autoimmune disease process in MLR+/+ mice. Numerous studies have since examined the disease characteristics and mechanisms of action. Several mechanisms, not at all mutually exclusive, that have been proposed for TCE-induced autoimmunity are consistent with current understanding of the etiology of autoimmune disease. It has been suggested that TCE reactive metabolites, such as dichloroacetyl chloride (Khan et al. 1995, 2001) and lipid peroxidation-derived aldehydes, which form after TCE exposure (Wang et al. 2008), covalently bind to host proteins (Wang et al. 2007) and become immunogenic. Those protein adducts act as neoanti- gens and result in recognition by and activation of autoreactive T cells and autoantibody production. In further support of the hypothesis, Cai et al. (2007) were able to produce an immune response to adducts
108 Contaminated Water Supplies at Camp LejeuneâAssessing Potential Health Effects derived from TCE reactive metabolites after immunization in mice, and Wang et al. (2008) activated T cells in vitro after incubation with the protein adducts. Consistently with the formation of TCE metabo- lites that form protein adducts, Griffin et al. (2000a) prevented adduct formation and reversed, at least in part, the autoimmune effects in TCE-treated MRL+/+ mice by cotreatment with diallyl sulfide, an inhibi- tor of CYP2E1 that prevents TCE metabolism. TCE treatment of MRL+/+ mice also has been suggested to stimulate CD4+ T cells directly (Gil- bert et al. 1999). The activated CD4+ T cells develop a surface antigen, referred to as CD44 (Griffin et al. 2000b), that is involved in cell adhesion and is highly expressed in MRL-lpr/lpr mice. Treatment of MLR+/+ mice with trichloroacetaldehyde hydrate or trichloroacetic acid, major TCE metabolites, also activated CD4 T cells (Blossom et al. 2004). Consistently with the ability of CYP2E1 inhibition to re- verse the autoimmune effects (Griffin et al. 2000a), the activated cells are less susceptible to a form of cellular apoptosis, referred to as activation-induced cell death, that is observed in many autoimmune dis- eases. TCE-mediated defects of activation-induced cell death were recently found to be associated with metalloproteinase 7, which later facilitated FasL, a receptor involved in apoptosis (Blossom and Gilbert 2006). The T cells activated by the protein adducts are believed to represent predominantly a Th1 pheno- type, rather than Th2, inasmuch as they produced higher concentrations of IFN-Î³, a Th1 cytokine, and lower concentrations of IL-4, a Th2 cytokine. Th1 cytokines are usually associated with systemic auto- immune diseases. Gilbert et al. (1999, 2004) also provided evidence that trichloroacetaldehyde hydrate can activate T cells through the formation of a Schiff base. Schiff-base-forming structures, such as alde- hydes and ketones, can substitute for physiologic donors of carbonyl groups and directly activate CD4 cells without engaging the T-cell receptor (Rhodes et al. 1995). The chronic effects of TCE exposure in MLR+/+ mice have been addressed in several studies. Griffin et al. (2000c) exposed MLR+/+ mice to TCE at 0.1, 0.5, or 2.5 mg/mL in drinking water (21, 100, or 400 mg/kg) for 4 or 32 weeks and showed CD4+ T-cell activation and induction of autoimmune hepati- tis at all doses. Cai et al. (2008) exposed mice to TCE at 0.5 mg/mL in drinking water for up to 48 weeks. In addition to increases antinuclear autoantibody titers, lymphocyte infiltration and immunoglobulin de- posits were found in the liver, pancreas, lungs, and kidneys (including glomeruli); this was consistent with SLE or an SLE-like disease. Blossom et al. (2007) treated MLR+/+ mice with trichloroacetaldehyde hydrate at 0.1, 0.3, or 0.9 mg/mL (about 13, 49, or 143 mg/kg per day) in drinking water for 40 weeks. Long-term exposure promoted alopecia and skin inflammation. The lesions did not appear similar to the cutaneous lupus seen in older MLR mice or the skin conditions in patients with systemic scleroderma; rather, they may have been associated with dermal infiltration of activated T cells. Taken together, the experimental studies suggest two mechanisms, not mutually exclusive, by which TCE modulates autoimmune disease. The first involves TCE reactive metabolites that covalently bind to host protein to produce neoantigens that stimulate the formation of autoreactive immune cells. The second involves activation of Th1 cells nonspecifically by TCE metabolites, which also leads eventually to the formation of autoreactive immune cells. Both processes, like autoimmune diseases in general, in- volve cellular apoptosis. The latter is a general mechanism that may be relevant to a variety of autoim- mune diseases, whereas the former may be more specific to particular diseases (such as lupus). PERCHLOROETHYLENE Data on the toxicity of PCE were summarized in a 1985 health assessment by EPA (1985) and an addendum issued in 1986 (EPA 1986). The California Environmental Protection Agency published a pub- lic-health goal for PCE in drinking water (CalEPA 2001) that included a brief review of toxicity data. ATSDR (1997c) also published a toxicologic profile of PCE, and a draft neurotoxicity assessment was available from EPA (2003). Literature reviews were available in particular subject areas (e.g., Beliles 2002; Klaunig et al. 2003; Wernke and Schell 2004). Such references were relied on for defining the body of literature available on PCE; in addition, a literature search was done to determine whether any relevant new publications were available. Conclusions drawn for the present report were based on a review of the
Review of Toxicologic Studies 109 body of available literature. The data are presented below by organ system, and toxic effects are consid- ered separately from carcinogenic effects. Hepatic Effects Toxicity PCE, like TCE, has a limited ability to cause acute, subacute, or chronic hepatic injury in rodents. Klaassen and Plaa (1966) assessed the acute cytotoxicity of PCE, TCE, and several other halocarbons in male Swiss-Webster mice given each chemical in a single intraperitoneal injection. PCE was a slightly less potent hepatotoxicant than TCE. A lethal dose of PCE was required to produce a substantail increase in serum alanine aminotransferase activity. Recently, Philip et al. (2007) reported that male Swiss- Webster mice given PCE at 150 mg/kg by aqueous gavage exhibited a transient increase in serum alanine aminotransferase activity. Higher alanine aminotransferase concentrations were manifested at 500 and 1,000 mg/kg. The extent of injury regressed substantially over a 30-day dosing period, apparently because of the onset of tissue repair and PCEâs inhibition of its own oxidative metabolism. Buben and OâFlaherty (1985) saw modest increases over controls in serum alanine aminotransferase, liver weight, and hepatic triglycerides in male Swiss-Cox mice dosed with PCE at 500-2,000 mg/kg per day for 6 weeks by corn- oil gavage; the lack of dose dependence reflected saturation of metabolic activation in this dosage range. Hayes et al. (1986) found no consistent dose-related effects on any hematologic or clinical-chemistry measure in male or female rats that ingested PCE at about 14, 400, or 1,440 mg/kg per day for 90 days. Rats may be less susceptible than mice, although the absence of hepatotoxicity in rats in this instance can also be attributed to differences in oral-exposure regimens. Ingestion of a bolus dose of PCE will result in a high tissue dose that exceeds the capacity of the liverâs defense and repair mechanisms. Consumption of the total dose in relatively small, divided doses might not exceed such a cytotoxicity threshold. PCE-induced hepatic injury is believed to be a consequence of oxidative metabolism of PCE (Lash and Parker 2001). The PCE oxidative pathway is described in Chapter 3 (see section on metabolic activation and inactivation of TCE and PCE). PCE is more poorly metabolized by cytochrome P-450s than TCE, but two additional intermediate metabolites of PCE also contribute to its hepatocytotoxicity: the initial oxidation product, PCE oxide (epoxide), and one of its convertants, trichloroacetyl chloride. The latter is transformed to trichloroacetic acid, the major metabolite of PCE. Some trichloroacetic acid can be dechlorinated to form dichloroacetic acid. Trichloroacetic acid and dichloroacetic acid are also products of TCE biotransformation. As described earlier, trichloroacetic acid is primarily responsible for activation of the nuclear receptor PPARÎ±, which stimulates peroxisomal enzymes and selected cyto- chrome P-450s involved in lipid metabolism. That results in peroxisome proliferation, which generates reactive oxygen moieties that can cause lipid peroxidation, cellular injury, and altered expression of cell- signaling proteins (Bull 2000). Lash et al. (2007) recently demonstrated that cytochrome P-450 inhibition resulted in reduced injury of hepatocytes isolated from male F344 rats and exposed to PCE. Glutathione depletion increased cellular injury, apparently because of a shift from glutathione conjugation to the oxi- dative metabolism of PCE. Humans should be less susceptible to hepatic injury by PCE than rodents because of lower inter- nal and target-organ doses of the parent compound and its bioactive metabolites. As described in Chapter 3, rats achieve a substantially higher internal dose of PCE than humans on inhaling it. Volkel et al. (1998) subjected rats and people to identical PCE inhalation regimens. Blood trichloroacetic acid concentrations were 3- to 10-times higher in the rats. Dichloroacetic acid was not detectable in human urine, but substan- tial amounts were found in rat urine. A study of the urinary excretion of total trichloro-metabolites by PCE-exposed workers led Ohtsuki et al. (1983) to conclude that the capacity of men to metabolize PCE was rather rather low. Lash and Parker (2001) noted that saturation of PCE metabolism occurred at lower doses in humans than in rodents. That implies that humans have lower capacity to form biologically ac- tive metabolites from moderate to high PCE doses. The difference is reflected in the finding of much
110 Contaminated Water Supplies at Camp LejeuneâAssessing Potential Health Effects lower concentrations of protein adducts in the blood of humans than in the blood of rats subjected to equivalent PCE inhalation exposures (Pahler et al. 1999). Stewart et al. (1977) found no evidence of hepa- totoxicity in six male and six female volunteers exposed randomly to PCE at 0, 25, or 100 ppm 5.5 h/day 5 days/week for 11 weeks. Serum alanine aminotransferase activity was not increased in 22 dry cleaners examined in Belgium (Lauwerys et al. 1983). A research group in Italy studied 141 employees exposed to PCE in small laundries and dry-cleaning shops (Gennari et al. 1992); no worker exhibited clinical signs of hepatic dysfunction or abnormal serum enzyme concentrations, although there did appear to be an in- crease in one isozyme of Î³-glutamyltransferase, which was said to be associated with hepatobiliary im- pairment. Another investigation of dry cleaners failed to reveal increases in serum enzyme concentrations but did show mild to moderate changes in hepatic parenchyma revealed by ultrasonography (Brodkin et al. 1995). Considerable experience in occupational settings demonstrates that humans, like rodents, may develop mild but reversible hepatic injury on exposure to high concentrations (ATSDR 1997b). Cancer Exposure to PCE by inhalation (NTP 1986a) and by oral gavage (NCI 1977) has shown increases in liver cancer in B6C3F1 mice (Table 4-1). Inhalation exposure of 50 B6C3F1 mice of each sex at 0, 100, and 200 ppm 6 h/day 5 days/week for 103 weeks caused increased incidence of hepatocellular neoplasms (adenomas and carcinomas combined) in males and females. The incidence in males was 17 of 49, 31 of 49, and 41 of 50, respectively; in females, it was 4 of 48, 17 of 50, and 38 of 50, respectively. As also shown in Table 4-1, exposure of male B6C3F1 mice to PCE at 536 and 1,072 mg/kg per day and of fe- male mice at 386 and 722 mg/kg per day in corn oil with epichlorohydrin stabilizer by oral gavage yielded significant increases in hepatocellular carcinomas (P < 0.001). Thus, there is clear evidence of hepatic carcinogenicity in B6C3F1 mice related to PCE exposure. No hepatic-cancer effects were seen in F344/N rats exposed by inhalation to PCE at 200 and 400 ppm for 103 weeks (NTP 1986a). Trichloroacetic acid is also a metabolite of PCE. As discussed in detail in the preceding section on TCE cancer bioassays, trichloroacetic acid induces peroxisome proliferation in B6C3F1 mouse liver but not in rat liver. That difference should be taken into account, as discussed in greater detail in the pre- ceding section, in considering the relevance of mouse hepatocellular tumors for humans. As shown in Table 4-2, gavage studies to determine carcinogenicity in Osborne Mendel rats (NCI 1977) were judged inadequate because of early mortality when PCE-induced toxic nephropathy reduced survival of dosed rats. There were many early deaths, so those results precluded conclusions regarding carcinogenicity of PCE in the rats. Renal Effects Toxicity PCE is somewhat more nephrotoxic in mice and rats than TCE, but high, subchronic oral bolus dosing with PCE is required to affect the kidneys adversely. Jonker et al. (1996), for example, gave fe- male Wistar rats TCE at 500 or 600 mg/kg per day by corn-oil gavage for 32 consecutive days. PCE elic- ited doubling of urinary protein and activities of several enzymes released from injured renal proximal tubule cells. TCE produced slight increases in just two of the enzymes. Coadministration of TCE and PCE resulted in additive nephrotoxicity. Philip et al. (2007) recently failed to see morphologic changes in the kidneys of male Swiss-Webster mice given PCE at 150, 500, or 1,500 mg/kg per day by aqueous ga- vage for 30 days. Green et al. (1990) gavaged male F344 rats with PCE at 1,500 mg/kg per day in corn oil for 42 days. There were increases in urine volume and urinary enzyme activities that were indicative of
TABLE 4-1 Animal Cancer Studies of PCE with Positive Outcomes Timing and Species Strain Dose or Concentration Route Duration Outcomes LOAEL References Mouse B6C3F1 0, 100, 200 ppm Inhalation 6 h/day Hepatocellular adenoma in 100 ppm NTP 1986a 5 days/week, males; hepatocellular 103 weeks carcinoma in males, females Mouse B6C3F1 Males: 0, 536, 1,072 Oral gavage 5 days/week, Hepatocellular 536 mg/kg per day (males); NCI 1977 mg/kg per day; females: 0, (corn oil) 78 weeks, carcinoma in males, females 386 mg/kg per day 386, 722 mg/kg per day; observed to 90 (females) epichlorohydrin stabilizer weeks Rat F344/N 0, 200, 400 ppm Inhalation 6 h/day 5 Stage 3 mononuclear 200 ppm (males, females) NTP 1986a; Mennear et al. days/week, cell leukemia;a rare renal 1986 103 weeks tubular adenoma, adenocarcinoma in males a Mononuclear-cell leukemia is common in aging F344 rats. TABLE 4-2 Animal Cancer Studies of PCE Determined to be Negative, Inadequate, or Incomplete Species Strain Dose or Concentration Route Timing and Duration Outcomes Comment NOAEL Reference Rat Osborne Males, 471, 941 mg/kg Oral gavage 78 weeks Early mortality due to Inadequate â NCI 1977 Mendel per day; females, (corn oil) PCE-induced toxic study 474, 949 mg/kg per day nephropathy Rat Sprague- 0, 300, 600 ppm, Inhalation 52 weeks, then held Hematologic examinations Short duration 600 ppm Rampy et al. Dawley 6 h/day, 5 days/week another 12 mo and tumor outcomes of exposure; 1978 negative unpublished Rat Sprague- 0, 500 mg/kg per day Oral gavage 4-5 days/week, No increase in total and Negative study 500 mg/kg Maltoni et al. Dawley (olive oil) 104 weeks malignant tumors at 141 per day 1986 weeks Rat Long-Evans â Oral gavage In-life studies completed; Study judged Inadequate â NTP 1986a Sherman overall findings âinadequate,â no final study Wistar unpublished report available F344 Mouse Ha:ICR 1st application, 163.0 Skin 1st application, 1st application, 4/7 mice Negative study; â Van Duuren et Swiss mg; 2nd application, 229 days to papilloma; with papillomas/total P > 0.05 al. 1979 54.0 mg 2nd application: papillomas; 0 papillomas 2nd application, 0 111
112 Contaminated Water Supplies at Camp LejeuneâAssessing Potential Health Effects mild renal proximal tubular-cell damage. Histopathologic examination revealed the presence of hyaline droplet accumulation and some regeneration in the animalsâ proximal tubules. Ingestion of daily doses of PCE estimated at 14, 400, or 1,400 mg/kg in drinking water for 90 days failed to produce renal damage in male or female Sprague-Dawley-derived CD rats (Hayes et al. 1986). Thus, ingestion of divided doses of PCE in water over the course of the day is much less nephrotoxic in rodents than ingestion of the total dose once a day. Subchronic and chronic inhalation of PCE has resulted in limited evidence of nephrotoxicity in rodents. Exposure of both sexes of F344 rats and B6C3F1 mice to PCE at 400 ppm 6 h/day for 28 days failed to increase renal weights or produce histopathologic changes. Tinston (1995) reported mild, pro- gressive glomerulonephropathy and increased pleomorphism of proximal tubular nuclei in male but not female rats that inhaled PCE at 1,000 ppm for up to 19 weeks. Nephropathy was seen in rats and mice chronically given high oral bolus doses of PCE in corn oil (NCI 1977). Karyomegaly occurred in renal tubules of male and female B6C3F1 mice exposed to PCE at 200-1,600 ppm by inhalation for 13 weeks (NTP 1986a); the NOAEL in mice was 100 ppm. Renal lesions were not seen in F344 rats exposed to PCE at 1,600 ppm. Dose-dependent karyomegaly was observed in each sex of rats exposed to PCE at 200 or 400 ppm and mice exposed at 100 and 200 ppm chronically (NTP 1986a); there were also low inci- dences of renal proximal tubular-cell hyperplasia in the male rats. The metabolism and mode of nephrotoxicity of PCE and TCE appear to be quite similar, although PCE and its metabolites are somewhat more potent. Renal effects of both halocarbons are due primarily to metabolites formed via the glutathione conjugation pathway (Lash and Parker 2001). The sites, enzymes, and products associated with PCE biotransformation are almost identical with those associated with TCE. (The TCE and PCE glutathione conjugation pathways were described earlier in Chapter 3.) The primary difference is that S-(1,2,2-trichlorovinyl)glutathione (TCVG) and S-(1,2,2-trichlorovinyl)-L-cysteine (TCVC) are produced from PCE, and DCVG and DCVC from TCE. TCVC can be detoxified by acetyla- tion or cleaved by renal cytosolic and mitochondrial Î²-lyases to trichlorothioketene, which loses a chlo- ride ion to form dichlorothioketene. The latter is a very reactive moiety that binds to cellular proteins and DNA. TCVC, like DCVC, can be enzymatically oxidized to form the very reactive S-(1,2,2- trichlorovinyl)-L-cysteine sulfoxide (TCVCS) (Krause et al. 2003). TCVCS was shown to be more nephorotoxic than TCVC in male Sprague-Dawley rats on intraperitoneal injection (Elfarra and Krause 2007). TCVC caused more pronounced necrosis of renal proximal tubular cells in male Wistar rats than did DCVC after intravenous injection (Birner et al. 1997). Lash et al. (2002) similarly found that PCE and TCVG were more toxic than TCE and DCVG to renal cortical cells from F344 rats in vitro. Cells from male rats were more sensitive than cells from females to PCE-induced and TCVG-induced mitochondrial state 3 respiratory inhibition and cytotoxicity. Isolated rat hepatocytes and their mitochondria were unaf- fected by PCE and TCVC. Increased glutathione concentrations increased TCE-induced and PCE-induced cytotoxicity in suspensions of rat renal cortical cells but not hepatocytes (Lash et al. 2007). In summary, PCEâs glutathione-pathway metabolites are more reactive and cytotoxic in the kidney than are TCEâs glu- tathione metabolites. PCE cytotoxicity is both sex-dependent and tissue-dependent. Occupational exposures to PCE vapor have led to several reports of mild renal tubular damage (ATSDR 1997b). Employees of dry-cleaning shops have been the subjects of a number of investigations. Increased concentrations of urinary lysozyme or increased Î²-glucuronidase activity was described in dry cleaners exposed to PCE at average concentrations of 10 ppm (Franchini et al. 1983) and 23 ppm (Vysko- cil et al. 1990) for 9-14 years. In a more comprehensive study of renal function, a number of urinary in- dexes indicative of early glomerular and tubular changes were increased over controls in 50 dry cleaners who inhaled PCE at an average concentration of 15 ppm for 10 years (Mutti et al. 1992). There was a lack of association between the extent of the changes and the intensity and duration of exposure. Verplanke et al. (1999) monitored several indexes of tubular and glomerular function in Dutch dry-cleaning workers but found an increase only in retinol-binding protein in their urine. Other groups of investigators have failed to find evidence of renal effects in such populations. A laboratory study of 10 male and 10 female adults who inhaled PCE at up to 150 ppm for as long as 7.5 h/day for 5 days did not show changes from pre-exposure baseline urinary and blood urea nitrogen concentrations (Stewart et al. 1981). Hake and
Review of Toxicologic Studies 113 Stewart (1977) described a dry cleaner who was found unconscious in a pool of PCE, where he had been for an estimated 12 h. Laboratory tests revealed hematuria and proteinuria that lasted for 10 and 20 days, respectively. Mild hepatic damage was revealed by transient increases in serum enzymes. On the basis of the foregoing human experiences, PCE has limited ability to cause diffuse changes along the nephron, although extremely high exposures can lead to pronounced changes. Cancer No renal carcinomas were observed in B6C3F1 mice exposed to PCE at 0, 100, and 200 ppm for 103 weeks (NTP 1986a); dose-related karyomegaly was found in both males and females, but it was not accompanied by tubular-cell hyperplasia as it was in rats (Table 4-1). F344/N rats develop nephrologic changes as a normal condition of ageing. Both sexes showed re- nal tubular-cell karyomagaly and males renal tubular-cell hyperplasia after exposure to PCE at 200 and 400 ppm for 103 weeks (NTP 1986a). That effect has been seen in other strains of rats exposed to chlo- rinated ethylenes, so it is not necessarily specific to PCE. Renal tubular-cell adenomas and adenocarci- nomas were detected in male, but not female, rats. The incidence of renal neoplasm in the males was 1 of 49 controls, 3 of 49 exposed at 200 ppm, and 4 of 49 exposed at 400 ppm. Even though the results were not statistically significant, it was noted that those particular tumors are rare in F344/N male rats, so they were believed to have been caused by PCE exposure. Pulmonary Effects Toxicity Little information was available on the pulmonary toxicity of PCE in laboratory animals or hu- mans. Epithelial degeneration was observed in mice that inhaled PCE at 300 ppm 6 h/day for 5 days (Aoki et al. 1994). That effect was more severe in the olfactory than in the respiratory mucosa. Mice ex- posed to PCE at 50 ppm for 3 h were more susceptible to two strains of inhaled bacteria than controls (Aranyi et al. 1986). It was hypothesized that the susceptibility occurred because PCE inhibited alveolar macrophage activity. Intermittent inhalation of of PCE at 1,600 ppm for 13 weeks produced congestion in the lungs of rats (NTP 1986a). The 800-ppm vapor concentration did not have that effect. Pulmonary congestion was seen in mice that inhaled PCE at 100 ppm or greater in the 103-week phase of the cancer bioassay. There was not an increased incidence of lung tumors in the mice or rats. The reason for the ap- parent lack of significant pulmonary toxicity or carcinogenicity in rodents may have been the small amounts of the cytochrome P-450 isozymes that metabolically activate PCE. Although CYP2E1 is abun- dant in mouse lung, it does not appear to be active in PCE metabolism in rat (Hanioka et al. 1995) or hu- man (White et al. 2001) cells, thereby inferring that CYP2E1 is unlikely to be a factor in metabolizing PCE in the rat or human lung. A number of studies of inhaled PCE have shown that vapor concentrations as low as about 200-300 ppm can cause mild irritation of the nasal passages of humans (ATSDR 1997b). Stewart et al. (1981) subjected four male volunteers to PCE at 0, 20, 100, and 150 ppm 7.5 h/day for 5 days. The subjects were exposed sequentially to each concentration for 1 week. Pulmonary-function measurements did not reveal any decrements. Pulmonary edema has been described in a person rendered unconscious by PCE fumes (Patel et al. 1973). Cancer No increases in lung proliferative lesions were seen in B6C3F1 mice of either sex after inhalation of PCE at 100 or 200 ppm for 103 weeks, nor were lung neoplasms seen in male or female F344/N rats exposed at 200 or 400 ppm for 103 weeks (NTP 1986a).
114 Contaminated Water Supplies at Camp LejeuneâAssessing Potential Health Effects Genotoxicity The genetic toxicity of PCE has been reviewed extensively by the California Environmental Pro- tection Agency (CalEPA 1992), the International Agency for Research on Cancer (IARC 1995), and ATSDR (1997c). In general, studies have not yielded evidence of genotoxicity of PCE. Results in pro- karyotic mutation assays (principally with Salmonella typhimurium and Escherichia coli) have been nega- tive with and without S-9 rat liver microsomal metabolic activation. More recently, PCE was negative in an S. typhimurium tester strain competent for CYP2E1 metabolizing capacity (Emmert et al. 2006). Me- tabolites of PCE have been shown to be mutagenic in vitro. The minor PCE urinary metabolite glu- tathione conjugate TCVG is mutagenic in S. typhimurium TA 100 with renal cytosol metabolic activation (Vamvakas et al. 1987). TCVG, the precursor of the cysteine conjugate, was also mutagenic to S. typhi- murium TA 100 with rat kidney microsomal metabolic activation (Vamvakas et al. 1989). Reproductive Effects Toxicity Only two studies have addressed the potential for reproductive toxicity of PCE: a study by Beliles et al. (1980) and a two-generation study by Tinston (1995). In the Beliles et al. (1980) study, male rats and mice were exposed to PCE by inhalation at 100 and 500 ppm 7 h/day for 5 days. No effects on sperm structure were seen in rats, but in the 500-ppm group of mice, there was a significant increase in the inci- dence of abnormal sperm heads 4 weeks after exposure. The timing of the appearance of the effects after exposure suggests that spermatocyte or spermatogonia were most sensitive to exposure to PCE. The NOAEL was 100 ppm. The two-generation study by Tinston (1995) involved exposure of male and female rats (Alpk:ApfSD) to PCE at 0, 100, 300, or 1,000 ppm 6 h/day 5 days/week for 11 weeks before mating and then daily during mating and through gestation to day 20. There was no exposure from gestation day 21 through postnatal day 6, and then exposure resumed. F1 parents were selected on postnatal day 29, and exposure continued for at least 11 weeks before mating and then through mating, gestation, and lactation until the F2 litters were weaned. Parental animals experienced CNS depression, decreased respiration at 300 and 1,000 ppm, decreased body weight at all concentrations during lactation, and nephrotoxicity at 1,000 ppm. Later growth in the 100-ppm and 300-ppm groups was similar to that in controls. There were reductions in live births, litter size, postnatal survival, and pup weight at 1,000 ppm. Pup kidney, liver, and testis weights were reduced at 300 and 1,000 ppm but not when adjusted for body weight. The NOAEL was considered to be 100 ppm. Cancer PCE has not been shown to cause testicular tumors in mice or rats in chronic carcinogenicity bio- assays. The potential oncogenicity of PCE was evaluated in male and female F344 rats that inhaled PCE at 0, 200, or 400 ppm 6 h/day 5 days/week for 2 years (NTP 1986a). The overall incidence of Leydig cell tumors was 70%, 80%, and 82% in the 0-, 200-, and 400-ppm groups, respectively. Haseman et al. (1998) reported that NTP control F344 rats have an extremely high spontaneous incidence (89.1%) of Leydig cell tumors. F344 rats have therefore been replaced in the NTP bioassay program with Wistar Han rats. As discussed in the foregoing PCE reproductive-cancer section, Leydig cell tumors in F344 rats are believed to be irrelevant to humans. In summary, the effects of PCE on sperm morphology and germ cells in rats and mice suggest an effect on male reproduction (Beliles et al. 1980; Tinston 1995), but more detailed studies are needed to clarify the effects and the relationship to magnitude of exposure. On the basis of the available studies, the
Review of Toxicologic Studies 115 LOAEL was 300 ppm for exposure 6 h/day 5 days/week for 11 weeks before mating and then daily dur- ing mating and through gestation to day 20. The NOAEL was 100 ppm. Leydig cell tumors reported in the chronic study (NTP 1986a) were discounted because of the high background rates of such tumors in F344 rats. Developmental Effects Pregnancy Outcomes Several studies in rodents have focused on the potential for developmental toxicity of PCE. Schwetz et al. (1975) exposed pregnant mice and rats to PCE at 300 ppm on gestation days 6-15 and found maternal and developmental toxicity, including lowered weight of mice, subcutaneous edema in mouse fetuses, and increased resorption in rats. Beliles et al. (1980) found only minor changes in devel- opment in rats exposed to PCE at 300 ppm 7 h/day 5 days/week, either before mating and throughout ges- tation or only during gestation. In rabbits exposed at 500 ppm before or during gestation, there were no significant maternal or developmental effects. Tepe et al. (1982) studied the effects of PCE exposure at 1,000 ppm in Long-Evans female rats before mating and during pregnancy or only during pregnancy to determine the more sensitive window. Increased relative maternal hepatic weight and reduced fetal body weight were seen after exposure to PCE at 1,000 ppm by inhalation during pregnancy. An increase in skeletal variations was seen in the group exposed before mating and during pregnancy, and soft-tissue variations (such as renal dysplasia) were seen more in the group exposed only during pregnancy. Narotsky and Kavlock (1995) evaluated the effects of PCE at 0, 900, or 1,200 mg/kg per day ad- ministered orally by intubation on gestation days 6-19. There were no live pups in the 1,200-mg/kg group; maternal ataxia and reduced weight, fewer pups per litter, full litter resorptions, and microphthal- mia or anophthalmia were seen at 900 mg/kg. Because of the high doses used and incomplete anatomic evaluation of pups, this study has little utility in hazard characterization. More recently, Carney et al. (2006), using a standard prenatal developmental-toxicity study pro- tocol (inhalation exposure 6 h/day 7 days/week on gestation days 6-20), reported reduced uterine and pla- cental weights, reduced body weight, and reduced ossification in the thoracic vertebral centra in rats at PCE concentrations of 250 and 600 ppm and maternal toxicity at 600 ppm. The LOAEL for reduced fetal body weight was 250 ppm in this study. Reduced fetal body weight in the rat can be considered analogous to âsmall for gestational ageâ in humans. An in vitro study by Saillenfait et al. (1995) reported concentration-dependent decreases in growth and differentiation indexes and increases in morphologic abnormalities in rat whole-embryo cul- ture (gestation day 10) in a medium containing PCE at 3.5 mM. However, the relevance of the data to human risk assessment is questionable. In summary, data from recent studies do not substantially alter the conclusions of EPA (1985), which were that data âdo not indicate any significant teratogenic potential of PCEâ and that other ob- served effects reflect primarily delayed development. The 2006 study by Carney et al. confirms the lack of teratogenicity of PCE, and the developmental effects reported at the lowest concentrations were rela- tively minor. The LOAEL for maternal effects was 600 ppm and for developmental effects was 250 ppm. The NOAEL for maternal effects was 250 ppm and for developmental effects was 65 ppm. Growth and Development Concerns about the neurotoxicity of PCE prompted investigations of the potential effects of expo- sure during development (Nelson et al. 1980; Manson et al. 1982; Fredriksson et al. 1993; Chen et al. 2002). Nelson et al. (1980) evaluated the effects of inhalation exposure to PCE at 900 ppm 7 h/day on gestation days 7-13 or 14-20. Dams gained less weight and had lower food consumption than controls
116 Contaminated Water Supplies at Camp LejeuneâAssessing Potential Health Effects during exposure. Animals were allowed to litter, and pups showed signs of neurobehavioral impairment on certain days of testing. Pups exposed on gestation days 14-20 initially performed more poorly than controls but later were superior on other tests. Significant reductions in acetylcholine were seen in both exposure groups, and reductions in dopamine were seen in the group exposed on gestation days 7-13. An- other group of rats exposed to PCE at 100 ppm showed no differences from controls in any of the behav- ioral tests. Manson et al. (1982) did a followup study on the animals from the Tepe et al. (1982) study to evaluate the potential for postnatal body-weight and skeletal or soft-tissue variants, carcinogenicity, and neurotoxicity. No effects on any of those characteristics were observed. Fredriksson et al. (1993) studied mice exposed orally to PCE (5 or 320 mg/kg per day) on postnatal days 10-16. Mice tested at on postnatal day 17 were unaffected; but at the age of 60 days, changes in all three spontaneous-activity variables (mo- tor activity, rearing, and total activity) and an attenuation of habituation were seen at both doses of PCE. Chen et al. (2002) exposed young rats beginning at weaning (body weight, 45-50 g) to PCE orally at 5 or 50 mg/kg per day 5 days/week for 8 weeks. Effects on pain threshold, locomotor activity, reduction in body-weight gain, and seizure susceptibility were seen at both doses. The behavioral effects reported in rats (Nelson et al. 1980; Chen et al. 2002) and mice (Fredriks- son et al. 1993) exposed to PCE prenatally or postnatally suggest that there may be sensitive windows for neurobehavioral impairment during development. Further study comparing the neurobehavioral, neuro- chemical, and neuroanatomic changes that follow developmental exposure to PCE are needed. (See the following section.) Neurologic Effects Neurotoxicity and Neurobehavioral Effects Reviews by ATSDR (1997c), the California Environmental Protection Agency (CalEPA 2001), and EPA (2003, 2004) were consulted for this review. Data on accidental and controlled human inhalation and oral exposures and on experimental animal exposures are available. Acute inhalation and oral exposure of humans has been shown to induce symptoms of CNS de- pression (dizziness and drowsiness) (ATSDR 1997c). Electroencephalograpic (EEG) changes have been shown after acute inhalation exposure (Hake and Stewart 1977) and after subchronic inhalation exposure (5 days/week for 1 month; Stewart et al.1981) to PCE at 100 ppm. Neurobehavioral changesâsuch as changes in flash-evoked visual potentials, deficits in vigilance, and deficits in eye-hand coordinationâ were seen in volunteers exposed to PCE at 50 ppm 4 h/day for 4 days (Altmann et al. 1990, 1992). Oral exposure to doses of PCE ranging from 2.8 to 4 mL (about 4.2 to 6 g) given orally as an anthelminthic resulted in narcotic effects and such associated changes as inebriation, perceptual distortion, and exhilara- tion (ATSDR 1997c). A number of animal studies have shown effects on neurologic symptoms and biochemical end points in the brain after exposure to PCE. Acute and short-term inhalation exposure of rats, mice, and dogs to high concentrations of PCE (over 1,000 ppm) produced neurologic signs typical of anesthetic ef- fects, such as hyperactivity, ataxia, hypoactivity, and finally loss of consciousness (summarized by ATSDR 1997c). Savolainen et al. (1977) reported effects of PCE on open-field behavior in rats exposed to PCE at 200 ppm 6 h/day for 4 days. Activity was increased at 1 h but not 17 h after the last exposure, and reduced RNA content and increased cholinesterase were measured in the brain. Mattsson et al. (1998) showed effects of PCE on flash-evoked potentials, somatosensory evoked potentials, and EEG results af- ter acute exposure of rats to PCE at 800 ppm 6 h/day for 4 days when the animals were tested after expo- sure on the fourth day. Exposure of male Swiss mice to PCE at 596, 649, 684, or 820 ppm for 4 h reduced the duration of immobility experienced by mice when immersed in water (De Ceaurriz et al. 1983); the LOAEL was 649 ppm, and the NOAEL was 596 ppm. Albee et al. (1991) reported EEG changes and de- creased latency of flash-evoked potentials and somatosensory evoked potentials in male rats exposed to PCE at 800 ppm 4 h/day for 4 days.
Review of Toxicologic Studies 117 The effects of intermediate and subchronic inhalation exposure to PCE have also been investi- gated in several animal studies. Mattsson et al. (1998) found effects on flash-evoked potentials after 13 weeks of exposure of F344 rats to PCE at 800 ppm; the NOAEL was 200 ppm. Male Sprague-Dawley rats exposed continuously to PCE at 600 ppm for 4 or 12 weeks were reported to have reduced brain- weight gain, decreased regional brain weight, and decreased DNA in the frontal cortex and brainstem (Wang et al. 1993). Specific glial proteins (S100 and glial fibrillary acidic protein) and neuronal cy- toskeletal proteins (neurofilament 68-kD polypeptide) were also decreased; exposure to 300 ppm had no effect (and 300 ppm was the NOAEL). The authors concluded that the frontal cerebral cortex is more sen- sitive to PCE exposure than other parts of the brain and that cytoskeletal elements are more sensitive than cytosolic proteins. Rosengren et al. (1986a) exposed male and female Mongolian gerbils to PCE at 60 or 300 ppm for 3 months followed by 4 months without exposure. Changes in S100 (astroglial protein) and reduction in DNA concentrations in various brain regions were observed at 300 ppm, and reduction in DNA in the frontal cortex was seen at 60 ppm. Those effects were replicated by Karlsson et al. (1987). Kyrklund et al. (1988, 1990) reported changes in brain cholesterol, lipids, and polyunsaturated fatty acids in rats after exposure to PCE at 320 ppm for 30 or 90 days. Honma et al. (1980a,b) reported a decrease in acetylcholine in the striatum and an increase in glutamine, threonine, and serine. Kjellstrand et al. (1984) reported increased plasma butyrylcholinesterase concentrations and reduced body weight in white male and female MRI mice exposed to PCE at 37 ppm or greater for 30 days. Hepatic weight was increased at all concentrations (9, 37, 75, and 150 ppm) and continued to be increased 150 days after exposure; changes in hepatic structure were detected during exposure but were reversible. Cessation of exposure reversed the increase in butyrylcholinesterase concentrations. In experiments with various exposure dura- tions, increases in butyrylcholinesterase and hepatic weight were seen after exposure at a time-weighted average of 150 ppm for 30 days. Three studies have investigated the inhalation exposure of rodents to PCE during development (see also the section âDevelopmental Effectsâ above). Nelson et al. (1980) exposed pregnant rats to PCE at 100 or 900 ppm on gestation days 7-13 or 14-20. No effects were seen at 100 ppm, but pup weight gain was decreased in weeks 3-5 after exposure at 900 ppm. Developmental delays of offspring were seen in the exposed groups, and offspring exposed earlier in development had changes in an ascent test and a ro- torod test with some increase in motor activity. Significant reductions in acetylcholine were found in as- says of the whole brain (minus the cerebellum) after both exposure periods, and there were reductions in dopamine after exposure on gestation days 7-13. The authors concluded that animals exposed late in pregnancy had more behavioral changes than those exposed earlier. Manson et al. (1982), following up on the Tepe et al. (1982) study, found no postnatal effects of exposure to PCE at 1,000 ppm before mating and during pregnancy or only during pregnancy. Pregnant guinea pigs exposed to PCE continuously at 160 ppm on gestation days 33-65 had slightly altered brain fatty acid composition (Kyrklund and Haglid 1991), but the group sizes were very small (four litters each), and the statistical analyses treated each pup as an independent unit. Year long exposures of Mongolian gerbils to PCE at 120 ppm altered phospholipid content in cerebral cortex and hippocampus (Kyrklund et al. 1984) and caused reductions in cerebellar and hippo- campal taurine and increases in hippocampal glutamine (Briving et al. 1986a). However, there was no examination of nervous system structure in those studies to allow correlation of biochemical and behav- ioral changes. No structural CNS changes were reported in rats and mice exposed to PCE by inhalation at 200 or 400 ppm for 2 years (NTP 1986a). The effects of oral exposure to PCE have been investigated in only a few studies. Moser et al. (1995) examined adult female F344 rats in a functional observation-screening battery after either a single dose or repeated doses over 14 days. A single dose of PCE at 1,500 mg/kg caused increased lacrimation and gait scores and decreased motor activity; the LOAEL was 150 mg/kg. Effects were greater 4 h after dosing than 24 h after dosing. No effects were seen 24 h after dosing with PCE at 1,500 mg/kg per day for 14 days. EPA (2003) concluded that the difference in effects between single and repeated dosing may re- flect behavioral adaptation to PCE exposure. Warren et al. (1996) reported a transient decrease in a 90- min fixed-ratio 40 schedule of reinforcement in male mice exposed to PCE at 480 mg/kg immediately
118 Contaminated Water Supplies at Camp LejeuneâAssessing Potential Health Effects before testing; no effect was seen in animals exposed at 160 mg/kg. Blood concentrations correlated with administered dose, but brain concentrations were similar in the two groups. Chen et al. (2002) reported changes in pain threshold, locomotor activity, and seizure susceptibility (after pentylenetetrazol infusion) after exposure to a single dose of PCE at 500 mg/kg in adult rats; at 50 mg/kg, there were changes only in seizure susceptibility. The effects of PCE exposure on younger animals were reported in two studies. Exposure of young rats (45-50 g) to PCE at 5 or 50 mg/kg per day 5 days/week for 8 weeks resulted in effects on pain threshold, locomotor activity, and seizure susceptibility; changes in locomotion at the high dose; and re- duced body-weight gain at 5 and 50 mg/kg (Chen et al. 2002). The review by EPA (2003) raised serious questions about the design and interpretation of the study because of its observational nature and the mi- nor degree of change in latency scores. Fredriksson et al. (1993) exposed 10-day-old MRI mice to PCE orally at 5 or 320 mg/kg per day for 7 days and found increased locomotor activity and total activity at 60 days in both dose groups. Rearing behavior was decreased in the high-dose group. Habituation in re- sponse was seen in all three measures, PCE attenuated the response in locomotion and total activity but not rearing. Although EPA (2003) raised issues with the data interpretation in the study and the similarity of the two doses of PCE on locomotion and total activity, the effects on rearing were dose-related. In ad- dition, its criticism of using the pup as the statistical unit ignored to some extent the fact that individual pups were treated in the study. Two studies that used intraperitoneal exposure have evaluated the neurologic effects of PCE. Umezu et al. (1997) determined that righting reflex was affected after a single intraperitoneal dose of PCE of 4,000 mg/kg but not 2,000 mg/kg in 8-week-old male ICR mice. Ability to balance on a wooden rod was decreased at 2,000 mg/kg but not at 1,000 mg/kg or lower. Response rate on a fixed-ratio 20 schedule was affected at 2,000 mg/kg but not at 1,000 or lower 30 min after treatment. With a fixed-ratio 20 pun- ishment schedule, mice showed an increased response rate at 1,000 mg/kg but not at 500 mg/kg or lower. Motohashi et al. (1993) reported dose-dependent changes in circadian rhythm of 6-week-old male Wistar rats measured at least 1 week after intraperitoneal doses of PCE at 100, 500, or 1,000 mg/kg per day for 3 days. Recovery occurred 3-5 days after exposure ended. Results of studies that use intraperitoneal dosing cannot easily be compared with those of oral or inhalation exposures without pharmacokinetic modeling and development of appropriate conversion metrics. Cancer Gliomas were found in two female and four male F344/N rats exposed to PCE at 400 ppm (high- est concentration tested) and in one control male (NTP 1986a). The incidence of the tumor was not statis- tically significant, and one glioma was observed in the control group. Thus, the brain tumors were not considered to have been induced by exposure to PCE. Immunologic Effects The effects of PCE on the immune system have been studied less than the effects of TCE. For ex- ample, much work has been performed on evaluating the effects of TCE, but not PCE, on autoimmunity. Most immunologic research on PCE has been on allergic sensitization and immunosuppression. Allergic Sensitization There is no evidence that PCE can directly induce asthma, but there are suggestive data that it might modulate asthma. Seo et al. (2008) reported that rats given PCE by a single intraperitoneal injection at 0.1 mL/kg showed increased production of regulatory cytokines, including IL-4, and induced histamine release from basophils in animals immunized with a protein allergen. Similar effects were induced by
Review of Toxicologic Studies 119 PCE in vitro in cells from animals immunized with a protein allergen. Thus, PCE may act as an adjuvant to enhance existing allergic respiratory disease. Epidemiologic studies have indicated that the presence of PCE in the home environment is associated with reduced numbers of IFN-Î³ containing type 1 T cells (Lehmann et al. 2002). This regulatory cytokine could conceivably skew the normal ratio of type 1 to type 2 T cells to favor the development of asthma in children by allowing a greater proportion of type 2 cells to develop. Earlier studies showed that VOCs may modulate immune cells to favor induction of allergic re- sponses in young children (Lehmann et al. 2001). Further study is needed to clarify whether PCE can in- duce or modulate allergic diseases. Immunosuppression PCE was found to inhibit natural-killer-cell and cytotoxic T-cell activity after in vitro treatment of isolated mouse and rat spleen cells but not in in vivo experiments (Schlichting et al. 1992). In other stud- ies, inhalation of PCE vapors (50 ppm) reduced bactericidal activity against inhaled Klebsiella pneumo- niae and reduced survival after inhalation challenge with Streptococcus zooepidemicus in mice (Aranyi et al. 1986). It was hypothesized that those effects occurred because PCE inhibited alveolar macrophage activity. Such pulmonary host-resistance models can be influenced by a number of factors in the lung, including pulmonary macrophage function and inflammation. The available evidence does not allow any definitive conclusions to be drawn about the immunosuppressive potential of PCE. Hematopoietic Cancer F344/N rats were exposed to PCE by inhalation at 0, 200, and 400 ppm 6 h/day 5 days/week for 103 weeks (Mennear et al. 1986; NTP 1986a). A statistically significant increase in mononuclear-cell leu- kemia in both sexes was shown at both test concentrations, but no apparent dose-response relationship was observed. The NTP concluded that there was clear evidence of carcinogenicity of PCE in male F344/N rats and some evidence in female F344/N rats. Mononuclear-cell leukemia is a common spontaneous disease of aging F344 rats with incidences in NTP historical control males and females reported to be 50.5% and 28.1%, respectively (Haseman et al. 1998). The condition can exceed 70% in F344 controls (Caldwell 1999; Ishmael and Dugard 2006). Mononuclear-cell leukemia exhibited by F344 rats apparently arises from large granular lymphocytes; that leukemic origin is very uncommon in humans (Caldwell 1999). Given the high background incidence of mononuclear-cell leukemia and other tumors in F344 rats, a series of workshops convened by the NTP considered possible alteratives to the F344 rat as a model for use in bioassays (King-Herbert and Thayer 2006). More recently, a posting on the NTP Web site stated that the outbred Wistar Han rat will be used in standard bioassays rather than the F344 rat because of its attractive characteristics, including an overall low incidence of spontaneous background tumors (NTP 2007). The incidence of mononuclear-cell leu- kemia in the NTP (1986a) study showed moderate but not clearly PCE-dose-related increases. Consider- ing those factors, induction of mononuclear-cell leukemia in F344 rats exposed to PCE is unlikely to be relevant to prediction of human leukemia risk. SUMMARY The purposes of this section are to summarize information from key studies of the more important health effects of TCE and PCE and to describe the scientific evidence of an association between adverse effects in humans and various exposure conditions. TCE and PCE, in contrast with most other chemicals of environmental-health interest, have been extensively studied from a health standpoint. Nonetheless, there remain potential health effects of exposure to TCE and PCE on which there are inconclusive data or no data at all. The committee used a number of criteria in assessing the evidence in human case reports
120 Contaminated Water Supplies at Camp LejeuneâAssessing Potential Health Effects and from clinical studies and from controlled investigations with laboratory animals. Criteria used in reaching professional judgments included quality and reliability of key supporting studies, consistency of findings of similar studies, biologic plausibility, toxicologic significance, dose dependence and duration dependence, relative bioavailability and effects after different routes of exposure, and human relevance as determined by toxicokinetic and toxicodynamic concordance. Additional criteria have been used by study authors in assessing the implications of animal cancer bioassay results. The significance of those findings increases with increasing prevalence of tumors in multiple species, strains, and sexes; tumors at multiple sites; occurrence with more than one exposure route; progression from preneoplastic to benign to malig- nant; metastases; dose dependence; and low or nonexistent spontaneous tumor incidence in the test spe- cies. The primary adverse health effects of TCE and PCE and the conditions under which they were observed are presented graphically below. Figures were prepared for inhalation of TCE (Figure 4-1) and PCE (Figure 4-2) and for ingestion of TCE (Figure 4-3) and PCE (Figure 4-4). The figures are intended to give an overall view of the lowest exposures at which chemically induced anomalies of target organs were reported in reputable studies. Exposure concentrations high enough to also produce anesthesia or narcosis or nonspecific signs of general toxicity (such as malaise, reduced food consumption or reduced body- weight gain, or decreased survival) are indicated. Later in this chapter, LOAELs for selected end points are compared with estimated ranges of TCE and PCE doses by simultaneous ingestion and inhalation ex- perienced by former residents of Camp Lejeune from exposure to contaminated water supplies. Trichloroethylene Hepatic Effects Toxicity TCE, even in very high oral doses, has little ability to damage the livers of rodents or humans. A typical LOAEL in mice is 500 mg/kg. That dose, when given five times a week for 6 weeks, resulted in a modest increase in release of cytoplasmic enzymes from some damaged hepatocytes. Mice receiving TCE at 100 mg/kg per day on this regimen exhibited only a reversible increase in hepatic weight. The latter effect is not considered to be toxicologically significant. It should be recognized that TCE (and the other VOCs) at Camp Lejeune must undergo metabolic activation to exert cytotoxicity or mutagenicity and that mice metabolize substantially more TCE than rats and rats more TCE than humans. Reports of hepatotoxicity in patients anesthetized with TCE are rare in the medical literature. No evidence of hepatic injury was manifested in a man rendered unconscious for 5 days by ingesting about 1,370 mg/kg in a suicide attempt. Cancer The ability of TCE to cause cancer of the liver and other organs has been the subject of a number of lifetime oral-exposure and inhalation-exposure studies in mice and rats. Daily administration of high doses by both exposure routes resulted in an increased incidence of hepatocellular carcinoma in one strain of one species, the B6C3F1 mouse. It is unlikely that that tumor response is relevant to humans, because mice metabolically activate a much larger fraction of doses of TCE than do humans, the incidence of spontaneous hepatic tumors in male B6C3F1 mice is greater than 42%, and peroxisome proliferation, be- lieved to be a major mechanism by which key TCE metabolites induce hepatic tumors, is negligible in humans. However, some have questioned whether PPARÎ± action is the only relevant mode of hepatic car- cinogenesis of such chemicals.
Review of Toxicologic Studies 121 2,400 ppm (5 days) Auditory deficits in rats (Crofton and Zhao 1993) 2,000 ppm (9 days) Decreased performance on neurophysiological tests in rats (Bushnell and Oshiro 2000) 1,000 ppm (2 hrs) Decreased performance on visual-motor tests in humans (Vernon and Ferguson 1969) 600 ppm (104 wks) Kidney tumors in ratsa (Maltoni et al. 1988a) 500 ppm (6 months) Glomerulonephritis and increased urinary enzymes in rats (Mensing et al. 2002) 376 ppm (12 weeks) Testicular and epididymal damage, reduced testosterone and fertility in ratsa (Kumar et al. 2000a,b) 350 ppm (12 weeks) Changes in visual evoked potentials in rats (Blain et al. 1992) 200 ppm (5 days) Mild fatigue and sleepiness in humans (Stewart et al. 1970) 150 ppm (2 years) Lung tumors (Fukuda et al. 1983) and increased liver weight in mice (Kjellstrand et al. 1981) 50 ppm (6 weeks) Altered sleep patterns in rats (Arito et al. 1994) aGeneral toxicityâfor example, reduced body weight, weight gain, or food consumptionâthat may influence effects were observed in the study. FIGURE 4-1 Effects of exposure to TCE by inhalation. Duration of exposure should be considered in comparing end points that occur after different exposures. 1,600 ppm (13 weeks) Lung congestion in ratsa (NTP 1986a) 1,000 ppm (19 weeks) Mild glomerulonephropathy and increased pleomorphic nuclei in rats (Tinston 1995) 900 ppm (Gestation days 7-13, 14-20) Neurobehavioral impairment and neurotransmitter alterations in ratsa (Nelson et al. 1980) Effects on flash and somatosensory evoked potentials and electroencephalograms in rats 800 ppm (4 days) (Albee et al. 1991; Mattsson et al. 1998) 400 ppm (13 wks in mice, 2 years in rats) Centrilobular necrosis, bile stasis, hepatocellular proliferation in mice (NTP 1986a) Kidney tubular cell adenomas and adenocarcinomas in male ratsa (NTP 1986a) 300 ppm (premating, mating, gestation, and lactation through F1 and F2 generations) Reduced pre- and postnatal survival, pup body and organ weights in rats (Tinston 1995) 250 ppm (Gestation days 6-20) Delayed development of offspring in ratsa (Carney et al. 2006) 200 ppm (2 years in rats, acute in humans) Mononuclear cell leukemia and renal tubular cell karyomegaly in rats,a (NTP 1986) mild mucosal irritation in humans (ATSDR 1997b) 100 ppm (2 years) Renal tubular cell karyomegaly and liver tumors in micea (NTP 1986a) 50 ppm (3 hours in mice, 4 days in humans) Immunosuppression in mice (Aranyi et al. 1986) changes in visual evoked potentials, vigilance, and eye-hand coordination in humans (Altmann et al. 1990, 1992) 37 ppm (30 days) Increased brain butyrylcholinesterase in micea (Kjellstrand et al. 1984) aGeneral toxicityâfor example, reduced body weight, weight gain, or food consumptionâthat may influence effects were observed in the study. FIGURE 4-2 Effects of exposure to PCE by inhalation. Duration of exposure should be considered in comparing end points that occur after different exposures.
122 Contaminated Water Supplies at Camp LejeuneâAssessing Potential Health Effects 1,500 mg/kg (acute) Increased urinary NAG and microproteins in humansa (Bruning et al. 1998) 1,000 mg/kg (2 years) Kidney tumors in rats and liver tumors in micea (NTP 1990a) 875 mg/kg (premating, mating and Impaired fertility in mice (NTP 1986b) throughout several pregnancies) 500 mg/kg (4-6 weeks) Increased urinary enzymes in rats (Jonker et al. 1996) and increased serum enzymes in mice (Buben and OâFlaherty 1985) 250 mg/kg (13 weeks) Proximal tubular cell proliferation in rats (Mally et al. 2006) 145 mg/kg (premating, mating and Impaired fertility in rats (NTP 1986c) throughout several pregnancies) 30 mg/kg (pregnancy and lactation Reduced 2-DG uptake in brain, reduced hippocampal myelin, and increased activity in rat offspring (Taylor et al. 1985; Noland-Gerbec et al. 1986; Isaacson and Taylor 1989) 22 mg/kg (4-6 months) Immunosuppression in mice (Sanders et al. 1982) aGeneral toxicityâfor example, reduced body weight, weight gain, or food consumptionâthat may influence effects were observed in the study. FIGURE 4-3 Effects of exposure to TCE by ingestion. Duration of exposure should be considered in comparing end points that occur after different exposures. 600 mg/kg (32 days) Increased urinary enzymes in rats (Jonker et al. 1996) 536 mg/kg (1.5 years) Liver tumors in male micea (NCI 1977) 500 mg/kg (6 weeks) Increased serum enzymes and liver triglycerides in mice (Buben and OâFlaherty 1985) 386 mg/kg (1.5 years) Liver tumors in female micea (NCI 1977) 150 mg/kg (Acute-30 days in Transient increase in serum enzymes in mice (Philip et al. 2007), transient effects in functional observation mice, acute in rats) battery in rats (Moser et al. 1995) 50 mg/kg (8 weeks Change in locomotor activity, pain threshold, and seizure susceptibility in ratsa (Chen et al. 2002) beginning at weaning) aGeneral toxicityâfor example, reduced body weight, weight gain, or food consumptionâthat may influence effects were observed in the study. FIGURE 4-4 Effects of exposure to PCE by ingestion. Duration of exposure should be considered in comparing end points that occur after different exposures.
Review of Toxicologic Studies 123 Renal Effects Toxicity TCE has little ability to cause renal damage in rodents subjected to high oral or inhalation expo- sures for extended periods. A LOAEL of 500 mg/kg was found for mild renal injury in rats gavaged daily for 1 month. LOAELs of 250 and 500 mg/kg for proximal tubular-cell proliferation and karyomegaly, respectively, have been reported. Those responses were observed in male rats exposed orally five times a week for 13 weeks. Nephrosis occurs more commonly and is more serious in rats than in mice in lifetime cancer bioassays. The damage is apparently caused by reactive metabolites of the glutathione conjugation pathway. That pathway is similar qualitatively, but not quantitatively, in rats and humans (rats metaboli- cally activate about 10 times as much). Some workers exposed chronically by inhalation and dermally to TCE sufficient to produce neurologic effects experience renal epithelial toxicity. Cancer Chronic exposure to TCE at 1,000 mg/kg per day orally or 600 ppm by inhalation causes satura- tion of the oxidative metabolic pathway, which leads to increased formation of metabolites via the glu- tathione pathway. Some of the metabolites are cytotoxic and mutagenic. Male rats, but not female rats and not mice of either sex, exhibit a low incidence of renal-cell carcinoma when subjected to TCE at the aforementioned doses for their lifetimes. Increased rates of renal-cell cancer are also reported in some workers exposed for years to concentrations of TCE high enough to produce CNS effects and renal injury. The recurring cytotoxicity and compensatory cellular proliferation are thought to be prerequisites for re- nal-cell carcinoma (that is, coupled with the initiating action of mutagenic glutathione metabolites they act as promoters). Pulmonary Effects Toxicity Mice appear to be uniquely sensitive to pulmonary injury by TCE vapor. No reports of lung dam- age after TCE ingestion were located. Vacuolation of Clara cells was observed in mice that inhaled TCE at concentrations as low as 20 ppm 6 h/day for 5 days. Clara cells are nonciliated bronchiolar mucosal cells that have high CYP2E1 and CYP2F2 activities. The cytochrome P-450s catalyze the oxidation of TCE to chloral and diacetyl chloride, two putative cytotoxic and weakly mutagenic metabolites. Clara cells are numerous and are present throughout mouse airways; they are much less frequent in rats and rare in humans. CYP2E1 activity and TCE metabolism are undetectable in human lung preparations. Cancer Chronic TCE exposure has caused increased incidence of lung cancer in three strains of mice but not in rats. Lung tumors have not been seen in mice or rats in five oral TCE bioassays. That may be be- cause presystemic elimination of the orally administered chemical reduced the TCE that reached pulmo- nary tissues. The TCE-induced mouse lung tumors are not considered relevant to humans since mouse lung tumors are associated with Clara cells containing high CYP2E1 metabolizing activity and human lung contains few Clara cells and undetectable CYP2E1 activity.
124 Contaminated Water Supplies at Camp LejeuneâAssessing Potential Health Effects Fertility, Reproductive, and Developmental Effects Effects of TCE on fertility and reproduction have been seen in several investigations in rodents. In most cases, there were signs of general toxicity (such as body-weight and organ-weight changes and CNS depression) at the same exposure concentrations. Male rats exposed to TCE at 376 ppm 4 h/day 5 days/week for 12 or 24 weeks exhibited reduced body-weight gain, spermatoxicity, and reduced fecun- dity. CYP2E1, chloral formation, and dichloroacetyl adducts were found in testicular Leydig cells and epididymides of rats and were indicative of production of cytotoxic oxidative metabolites of TCE in the cells that were damaged. CYP2E1 has been found in human epididymal epithelium and Leydig cells. Some TCE oxidative metabolites have been identified in seminal fluid of TCE-exposed mechanics, al- though the relative metabolic capacities of human and rodent tissues have not been established. DuTeaux et al. (2004a,b) reported a dose-dependent reduction in the ability of sperm from TCE-treated rats to pene- trate ova from untreated females in vitro. The male rats ingested TCE at estimated doses of 1.6-3.7 mg/kg per day in drinking water for 14 days. Replication of those findings and further studies of the toxicologic and human significance of that sperm effect are warranted. Pregnancy outcomes were generally not affected by exposure to TCE at concentrations high enough to be maternally toxic, and there was no evidence of second-generation effects. Previously, there had been reports of cardiovascular defects in offspring of rodents exposed to TCE during gestation. More recently, well-conducted definitive experiments and a robust database have ruled out such developmental anomalies. The possibility of developmental neurotoxicity and immunotoxicity was raised in several pub- lications. Further research is needed to determine whether those results can be duplicated and, if so, to expand the scope of investigation and assess the human relevance. Cancer Leydig cell adenoma has been found in male rats in a 2-year oral and a 2-year inhalation cancer bioassay of TCE. It is the most frequently encountered testicular tumor in mice and rats. The spontaneous incidence in old F344 rats is as high as 90%. Most human testicular cancers originate in germ cells or Ser- toli cells and occur in young or middle-aged men. Leydig cell adenoma is rare in men, so spontaneous or TCE-induced Leydig cell adenoma is of questionable relevance to humans. Neurologic Effects TCE, like many other lipophilic VOCs, inhibits CNS functions as long as it is present at a suffi- cient concentration in neuronal membranes. Acute effects in humans are usually reversible and range from fatigue and dizziness to intoxication and anesthesia. A number of studies of human subjects have concurred that the inhalation LOAEL for impairment of motor or cognitive functions is 100-200 ppm for several hours. Residual neurotoxic effects (such as trigeminal and olfactory nerve impairment) have been reported in some workers exposed for years to vapor at concentrations that were probably in that range. Auditory deficits, reduced performance of tasks, and other effects were observed in more highly exposed rats, but tolerance usually developed over days or weeks of exposure. LOAELs of 350 and 50 ppm have been reported for changes in visual evoked potentials in rabbits and decreased wakefulness in rats, respec- tively. The toxicologic significance of those responses in rodents that inhaled TCE several hours a day for weeks has not been established. No definitive oral neurologic studies of TCE were located. Immunologic Effects TCE causes allergic sensitization in animal studies, including contact dermatitis and exacerbation of asthma. Some of those effects have been reported in humans after chronic occupational exposure to
Review of Toxicologic Studies 125 VOCs by inhalation at relatively high concentrations, but further studies are needed to determine whether TCE can induce or modulate allergic diseases in humans. Immunosuppression has also been shown in animal studies after TCE exposure, but it is unclear whether the effects are relevant to humans. Workers exposed to TCE showed increases in IL-2 and IFN-Î³ and an increase IL-4, but interpretation of these changes is difficult, and the data are too sparse to support definitive conclusions. Toxicologic studies have also shown exacerbation of autoimmune diseases in a genetically modified mouse model (MRL+/+). The relevance of those findings to humans is unclear, although epidemiologic studies have shown a relation- ship between solvent exposure and scleroderma, glomerulonephritis, and other immune-related diseases (see Chapter 5). Tetrachloroethylene Hepatic Effects Toxicity PCE, like TCE, has little ability to cause acute, subacute, or chronic hepatotoxicity in rodents or humans. PCE is somewhat more potent because of formation of some additional reactive metabolites. An acute oral LOAEL of 150 mg/kg was reported by Philip et al. (2007), but the serum concentration of a liver-specific enzyme in mice progressively declined as the mice were treated over 30 consecutive days. A NOAEL of 1,440 mg/kg per day was reported in rats that consumed PCE in drinking water for 90 days (Hayes et al. 1986). As described in Chapter 3, ingestion of a chemical in divided doses over several hours reduces its potency. In addition, rats are less susceptible than mice because of their lower capacity for activating PCE metabolically. Humans have even lower capacity than rats. Cancer There is clear evidence that near-lifetime inhalation or ingestion of PCE, like that of TCE, results in increased incidence of liver cancer in B6C3F1 mice. Similarly exposed rats do not develop hepatic tu- mors. PCEâs LOAEL is 386 mg/kg for 78 weeks compared with TCEâs LOAEL of 1,000 mg/kg for 103 weeks. Trichloroacetic acid, a major metabolite of both PCE and TCE, produces peroxisome proliferation in mouse liver but not rat or human liver. The very high spontaneous hepatic-tumor incidence in B6C3F1 mice and formation of substantially greater quantities of reactive metabolites suggest that mouse hepatic tumors may be of little relevance to humans. Renal Effects Toxicity PCE is somewhat more toxic to the kidneys than TCE. A LOAEL of PCE of 600 mg/kg per day for renal damage was found in rats gavaged for 32 consecutive days. In contrast, consumption of PCE at up to 1,400 mg/kg per day in drinking water for 90 days failed to damage ratsâ kidneys. That discrepancy can be attributed largely to the kidneysâ receipt of lower tissue doses when exposure was in drinking wa- ter. A NOAEL of 400 ppm and a LOAEL of 1,000 ppm are described for nephrotoxicity in rats that in- haled PCE several hours a day for a month or more. Karyomegaly was seen in the renal tubular cells of mice and rats that inhaled PCE chronically at as low as 100 and 200 ppm, respectively; the nuclear enlargement may be a predecessor of neoplasia, but a definite link has not been established. Renal effects of PCE are due primarily to metabolites formed via the glutathione conjugation pathway. Equivalent inha-
126 Contaminated Water Supplies at Camp LejeuneâAssessing Potential Health Effects lation exposures of rats and humans to PCE at 160 ppm for 6 h showed that biotransformation by the glu- tathione metabolic pathway was 10 times greater in the rats (Volkel et al. 1998). Cancer Chronic inhalation of PCE at 200 or 400 ppm produced renal tubular-cell karyomegaly, hyperpla- sia and a low incidence of tubular-cell adenoma and carcinoma in male rats. Renal tumors did not occur in female rats or in mice of either sex, although these animals did exhibit karyomegaly. Pulmonary Effects Toxicity There is little evidence of lung injury by inhaled PCE in laboratory animals or humans. Inhalation experiments with human subjects indicate a NOAEL of 150 ppm and a LOAEL of 200-300 ppm for mild irritation of nasal passages. Pulmonary-function measurements do not reveal decrements at those concen- trations. Intermittent inhalation of PCE at 1,600 ppm for 13 weeks produced pulmonary congestion in rats; 800 pm did not. There is one report (Aoki et al. 1994) of epithelial degeneration in mice that inhaled PCE at 300 ppm 6 h/day for 5 days. The change was more severe in the olfactory than in the respiratory mucosa. Cancer No increases in proliferative lesions or neoplasms of the respiratory tract have been seen in a chronic oral or inhalation cancer bioassay in mice and rats. Although CYP2E1 is abundant in mouse lung, that cytochrome P-450 isozyme is not active as a catalyst of PCE metabolism in the respiratory tract of other rodents or humans. Other Cancers An increased incidence of mononuclear-cell leukemia was found in male and female F344 rats that inhaled PCE at 200 or 400 ppm for 103 weeks. The increases were not dose-dependent and were within the incidence range of mononuclear-cell leukemia often seen in control F344 rats. The NTP is no longer using the F344 strain in its cancer bioassay program, because of its high rates of spontaneous can- cer of several types. Mononuclear-cell leukemia is rare in people. Thus, that form of leukemia in F344 rats has been judged not to be relevant to humans. Animal cancer bioassay outcomes relevant to human leukemia, multiple myeloma, and non-Hodgkin lymphoma have not been reported. Fertility, Reproductive, and Developmental Effects Information on potential effects of PCE on fertility and reproduction is limited. Inhalation of PCE for 5 days did not affect sperm morphology in rats but did result in increased incidence of abnormal sperm heads in mice. The NOAEL and LOAEL for that effect were 100 and 500 ppm, respectively. Long- term exposure of male and female rats to PCE vapor for two generations resulted in CNS depression, de- creased body weight during lactation, and nephrotoxicity at 1,000 ppm. There were reductions in live births, litter size, survival, and body weight in the F2 progeny at that vapor concentration. Those adverse effects may be secondary to maternal body-weight loss and toxicity. More data are needed to clarify the effects of PCE on reproductive function.
Review of Toxicologic Studies 127 A number of oral and inhalation studies of potential developmental effects of PCE have been conducted in rodents. Experimental protocols have included inhalation of PCE at 300-1,000 ppm before, during, or after pregnancy. Manifestations of developmental delay (such as reduced ossification of verte- brae and soft-tissue dysplasias) have been reported in pups at the relatively high concentration. Ingestion of PCE at 900 mg/kg per day on days 6-19 of gestation, for example, resulted in increased resorptions, reduced weight, and microphthalmia or anophthalmia in rat pups. That daily dose was so high that mater- nal ataxia and weight loss occurred. Developmental effects at lower concentrations were relatively minor and were not indicative of teratogenicity. Neurotoxicity Neurologic Effects Ingestion and inhalation of sufficient doses of PCE produce CNS depression in rodents and hu- mans. Because PCE is more lipophilic than TCE, it is moderately more potent as a CNS depressant. Defi- cits in neurophysiologic functions have been reported in volunteers exposed to PCE at as low as 50 ppm for 4 h/day for 4 days (Altmann et al. 1990, 1992). A number of animal studies have revealed neurobe- havioral and neurochemical changes in the brains of animals that inhaled PCE at several hundred parts per million for various periods. Mattsson et al. (1998), for example, found altered flash-evoked potentials in rats after 13 weeks of exposure at 800 ppm, but not at 200 ppm. Wang et al. (1993) measured decreases in regional brain weight, DNA content, and glial proteins in rats exposed continuously to PCE at 600 ppm for 4 or 12 weeks. Few researchers, however, have evaluated PCE-induced neurobehavioral and neuro- chemical changes in the same animals, so interpretation of many of the data is difficult. Neurodevelopmental Effects Concerns about possible neurodevelopmental effects in children exposed to PCE prompted sev- eral investigations in animals. Chen et al. (2002), for example, described changes in locomotor activity, pain threshold, and pentylenetetrazol-induced seizure thresholds in young rats dosed orally with PCE at 50 mg/kg per day for 8 weeks. Exposure of pregnant rats to PCE at 900 ppm resulted in pups with dimin- ished brain acetylcholine and dopamine concentrations and with neurobehavioral changes on certain days of testing; inhalation of PCE at 100 ppm was without effect. Such reports suggest that there may be peri- ods of neurologic development during which sufficiently high PCE exposures are detrimental. Additional research is needed to determine whether gestational, neonatal, or childhood exposure to such solvents can impair CNS development and function. Immunologic Effects Little information is available on the potential of PCE to suppress the immune system or to in- duce autoimmune diseases. In one study, PCE was found to suppress natural-killer-cell and T-cell activity in vitro but to have no effect on rats in vivo. In a second study, inhalation of PCE at 50 ppm reduced bac- tericidal activity in mice subjected to inhaled microorganisms. Further investigations of PCE are war- ranted in light of the apparent effects of TCE on the immune system. HAZARD EVALUATION OF TRICHLOROETHYLENE AND PERCHLOROETHYLENE EXPOSURE FOR SELECTED END POINTS The committee used several approaches to consider the health significance of the solvents found in the water supply at Camp Lejeune. Hazard can be defined as the intrinsic characteristic toxicity of a
128 Contaminated Water Supplies at Camp LejeuneâAssessing Potential Health Effects chemical compound. The hazard evaluation provides information on the inherent toxic potential of an ex- posure and is not meant to provide a quantitative estimate of risk. This approach compares the lowest doses of TCE and PCE at which adverse effects were observed in laboratory animals (the LOAELs) with a range of estimated doses from the Camp Lejeune water supply. It is one line of evidence in assessing possible relationships between exposure to TCE and PCE in water at Camp Lejeune and potential health effects. The lowest dose at which an adverse health effect was observed, the LOAEL, may be subject to some uncertainty, depending on a number of factors, including the doses that were studied, the end point chosen, and the method used to assess the end point; for example, death as an observed LOAEL end point is more certain than a subtle change in an end point that is reversible and of unknown biologic signifi- cance. LOAELs from animal studies, on average, are associated with a 10% increase in response rate and can be associated with various risk levels because the statistical power of the studies does not allow ob- servation of lower levels of exposure. Thus, LOAELs do not define a level below which no adverse ef- fects can occur. Nevertheless, determination of a LOAEL generally provides a useful measure of toxic potency. NOAELs are hampered by more uncertainty. A NOAEL is the highest experimental dose at which an adverse effect did not occur. An experimentally determined NOAEL may be substantially lower than the actual NOAEL if the doses administered were too low. The present hazard evaluation was based on LOAELs for selected toxicity end points as described below. The toxicologic databases on TCE and PCE are extensive, but some data gaps remain for a few end points. LOAELs observed in animal studies selected for this dose comparison represent a range of adverse effects and oral doses. The particular end points were chosen in part because it was assumed that they may be relevant to humans. For TCE, renal tumors in rats were chosen for a chronic high-dose end point (LOAEL, 1,000 mg/kg per day for lifetime oral exposure [NTP 1990a]), renal toxicity in rats was chosen for the medium dosage range (LOAEL, 250 mg/kg per day for 13 weeks [Mally et al. 2006]), and immunosuppression in a sensitive strain of mice was chosen at the lower end of the dosage spectrum (LOAEL, 22 mg/kg per day in drinking water for 4 or 6 months [Sanders et al. 1982]) (see Figure 4-3 and Table 4-3). For PCE: renal toxicity in rats (600 mg/kg per day for 32 days [Jonker et al. 1996]) was se- lected at the upper end of a series of LOAELs, and neurologic changes in young rats (50 mg/kg per day for 8 weeks [Chen et al. 2002]) at the lower end of LOAEL doses (see Figure 4-4 and Table 4-4). Uncertainty is associated with the TCE and PCE water concentrations used in the hazard evalua- tion because they are based on the relatively few mixed water samples analyzed (see Chapter 2). Only a small set of water-quality measurements are available, and those were taken during the 5 years before the contaminated wells were closed, so it is unknown how well they represented the conditions during the preceding decades. In addition, concurrent exposures to organic solvents may have occurred at Camp Le- jeune. Studies of mechanisms of VOC interactions (see Chapter 3) indicate that such concurrent exposure is not likely to result in greater than an additive effect. Relatively low doses of multiple VOCs are unlikely to affect the magnitude of adverse health effects appreciably. Additivity is not formally incorpo- rated into this appraisal. The exercise below is not a health risk assessment. Several assumptions (described below) were used to derive the comparisons, so there is uncertainty and variability in the values. The intent is to pro- vide general comparisons of the lowest doses at which specific adverse health effects were observed in experimental toxicologic studies with a range of estimated contaminant concentrations that may have oc- curred in the Camp Lejeune water supply. The following describes the assumptions in the evaluation and illustrative calculations. To pro- vide a standardized basis for comparison, the lowest doses at which a specific adverse effect was seen in toxicologic studies and the exposure estimates are both expressed in standard terms of milligrams of chemical per kilogram of body weight per day (mg/kg per day). Standard assumptions commonly used for hazard evaluations are that adults weigh an average of 70 kg and drink an average of 2 L of water per day and that children weigh an average of 10 kg and drink 1 L of water per day. Exposure via inhalation and dermal absorption of VOCs from water during showering, bathing, dishwashing, and other household ac-
Review of Toxicologic Studies 129 TABLE 4-3 LOAELs from Animal Studies Used for Comparison with Estimated Daily Human Doses to TCE Related to Water-Supply Measured Concentrations Range of Doses End Point LOAEL, mg/kg per day High Kidney cancer, rats 1,000 Medium Kidney toxicity, rats 250 Low Immunosuppression, mice (sensitive strain) 22 TABLE 4-4 From Animal Studies Used for Comparison with Estimated Daily Human Doses to PCE Related to Water-Supply Measured Concentrations Range of Doses End Point LOAEL, mg/kg per day High Kidney toxicity, rats 600 Low Neurotoxicity, rats 50 tivities has been shown experimentally to account for as much exposure as that from drinking water that contains the chemicals (see Chapter 3). Therefore, to account for potential inhalation and dermal uptake in addition to ingestion in drinking water, an intake of 4 L/day is assumed for adults and 2 L/day for chil- dren. This calculation, therefore, takes into account all three routes of exposureâingestion, inhalation, and dermalâof both adults and children. Considerable toxicologic data on VOCs are available from inha- lation studies. The range of adverse effects is presented in Figures 4-1 and 4-2, but absorbed doses were usually not determined. Duration of exposure is usually specified in animal studies. A conservative as- sumption used in this hazard evaluation is that humans receive the stated dose daily, although that is very unlikely inasmuch as data presented in Chapter 2 indicate that daily exposures were highly variable. It is important to note that the evaluation has not taken into account uncertainties and additional considerations (see Chapter 3) related to potentially sensitive populations (such as fetuses and the eld- erly), possible human interindividual variability in response related to sex and genetic background, such lifestyle factors as level of exercise , underlying diseases, and VOC interactions. Nevertheless, as dis- cussed in Chapter 3, rodents absorb a greater fraction of inhaled VOCs and metabolically activate a sub- stantially greater proportion of their internal dose and are therefore more susceptible than humans to most adverse effects of TCE and PCE. Chapter 2 summarizes the water-supply data available from the Tarawa Terrace and Hadnot Point water systems. Among the measurements with reported values, TCE concentration in mixed water sam- ples from the Hadnot Point water supply ranged from 1 to 1,400 Âµg/L (see Table 2-11). Water samples with detectable PCE from the Tarawa Terrace water supply ranged from 1 to 215 Âµg/L (Maslia et al. 2007). Given the sparse information regarding the range and magnitude of contaminant concentrations in the Camp Lejeune water supply, values that correspond to half the highest measured value, the highest measured value, and twice the highest measured value were selected for this exercise: TCE at 700, 1,400, and 2,800 Âµg/L and PCE at 100, 200, and 400 Âµg/L. The following calculation was carried out to obtain an estimate of human daily exposure: esti- mated human daily dose (mg/kg per day) = [mixed water concentration (Âµg/L) Ã estimated daily intake (oral, inhalation, and dermal) (L/day)]/[body weight (kg)]. A sample calculation follows. For Hadnot Point, the highest measured concentration of TCE in mixed water was 1,400 Âµg/L. For an adult human, the daily dose received from water containing TCE at 1,400 Âµg/L is estimated to be 1,400 Âµg/L Ã 4 L/day = 80 Âµg/kg per day = 0.08 mg/kg per day. 70 kg Half the highest measured TCE concentration in the water supply (700 Âµg/L) yields an estimated dose of 0.04 mg/kg per day for adults, and twice the highest measured concentration of TCE (2,800 Âµg/L) yields
130 Contaminated Water Supplies at Camp LejeuneâAssessing Potential Health Effects an estimated dose of 0.2 mg/kg per day for adults. For a child, the daily dose received from water contain- ing TCE at 1,400 Âµg/L is estimated to be 1,400 Âµg/L Ã 2 L/day = 280 Âµg/kg per day = 0.3 mg/kg per day. 10 kg Half the highest measured TCE concentration in the water supply (700 Âµg/L) yields an estimated dose of 0.1 mg/kg per day for a child, and twice the highest measured concentration of TCE (2,800 Âµg/L) yields an estimated dose of 0.6 mg/kg per day for a child. Table 4-3 shows the LOAELs from animal studies used to compare with the estimated human TCE doses related to a range of possible water-supply exposure concentrations. A comparison of LOAELs for health end points selected from TCE animal studies with the exposure estimates is summa- rized here: ï· Kidney cancer. The LOAEL of TCE for lifetime oral exposure leading to kidney cancer in the rat is 1,000 mg/kg per day (NTP 1990a). The estimated human adult dose at Camp Lejeune is 25,000 times lower than the LOAEL for exposure at half the highest water-supply concentration, 12,500 times lower than the LOAEL for exposure at the highest concentration, and 5,000 time lower than the LOAEL for ex- posure at twice the highest concentration for a lifetime exposure. For a child, the comparable estimates are 10,000, 3,350, and 1,700 time lower than the LOAEL, respectively. ï· Renal toxicity. The LOAEL of TCE for renal toxicity in the rat dosed orally for 13 weeks is 250 mg/kg per day (Mally et al. 2006). The estimated human adult dose at Camp Lejeune is 6,250 times lower than the LOAEL for exposure at half the highest water-supply concentration, 3,125 times lower than the LOAEL for exposure at the highest concentration, and 1,250 times lower than the LOAEL for exposure at twice the highest concentration. For a child, the comparable estimates are 2,500, 830, and 415 times lower than the LOAEL, respectively. ï· Immunosuppression. The LOAEL of TCE for immunosuppression in a sensitive strain of mouse ingesting TCE for 4 or 6 months is 22 mg/kg per day (Sanders et al. 1982). The estimated human adult dose at Camp Lejeune is 550 times lower than the LOAEL for exposure at half the highest water-supply concentration, 275 times lower than the LOAEL for exposure at the highest concentration, and 110 times lower than the LOAEL for exposure at twice the highest concentration. For a child, the comparable esti- mates are 220, 75, and 40 times lower than the LOAEL, respectively. These differences are relatively smaller than for kidney cancer and kidney toxicity. As stated earlier in the chapter, uncertainties exist re- garding this end point since there is relatively little toxicologic information on TCE and immune effects. Additional research may be needed on the potential immunosuppressive effects of TCE. For PCE, the daily dose received from water at the maximum measured concentration (200 Âµg/L) in the water supply for an adult human is estimated to be 200 Âµg/L Ã 4 L/day = 0.01 mg/kg per day. 70 kg Exposure to half the highest measured water supply concentration (100 Âµg/L) yields a dose of 0.006 mg/kg per day for an adult human and exposure to twice the highest measured water supply concentration (400 Âµg/L) yields a dose of 0.02 mg/kg per day. For a child, the daily dose received from water contain- ing PCE at the maximum measured concentration (200 Âµg/L) is estimated to be 200 Âµg/L Ã 2 L/day = 0.04 mg/kg per day. 10 kg
Review of Toxicologic Studies 131 Exposure to half the highest measured water supply concentration (100 Âµg/L) yields a dose of 0.02 mg/kg per day for a child and exposure to twice the highest measured water supply concentration (400 Âµg/L) yields a dose of 0.08 mg/kg per day. A comparison of LOAELs for each of the two health end points selected from PCE animal studies (Table 4-4) with the estimated doses from the water supply is summarized here: ï· Renal toxicity. The LOAEL for renal toxicity in the rat dosed orally with PCE for 32 days is 600 mg/kg per day (Jonker et al. 1996). The estimated human adult dose at Camp Lejeune is 100,000 times lower than the LOAEL for exposure at half the highest water-supply concentration, 60,000 times lower than the LOAEL for exposure at the highest concentration, and 30,000 times lower than the LOAEL for exposure at twice the highest concentration. For a child, the estimates are 30,000, 15,000, and 7,500 times lower than the LOAEL, respectively. ï· Neurotoxicity. The LOAEL of PCE for neurotoxic effects in rats is 50 mg/kg per day for 8 weeks (Chen et al. 2002). The estimated human adult dose at Camp Lejeune is 8,300 times lower than the LOAEL for exposure at half the highest water-supply concentration, 5,000 times lower than the LOAEL for exposure at the highest concentration, and 2,500 times lower than the LOAEL for exposure at twice the highest concentration. For a child, the comparable estimates are 2,500, 1,250, and 625 times lower than the LOAEL, respectively. As noted earlier in this chapter, there is a need for additional research to clarify the neurotoxic effects of PCE. The comparisons above included health end points observed in animals that were considered relevant to humans. Renal toxicity and cancer, neurotoxicity, and immune-related effects have been re- ported in some epidemiology studies and in clinical reports. The dose comparisons1 suggest considerable differences between the estimated doses from human exposure to contaminated water supplies at Camp Lejeune under conservative assumptions of exposure and the lowest doses associated with the develop- ment of renal toxicity, kidney cancer, neurotoxicity, and immunosuppression in rodents. The drinking- water doses at Camp Lejeune are substantially lower. As pointed out in this section, however, each and 1 One member, Lianne Sheppard, objected to inclusion of the hazard evaluation in the report as written and offered the following explanation: âComparison of toxicology-based LOAEL values with estimated exposures to the Camp Lejeune population uses questionable logic to support inference that adverse health effects are unlikely to have occurred. Although LOAEL estimates give evidence about the presence of a hazard, they should not be used to make inference about the absence of hazard at lower doses. The absence of evidence of a hazard (e.g., at levels be- low the LOAEL) cannot be equated with evidence of the absence of hazard (Altman and Bland 1995; Fleming 2008). Because of their small sample size, animal studies are only able to identify hazards that induce high levels of response (on average 10% increase in response for the LOAEL). Moreover, levels of excess response considered acceptable in humans are much lower than 1 in 10, typically on the order of 1 in 10,000 to 1 in 1 million (EPA 2005). While low-dose extrapolation involves additional untestable assumptions, dividing the LOAELs by 1,000 to 100,000 provides an alternative approach to the informal hazard evaluation presented above. This second approach compares Camp Lejeune exposures with an acceptable hazard in humans, as extrapolated from toxicologic studies. The results lead to strikingly different conclusions because they yield acceptable hazards that are both larger and smaller than the estimated exposures; indeed, some are several orders of magnitude lower than Camp Lejeune expo- sures. Alternatively, standard practice would replace informal hazard evaluation with a formal risk assessment, although this task was outside the committee charge. Despite my reservations on this one area of the assessment, I support the overarching findings and recommendations of the report.â Other members disagree with Dr. Sheppardâs characterization that the hazard evaluation is based on questionable logic. The reasons for this are stated in the text. The validity of results using the approach she outlines above is ques- tioned by some committee members. There were varying views among committee members on the value of the in- formation generated by the hazard evaluation effort, ranging from members who found it quite useful because it provided a rough benchmark for speculating about the likelihood of adverse health effects, to members who placed less reliance on results, given limited exposure information and their uncertainty about the applicability of toxi- cologic information. Regardless of the approach taken to the hazard evaluation, however, all committee members strongly support the overarching findings and recommendations of the report.
132 Contaminated Water Supplies at Camp LejeuneâAssessing Potential Health Effects every source of uncertainty (e.g., interindividual variability, lifestyle, genetic background, exposure as- sessment, completeness of the database) has not been factored into this estimate since it is a hazard evaluation procedure and not a health risk assessment. ALLOWABLE LIMITS OF VOLATILE ORGANIC COMPOUNDS IN DRINKING WATER Current regulatory standards termed maximum contaminant levels (MCLs) for several VOCs in drinking water, including TCE and PCE, were developed by EPA in the middle 1980s (50 Fed. Reg. 46880 ; 52 Fed. Reg. 25690 ; Cotruvo 1988). Under the U.S. Safe Drinking Water Act, the public-health goal or maximum contaminant level goal (MCLG) for a compound was initially determined. The MCLG is the concentration that would result in âno known or anticipated adverse effect on healthâ with a large margin of safety. Second, an MCL, or enforceable standard, was set as close as feasible to the MCLG; technical and economic factors were taken into consideration. EPA consulted the International Agency for Research on Cancer guidelines when assessing epidemiologic and animal cancer data and in its own qualitative weight-of-evidence scheme for determining the potential for a compound to increase cancer risk in humans. TCE and PCE fell into category I in the latter scheme, in which the MCLG by definition equals zero as an aspirational goal. Economic considerations for water treatment were also de- liberated. Technical feasibility focused on analytic considerations; the lowest concentrations that can be reliably detected within specified limits of precision and accuracy during routine laboratory operations (practical quantitation limits) were determined. With that approach, an MCL of 0.005 mg/L (5 Âµg/L or 5 ppb) was set for selected VOCs, including TCE and PCE. In 2005, EPA issued new guidelines for carcinogen risk assessment in which incorporation of in- creased scientific understanding of the biologic mechanisms that can cause cancer was supported for in- clusion in risk assessments with other improved risk-assessment practices (EPA 2005). In the more than 20 years since the original MCLs were established, considerable kinetic and biologic mechanism-of- action information on TCE and PCE has been published, as reviewed in the present report. There are dif- ferent approaches to risk assessment that yield different results. At least one recent study has explored different approaches, including the use of contemporary published elements of TCEâs biologic mode of action and a cancer-risk model that was the best fit to the data (Clewell and Andersen 2004). The latter approach yielded a TCE concentration of 265 Âµg/L in drinking water; below this concentration, a car- cinogenic hazard to human health was deemed unlikely. This is one example of the possible application of toxicologic and mechanistic biologic data to a cancer health risk assessment for TCE, which yields a value greater than one based on analytical limits of detection. EPA is currently updating its risk assess- ments on TCE and PCE and is considering new data and different assessment approaches as part of its reassessments. In summary, the few TCE and PCE measurements available from mixed drinking-water samples at Camp Lejuene (see Chapter 2) indicated that some samples exceeded the MCLs derived as briefly described above. CONCLUSIONS TCE and PCE are well-studied compounds compared with most other compounds of environ- mental concern. On the basis of the review presented above, the committee concludes that the strongest evidence of health effects of relevance to humans are renal toxicity, kidney cancer, neurobehavioral ef- fects, and immunologic effects, which have generally been observed at high concentrations in a work- place setting and in exposure to tens to thousands of milligrams per kilogram of body weight in animal studies. Discussion of the toxicologic evidence in context with the epidemiologic evidence on TCE and PCE (presented in Chapter 5) is provided in Chapter 7. The evidence on renal toxicity and cancer is par- ticularly convincing because concordance has been found in the bioactivation of TCE and PCE and in their modes of action in rodents and humans. However, gaps in the toxicologic database preclude drawing conclusions about some other health effects related to the nervous system and the immune system, par-
Review of Toxicologic Studies 133 ticularly with regard to potential effects on the developing or young animal. Implicit inherent limitations of toxicologic studies are that relatively homogeneous populations of laboratory animals are used and ex- posures are typically to single chemicals. On average, the lowest increase in effect that can usually be de- tected (LOAEL) is around 10% due to statistical power related to the number of animals that can be tested in any one study. In the instances of TCE and PCE, however, rodents are more susceptible to toxic ef- fects. A central issue in toxicology (and at Camp Lejeune) is whether doses were sufficient to produce specific adverse effects. The lowest doses at which adverse health effects have been seen in animal or clinical studies are many times higher than the worst-case (highest) assumed exposures at Camp Lejeune. However, that does not rule out the possibility that other, more subtle health effects that have not been well studied could occur, although it somewhat diminishes their likelihood. Another important issue is whether any adverse effects that may have occurred were reversible or permanent and (still) detectable when an epidemiology study might be conducted. Observations in animal studies indicate that very high acute or chronic doses of TCE or PCE are necessary to injure renal proxi- mal tubular cells. Results of occupational-exposure studies indicate that relatively high, chronic exposures result in modest, reversible changes in the most sensitive indexes of renal injury in workers. Thus, it is unlikely that renal toxicity would be a useful end point to examine in future epidemiology study of Camp Lejeune residents. A similar conclusion can be drawn with regard to the occurrence and detection of he- patic toxicity. Reproductive and developmental effects in rodents were quite modest and often secondary to general toxicity, decreased food intake, and reduced body-weight gain resulting from high maternal doses of TCE and PCE. The toxicologic data provide strong evidence that neither solvent is associated with congenital malformations in rats. Thus, on the basis of this review, reproductive effects and hepa- torenal toxicity are probably not of great concern at Camp Lejeune. There is reasonable interspecies concordance between rats and humans in the bioactivation of TCE and PCE and in their mode of induction of kidney cancer. A low incidence of kidney cancer has been seen in workers exposed for many years to TCE at concentrations high enough to cause dizziness, headache, and other reversible neurologic effects. The background incidence of kidney cancers in unex- posed persons is minimal. Nevertheless, there is little likelihood of identifying any increased incidence of renal tumors in the relatively small population that may be available for study at Camp Lejuene. Irreversible neurobehavioral effects associated with solvent exposure generally are chronic and result from high doses. Solvent abusers and workers chronically exposed to high vapor concentrations may exhibit various neurobehavioral effects and residual brain damage. Fetuses, infants, and young chil- dren exposed to such organic solvents as TCE and PCE at lower concentrations may experience subtle neurodevelopmental effects, but no relevant investigations were identified. There are few data from ani- mal studies on this topic. Immune suppression and autoimmunity related to TCE exposure have been demonstrated in some sensitive animal models. TCE-induced glomerulonephritis and scleroderma occur in low incidences in highly exposed worker populations. Much less is known about the potential immunologic effects of PCE (particularly as related to exposures during development), which may warrant further consideration for inclusion in studies of populations exposed to TCE or PCE.