4
Review of Toxicologic Studies

This chapter summarizes findings of animal studies of trichloroethylene (TCE) and tetrachloroethyle (perchloroethylene, PCE) toxicity and relevant end points. The review was based in part on previously 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 concordance 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 concern. To manage the vast amount of information on each, we provide different degrees of review according to the findings from the exposure assessment regarding the frequency and concentrations of the contaminants 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 review of the body of available peer-reviewed literature. Because TCE and PCE have some of the same metabolites and effects, salient finding of studies of PCE are discussed in relevant sections of the TCE review. 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-



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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

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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-2′-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-2′-deoxyguanosine adducts in liver DNA were significantly in- creased throughout much of the study with TCE at 1,200 mg/kg. Buben and O’Flaherty (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 O’Flaherty (1985) reported that plots of their mouse subchronic-hepatotoxicity data against urinary-metabolite excretion values indicated that TCE’s 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 acid’s activating a nuclear protein known as the peroxisome-proliferator-activated receptor alpha (PPARα). PPARα-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

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Contaminated Water Supplies at Camp Lejeune—Assessing Potential Health Effects 92 (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 TCE’s 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 TCE—trichloroacetic acid, dichloroacetic acid, and chloral hydrate—have 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 PPARα 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

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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 µM (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 rats’ 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 PPARα. Stimulation of PPARα 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). PPARα-null mice are refractory to those effects, which are associated with carcinogenesis. Mice expressing human PPARα fail to show increases in markers of cell proliferation and are resistant to liver cancer if treated with PPARα agonists (Morimura et al. 2006; Yang et al. 2008). The concentration of PPARα 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 compound’s mode of action, including induc-

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Contaminated Water Supplies at Camp Lejeune—Assessing Potential Health Effects 94 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 PPARα mode of action for liver cancer in mice is not relevant to humans (NRC 2006). However, others have raised questions about the interpretation of PPARα 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 PPARα 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-β- 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 animals’ 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 γ-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

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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-β-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 γ-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 β-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 α and α1- 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 α1-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-β-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 α 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 α1- and β2-microglobulin, N-acetyl-β-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.

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Contaminated Water Supplies at Camp Lejeune—Assessing Potential Health Effects 96 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 α2µ-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,

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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 µM 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 TCE’s 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, γ-glutamyl transpeptidase, and cysteine conjugate β-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

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Contaminated Water Supplies at Camp Lejeune—Assessing Potential Health Effects 98 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 observed—at low rates—in 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-

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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-2′-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-2′-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-2′-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-

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Contaminated Water Supplies at Camp Lejeune—Assessing Potential Health Effects 100 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 β-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-

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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.

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Contaminated Water Supplies at Camp Lejeune—Assessing Potential Health Effects 124 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

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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-γ 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. PCE’s LOAEL is 386 mg/kg for 78 weeks compared with TCE’s 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 rats’ kidneys. That discrepancy can be attributed largely to the kidneys’ 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-

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Contaminated Water Supplies at Camp Lejeune—Assessing Potential Health Effects 126 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.

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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

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Contaminated Water Supplies at Camp Lejeune—Assessing Potential Health Effects 128 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-

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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 exposure—ingestion, inhalation, and dermal—of 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 µg/L (see Table 2-11). Water samples with detectable PCE from the Tarawa Terrace water supply ranged from 1 to 215 µg/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 µg/L and PCE at 100, 200, and 400 µg/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 (µg/L) × 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 µg/L. For an adult human, the daily dose received from water containing TCE at 1,400 µg/L is estimated to be 1,400 µg/L × 4 L/day = 80 µg/kg per day = 0.08 mg/kg per day. 70 kg Half the highest measured TCE concentration in the water supply (700 µg/L) yields an estimated dose of 0.04 mg/kg per day for adults, and twice the highest measured concentration of TCE (2,800 µg/L) yields

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Contaminated Water Supplies at Camp Lejeune—Assessing Potential Health Effects 130 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 µg/L is estimated to be 1,400 µg/L × 2 L/day = 280 µg/kg per day = 0.3 mg/kg per day. 10 kg Half the highest measured TCE concentration in the water supply (700 µg/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 µg/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:  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.  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.  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 µg/L) in the water supply for an adult human is estimated to be 200 µg/L × 4 L/day = 0.01 mg/kg per day. 70 kg Exposure to half the highest measured water supply concentration (100 µg/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 µg/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 µg/L) is estimated to be 200 µg/L × 2 L/day = 0.04 mg/kg per day. 10 kg

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Review of Toxicologic Studies 131 Exposure to half the highest measured water supply concentration (100 µg/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 µg/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:  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.  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: “Comparison 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.” Other members disagree with Dr. Sheppard’s 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.

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Contaminated Water Supplies at Camp Lejeune—Assessing Potential Health Effects 132 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 [1985]; 52 Fed. Reg. 25690 [1987]; 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 “no known or anticipated adverse effect on health” 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 µg/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 TCE’s 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 µg/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-

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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.