Page 10: Guidelines for Canadian Drinking Water Quality: Guideline Technical Document – Tetrachloroethylene

9.0 Health effects

9.1 Effects in humans

A large number of studies have been conducted to investigate the effects of tetrachloroethylene on human health. Only the most relevant information from these studies is included in this section. An in-depth authoritative review of the effects of tetrachloroethylene on human health was recently performed by U.S. EPA (2012b), which is readily available and can be consulted for further details on individual studies.

9.1.1 Acute toxicity

The discussion in this section focuses primarily on several controlled dosing studies that investigated the acute toxicity of tetrachloroethylene, rather than on case studies of accidental or intentional overexposure that tend to have weak exposure estimates.

Tetrachloroethylene may exert toxicity as an irritant and neurotoxic substance upon acute inhalation exposure. Transient self-reported symptoms, including mild eye and nasal irritation, dizziness, light-headedness, speech difficulties and nausea, were observed in clinical studies of short-term exposure to tetrachloroethylene by inhalation at levels of 100 and 200 ppm (Stewart et al., 1961, 1970). However, these studies did not include controls; studies comparing exposed groups with controls (Stewart et al., 1977) or using non-exposed days in subjects as controls (Stewart et al., 1981) did not find any differences in these symptoms between exposed and control groups. Various neurophysiological and neurobehavioural effects were also observed. The neurophysiological effects included changes in electroencephalogram wave amplitude and frequency at 100 ppm that were consistent with drowsiness and anesthesia (Stewart et al., 1981); changes in peak latencies of visually evoked potentials (Altmann et al., 1990, 1992) and visual contrast sensitivity (Altmann et al., 1990) at 50 ppm; and difficulties in obtaining normal scores in balance tests at 100 ppm (Stewart et al., 1970) and > 200 ppm (Stewart et al., 1961). Decrements in coordination and vigilance tests were observed at > 50 ppm in some acute studies (Stewart et al., 1977, 1981; Altmann et al., 1992), but not all of them (Stewart et al., 1970).

In general, acute studies do not seem to support liver and kidney effects after short-term exposure, except in extreme cases of high-level accidental overexposure. These effects were limited to serum biomarkers of liver damage (Saland, 1967) and mild chemical hepatitis (Stewart, 1969) after extremely high occupational exposures and transient renal failure and tubular necrosis after ingestion (Choi et al., 2003). At lower exposure levels (20-200 ppm), no effects on biomarkers of liver and kidney damage were observed (Stewart et al., 1961, 1970, 1977, 1981).

Mortality from short-term exposures to tetrachloroethylene has been reported in cases of extreme accidental overexposure by inhalation. Although none of the case reports present estimated inhalation levels, concentrations of tetrachloroethylene in various tissues were listed. These tissue levels included 66 mg/L in blood and 31-79 mg/kg in heart, lung and brain in a 2-year-old boy (Garnier et al., 1996). In two adult workers, tissue concentrations were measured at 44 mg/L in blood, 3 mg/kg in lung and 360 mg/kg in brain (Lukaszewski, 1979); and 4.5 mg/L in blood and 69-240 mg/L in brain, kidney and liver (Levine et al., 1981).

9.1.2 Subchronic and chronic toxicity and carcinogenicity

The subchronic and chronic health effects associated with tetrachloroethylene exposure in humans have been investigated primarily in occupational settings and in limited studies of non-occupational exposure through contamination of drinking water or indoor air. In these studies, investigations of non-cancer effects focused primarily on the nervous system, with a few studies also considering hepatic, renal and immunological effects. Associations between tetrachloroethylene exposure and a variety of types of cancer have also been investigated. Neurological effects

Various studies have investigated subjective neurological symptoms and objective measures of neurophysiological, neurobehavioural and neuroendocrine effects of tetrachloroethylene in workers and the general public. The studies of general public exposure examined several cohorts. The first was a cohort of adults who had been previously exposed to tetrachloroethylene in drinking water—both prenatally and until the age of 5 years—when it was leached from the vinyl lining of asbestos-concrete pipes (Getz et al., 2012). Exposed participants in Schreiber et al. (2002) and Storm and Mazor (2004) were adult and child apartment dwellers whose residences were located in buildings containing dry cleaning facilities; the larger cohort from the New York State Department of Health (Getz et al., 2012) and Storm et al. (2011) was gathered in a similar manner, but was composed of different subjects. Another cohort with residential exposures was a group of adults exposed by living in a neighbourhood containing dry cleaning facilities. The majority of occupationally exposed cohorts were composed of dry cleaners (Lauwerys et al., 1983; Seeber, 1989; Cai et al., 1991; Ferroni et al., 1992; Nakatsuka et al., 1992; Cavalleri et al., 1994; Echeverria et al., 1995; Gobba et al., 1998), of which only two contained the same participants (Cavalleri et al., 1994; Gobba et al., 1998). One additional occupationally exposed cohort was a group of workers in a daycare centre that was located in a building containing dry cleaning facilities (Schreiber et al., 2002). Children of dry cleaners were also investigated in one study (Perrin et al., 2007).

Neurophysiological effects observed in the studies involved visual function, with decrements in colour vision and contrast sensitivity being the two effects that were primarily noted. The effects of tetrachloroethylene exposure on colour vision were investigated in the cohort of adults exposed in drinking water prenatally and until 5 years of age (Getz et al., 2012) and in inhalation studies of individuals with current exposures (Nakatsuka et al., 1992; Cavalleri et al., 1994; Gobba et al., 1998; Schreiber et al., 2002; New York State Department of Health, 2010; Storm et al., 2011), with most using the Lanthony desaturated 15 hue panel (a test that requires participants to arrange coloured caps according to colour similarity) for testing. The weight of evidence suggests an association between tetrachloroethylene exposure and decrements in colour discrimination. Significantly higher (i.e., worse) colour confusion index (CCI) scores were observed in adults exposed to tetrachloroethylene in drinking water early in life (at all exposure levels using the Farnsworth desaturated 15 hue panel, but only for estimated exposure concentrations of ≥ 40 µg/L using the Lanthony panel) (Getz et al., 2012). Adults occupationally exposed by inhalation to mean tetrachloroethylene levels of 4.35-7.27 ppm (Cavalleri et al., 1994; Gobba et al., 1998) also had significantly elevated CCI scores. A trend towards higher CCI scores, but without statistical significance, was measured in people living in apartment buildings containing dry cleaning facilities with mean and median daytime exposure concentrations of 1198 and 620 µg/m3, respectively (Schreiber et al., 2002). The majority of the decrements in colour vision that were observed were in the yellow-blue spectrum, which is the spectrum that is expected to be affected by exposure to solvents (Iregren et al., 2002). A dose-response relationship was observed (Cavalleri et al., 1994; Gobba et al., 1998; Getz et al., 2012); furthermore, temporality is suggested by increases in CCI scores in a group of workers whose exposure was measured to be higher (geometric mean level of 4.35 ppm) in a 2-year follow-up (Gobba et al., 1998). Despite these positive observations, no difference in colour vision was observed between exposed and unexposed workers in a study of dry cleaners exposed to geometric mean time-weighted average concentrations of 15.3 ppm (men) and 10.7 ppm (women) (Nakatsuka et al., 1992) or in daycare workers located in a building containing a dry cleaning facility, with a mean exposure concentration of 2150 µg/m3 (Schreiber et al., 2002). These studies, however, might have had a decreased ability to detect changes in colour vision, because the version of the hue panel used in the Nakatsuka (1992) study (the non-desaturated version of the Lanthony 15 hue panel) has a lower ability to detect more subtle decrements in colour vision, and exposure levels in the Schreiber et al. (2002) study were very low for occupational exposures. Despite these minor inconsistencies, the epidemiological weight of evidence tends to indicate that colour vision deficits occur in humans exposed to tetrachloroethylene at occupationally and environmentally relevant exposure levels. These effects might be sustained over time, as was demonstrated in adults whose exposure occurred prenatally and in early childhood (Getz et al., 2012). Moreover, no improvement in CCI score was observed after 2 years in workers whose geometric mean exposure levels dropped from 2.95 to 0.66 ppm (Gobba et al., 1998).

The second neurophysiological effect that was explored for tetrachloroethylene was contrast sensitivity. In adults exposed at a young age to tetrachloroethylene in drinking water, contrast sensitivity was decreased at intermediate and high spatial frequencies, but significance was reached only at the highest frequency and was restricted to the high-exposure group (estimated concentrations of ≥ 40 µg/L) (Getz et al., 2012). Changes in contrast sensitivity were significant at all spatial frequencies in apartment dwellers and daycare operators who were exposed to tetrachloroethylene (at geometric mean concentrations of 1198 and 2150 µg/m3, respectively) from dry cleaning facilities within the building (Schreiber et al., 2002). The analysis of residential exposures, however, included test scores for both adults and children; reanalysis identified no effects on contrast sensitivity when looking solely at adults, and effects on visual contrast sensitivity tests in children might have been affected by diagnosed developmental delay and attention deficit disorder (Storm and Mazor, 2004). A more recent study of apartment dwellers observed a trend of fewer adults and—to an even greater extent—children reaching maximum visual contrast sensitivity scores at multiple spatial frequencies with increasing indoor air exposure (geometric mean concentrations of 11.6 µg/m3 in low exposed adults, 477.9 µg/m3 in high exposed adults, 12.4 µg/m3 in low exposed children and 335.8 µg/m3 in high exposed children) (Storm et al., 2011). The validity of these results, however, is limited by differing characteristics of individuals in the high-exposure category that could affect visual contrast sensitivity scores, including minority status, lower education levels and lower income levels.

Various neurobehavioural and neuropsychological effects were investigated in the cohort of adults who had been exposed to tetrachloroethylene in drinking water during childhood. A neurobehavioural battery was administered to a small subset of the initial study population, which identified significant but inconsistent decrements in memory and attention/executive function (Janulewicz et al., 2012). Exposed subjects had decreased scores in one memory test (delayed recall in the Wechsler memory test, but not in others); furthermore, impacts on attention were identified in only one part of the Trail Making Test, but these effects lost significance when adjusted for sex and education (Janulewicz et al., 2012). Due to a limited number of participants in the neurobehavioural battery, comparisons were made solely between exposed and unexposed participants (and no average exposure estimates were provided); however, a greater sample size in other studies within this cohort allowed for division of participants by exposure subgroups to identify trends in effects. Slight increases in a tendency towards risky behaviour (likelihood for regular smoking, drinking or use of various drugs in adolescence or recently) were observed, particularly in the highest tertile of exposure (estimated cumulative prenatal plus postnatal exposures of ≥ 77.6 g) (Aschengrau et al., 2011). An increase in self-reported behavioural and learning disorders did not demonstrate a dose-response relationship, and borderline significance observed in the low-exposure group (estimated cumulative exposures of < 10 g prenatally and < 66.7 g postnatally, corresponding to 40 µg/L) was lost when adjustment for maternal age, race, education, sex, prematurity and low birth weight was performed (Janulewicz et al., 2008). Weak associations were observed between exposure and self-reported mental illness, with a significantly increased risk of bipolar disorder in the highest tertile (but with no dose-response relationship, as a higher risk was observed in the lowest tertile than in the middle tertile), an increase of borderline significance in post-traumatic stress disorder in the highest tertile that was lost when adjusted, and a non-significant increase in schizophrenia that could not be analyzed by tertile due to the low incidence of the disease (Aschengrau et al., 2012). No exposure estimates were provided for any of the tertiles. Therefore, the Aschengrau et al. (2012) findings were not sufficient to confirm those of a preliminary study of schizophrenia in a study with a large number of participants in Israel, which found a significantly increased risk of being diagnosed with schizophrenia in subjects whose parent's occupation was dry cleaning (as identified on subjects' birth certificates; therefore, no exposure levels were provided) (Perrin et al., 2007).

Table 2. Neurobehavioural test results for populations exposed via inhalation
Neurobehavioural function Test Results
Attention, vigilance and executive function Continuous performance test
  • Significantly prolonged response (Altmann et al., 1995)
Simple reaction time
  • Significantly prolonged response (Altmann et al., 1995)
  • Significantly prolonged response (Ferroni et al., 1992)
  • No effect (Lauwerys et al., 1983)
Trail making test
  • No effect (Echeverria et al., 1995)
Shape comparison
  • Significantly prolonged response (Ferroni et al., 1992)
Choice reaction
  • Significantly lower score (Seeber, 1989)
  • No effect (Lauwerys et al., 1983)
Perceptual functions (digit recognition)
  • Significantly prolonged response (Seeber, 1989)
Attention and short-term memory Visual retention test
  • Significantly fewer correctly identified (Altmann et al., 1995)
Visual reproduction test
  • Significantly lower score with lifetime exposure, but not 3-year or current exposure (Echeverria et al., 1995)
Pattern memory test
  • Significantly lower score with lifetime and 3-year exposure, but not current exposure; prolonged response time with lifetime and current exposure (Echeverria et al., 1995)
Pattern recognition test
  • Significantly lower score with lifetime exposure, but not 3-year or current exposure (Echeverria et al., 1995)
Digit span test
  • No effect (Echeverria et al., 1995)
  • Significantly lower score (with or without cancellation task) (Seeber, 1989)
Visuospatial and visuomotor Finger tapping test
  • No effect on right or left; difference for alternating, but not significant (p < 0.1 but not p < 0.05) (Altmann et al., 1995)
Visual-motor tracking
  • No effect (Altmann et al., 1995)
Digit symbol test
  • No effect (Echeverria et al., 1995)
  • No effect (Ferroni et al., 1992)
  • Significantly lower score (Seeber, 1989)

Neurobehavioural effects were also investigated in populations exposed through inhalation, from either residential (Altmann et al., 1995) or occupational (Lauwerys et al., 1983; Seeber, 1989; Ferroni et al., 1992; Echeverria et al., 1995) exposures. Detailed results from specific tests can be found in Table 2. Differences between exposed and unexposed groups were observed at a median concentration of 1.36 mg/m3 (Altmann et al., 1995) in those residentially exposed. Higher levels were required for effects to be observed in occupational studies. Responses were observed at median levels of 15 ppm (Ferroni et al., 1992) and at average concentrations of 83.4 and 363.8 mg/m3 (Seeber, 1989), but dose-response relationships were not observed in either study. In the study by Echeverria et al. (1995), effects were observed—typically with a dose-response relationship—in the lifetime or 3-year cumulative exposure groups (exposure quantification not published for these groups), but not at current exposures of up to 41.8 ppm. No effects were observed at a mean level of 20.8 ppm in an additional study (Lauwerys et al., 1983). The weight of evidence from these epidemiological studies suggests that tetrachloroethylene exposure at elevated levels might be associated with adverse effects on attention, vigilance and executive function, as well as on attention and short-term memory function, but not on visuomotor or visuospatial function.

Three studies considered the effects of chronic tetrachloroethylene exposure on subjective neurological symptoms by administering questionnaires that focused on symptoms that could possibly be related to disturbances of the nervous system (e.g., dizziness and fainting, drunkenness, forgetfulness, loss of concentration, increased excitability). Significantly higher prevalence of self-reported neurological symptoms was observed in two of these studies in workers exposed by inhalation to a geometric mean level of 19.9 ppm (Cai et al., 1991) and average concentrations of 83.4 and 363.8 mg/m3 (Seeber, 1989), but no dose-response relationships were observed. Although the prevalence of neurological signs was increased in workers exposed to a mean level of 20.8 ppm in the third study (Lauwerys et al., 1983), these increases were not significant. Furthermore, in the one study in which a clinical examination was administered by a physician, no differences were observed between exposed (geometric mean level of 19.9 ppm) and unexposed workers (Cai et al., 1991).

In an attempt to identify neuroendocrine effects of tetrachloroethylene, Ferroni et al. (1992) measured prolactin levels in female dry cleaning workers. Basal serum prolactin levels were significantly elevated during the proliferative phase of the menstrual cycle in tetrachloroethylene-exposed women (median level of 15 ppm); however, no dose-response relationship was observed. Other phases of the menstrual cycle were not discussed. The authors postulated that this effect might suggest that pituitary function is affected by tetrachloroethylene.

Of the endpoints discussed in this section, two studies (Cavalleri et al., 1994; Echeverria et al., 1995) were considered to be most relevant for the risk assessment. The reason for this decision is outlined in Section 9.5.2. A more in-depth description of these two studies is included below.

The first study (Cavalleri et al., 1994) investigated the effect of occupational exposure on colour vision by dividing 35 workers from 12 dry cleaning shops around Modena, Italy, into high-exposure (dry cleaners; mean 8-hour time-weighted average exposure level of 7.27 ppm) and moderate-exposure (ironers; mean 8-hour time-weighted average exposure level of 4.8 ppm) groups. Although breathing zone air samples were obtained for all participants, exposure was measured only during one workshift (which was considered to be representative of other days). Exposed workers had been exposed to tetrachloroethylene for an average of 8.8 years at the time of testing. The same number of controls (matched to exposed subjects based on sex, age, and alcohol and cigarette consumption) was gathered from workers unexposed occupationally or vocationally to solvents or other substances toxic to eyes. Exposed and control participants were excluded if they consumed more than an average alcohol intake of 50 g/day, smoked more than 30 cigarettes/day or had visual acuity of less than 6/10 with lenses. Testing was performed under standard conditions using the Lanthony desaturated 15 hue panel by a technician who was blinded to the exposure level and workers' tasks. Only three workers received a perfect score on the test, which was significantly less than the 13 controls that performed the test perfectly. Mean CCI score was significantly higher (worse) than controls in dry cleaners, but not in ironers; therefore, the no-observed-adverse-effect concentration (NOAEL) for the workers was 4.8 ppm. A multivariate analysis confirmed that the effect was most likely due to tetrachloroethylene and not other characteristics that can affect CCI score (e.g., age, alcohol, smoking). One weakness of the study was that some overlap in exposure occurred between the two exposure groups (range of exposures: ironers, 0.52-11.28 ppm; dry cleaners, 0.38-31.19 ppm); however, the authors also found a significant correlation between workers' time-weighted average exposure levels and their CCI scores, further supporting a dose-response relationship. An additional weakness with the use of the study in a dose-response assessment is that the significant association between tetrachloroethylene and CCI scores is lost upon exclusion of the three workers with the highest exposure levels (> 12.5 ppm) (Cavalleri et al., 1994; Benignus et al., 2009).

The second study that was considered further in the risk assessment was of 65 dry cleaners with no previous central nervous system problems who had worked in 1 of 23 shops in the Detroit, Michigan, area for at least 1 year (Echeverria et al., 1995). Workers were assigned to low-exposure (n = 24), moderate-exposure (n = 18) or high-exposure (n = 23) groups based on an industrial hygiene walkthrough. Breath samples were obtained from participants, and 15-minute breathing zone samples were obtained from one low-, medium- and high-exposed participant in 19 of the shops; the air sampling results identified average exposure levels of 0.6, 12.1 and 41.8 ppm, respectively, in wet transfer shops and of 0, 4.3 and 11.4 ppm, respectively, in dry-to-dry shops. In addition to current exposures, chronic and 3-year indices of exposure were estimated by summing the product of duration of employment and job title over 3 years or a lifetime. Controls were considered to be participants in the low-exposure group and were matched with highly exposed participants on education and age. Neurobehavioural testing of participants was performed on the 1st or 2nd day of the work week using a battery of tests (Profile of Mood States, Digit Span, Visual Reproductions Subtest, Wechsler Memory Scale, Neurological Evaluation Scale [NES] Pattern Memory, NES Pattern Recognition, Trail Making Test, NES Symbol-Digit Substitution Test and a vocabulary test). Regression models considered the potential effect of other variables on test results, including age, years of education, verbal skill, frequency of alcohol consumption, hours of sleep, fatigue, mood, symptoms, medication, and current or previous occupational or avocational exposure to neurotoxicants. Associations between tetrachloroethylene exposures and tests were found for the Visual Reproductions Subtest (lifetime index), Pattern Memory (lifetime and 3-year indices) and NES Pattern Recognition (lifetime index). No associations with tetrachloroethylene exposure were found for the other tests, and no associations were identified on any of the tests for current tetrachloroethylene exposures. Other non-cancer effects

Studies in dry cleaners have suggested that chronic inhalation exposure to tetrachloroethylene may be associated with damage to the liver, as demonstrated by elevated serum gamma-glutamyltransferase activity (Gennari et al., 1992; Brodkin et al., 1995) and alterations in the liver parenchyma observed using ultrasonography (Brodkin et al., 1995). However, other studies in similar populations did not demonstrate changes in levels of enzymes potentially indicative of liver damage (Lauwerys et al., 1983; Cai et al., 1991).

Adverse effects of tetrachloroethylene on kidney tubules are also suggested by occupational epidemiology studies. Elevated levels of urinary retinol binding protein (RBP) were observed in two studies (Mutti et al., 1992; Verplanke et al., 1999); although the levels were within normal ranges in the Verplanke et al. (1999) study, a significant increase in the incidence of abnormal RBP levels was reported by Mutti et al. (1992). No effect on RBP level was observed in one study with a smaller sample size (Lauwerys et al., 1983). Data to support the potential for tubular effects come from observations of proteinuria in occupationally exposed populations. Decreased integrity of the proximal tubule brush border was suggested by increases in brush border antigens and tissue non-specific alkaline phosphatase (ALP) (Mutti et al., 1992). Elevated urinary concentrations of β2μ-globulin (Mutti et al., 1992) and lysozyme (Franchini et al., 1983; Vyskocil et al., 1990) in dry cleaners were suggestive of decreased reabsorption in renal tubules; it should be noted, however, that two other studies—albeit with smaller sample sizes—did not note significant increases in urinary β2μ-globulin (Lauwerys et al., 1983; Vyskocil et al., 1990). β2-glucuronidase levels were also significantly increased in female dry cleaning shop workers (Franchini et al., 1983). No effects were observed on urinary levels of N-acetyl-β-D-glucosaminidase (Solet and Robins, 1991; Mutti et al., 1992; Verplanke et al., 1999; Trevisan et al., 2000) or alanine aminopeptidase (Verplanke et al., 1999).

The limited data available suggest that immunological and hematological effects are possible following tetrachloroethylene exposure, as noted by alterations in serum immunoglobulin E and circulating leukocytes, in addition to many other markers (Andrys et al., 1997; Lehmann et al., 2001, 2002; Emara et al., 2010). Studies of respiratory effects (Delfino et al., 2003a, 2003b; Tanios et al., 2004) and connective tissue effects (e.g., scleroderma) (Hinnen et al., 1995; Goldman, 1996; Lacey et al., 1999; Garabrant et al., 2003; Maître et al., 2004) are limited and inconsistent; thus, no firm conclusions can be drawn for these endpoints. Cancer

The potential for tetrachloroethylene to cause cancer has been examined mostly in occupational studies within the dry cleaning industry, but also in studies pertaining to drinking water. Studies of exposure to tetrachloroethylene via leaching from vinyl linings in pipes have revealed significant increases in leukemia in highly exposed individuals (i.e., participants with exposure levels ≥ 90th percentile) (Aschengrau et al., 1993). Several analyses of this population also suggest that exposure to tetrachloroethylene may be associated with breast cancer in highly exposed women (Aschengrau et al., 1998, 2003; Vieira et al., 2005; Gallagher et al., 2011), but observed increases were not significant in most of the studies and only of borderline significance in one of the studies (Gallagher et al., 2011). Other cancers examined were not associated with tetrachloroethylene exposure, including bladder, kidney and colon—rectum cancer (Aschengrau et al., 1993; Paulu et al., 1999). A study in a different population revealed tetrachloroethylene-related increases in high-grade non-Hodgkin's lymphoma with or without Burkitt's lymphoma among females exposed to concentrations above 5 ppb (5 µg/L) in drinking water (Cohn et al., 1994); however, the relative influence of trichloroethylene on these results is unclear due to the high degree of correlation between both chemicals in the study population.

Other community-based studies of contaminated water (containing many chemicals, including tetrachloroethylene) have noted increases in bladder cancer (Mallin, 1990) and mortality from liver cancer (Lee et al., 2003), whereas another reports no effects on liver cancer, non-Hodgkin's lymphoma, Hodgkin's disease, multiple myeloma or leukemia (Vartianinen et al., 1993). Other studies have noted a positive association between childhood leukemia and access to water potentially contaminated with chlorinated compounds, including tetrachloroethylene (Cutler et al., 1986; Lagakos et al., 1986); however, when likelihood of exposure to the contaminated wells was considered, no increase in risk was observed (Costas et al., 2002). The aforementioned studies involved co-exposure to various other chemicals, including trichloroethylene; thus, it is not possible to decipher whether these effects were caused directly by tetrachloroethylene exposure or by exposure to other chemicals.

As individuals working in the dry cleaning industry are primarily exposed to tetrachloroethylene, studies of dry cleaners can provide useful information on the carcinogenicity of tetrachloroethylene. These studies have noted increases in the incidence of or mortality from lung cancer (Brownson et al., 1993; Calvert et al., 2011), bladder cancer (Brown et al., 1987), tongue cancer (Calvert et al., 2011), oesophagus cancer (Calvert et al., 2011) and renal cell carcinoma (McCredie and Stewart, 1993; Mandel et al., 1995) in dry cleaners or in individuals classified as being exposed to dry cleaning solvents. Non-significant effects for laryngeal and oesophageal cancers (Vaughan et al., 1997) have also been reported. Other studies—based primarily on census data, cancer registries and other similar databases—that grouped dry cleaners into broader occupational categories have reported increases in the incidence of lung cancer (Andersen et al., 1999; Pohlabeln et al., 2000; Travier et al., 2002; Ji and Hemminki, 2005; Pukkala et al., 2009; Seldén and Ahlborg, 2011), liver cancer (Lynge and Thygesen, 1990; Pukkala et al., 2009), breast cancer (Band et al., 2000), kidney cancer (Ji et al., 2005), upper aerodigestive and pharynx cancer (Ji and Hemminki, 2005), stomach cancer (Pukkala et al., 2009), cervix uteri cancer (Pukkala et al., 2009), oral cancer (Pukkala et al., 2009), leukaemia (Morton and Marjanovic, 1984; Travier et al., 2002), non-Hodgkin's lymphoma (Seldén and Ahlborg, 2011) and Hodgkin's disease (Andersen et al., 1999). However, similar studies have also reported no excess risk of liver and kidney cancer (Lynge et al., 1995) and non-Hodgkin's lymphoma (Cano and Pollán, 2001).

Although no significant excess risk of bladder cancer was observed in case-control studies including dry cleaning workers (Smith et al., 1985; Lynge et al., 1995; Zheng et al., 2002), an association between tetrachloroethylene exposure and bladder cancer has been considered consistent because of observations from three large cohorts of dry cleaning workers with several follow-up periods (Guha et al., 2012). A significant increase in the incidence of bladder tumours was observed in Nordic dry cleaners (44% increase), with a higher increase (69%) in workers from Denmark and Norway, the countries with better job classification data in the study (Lynge et al., 2006). Interviews with these participants or their family members identified that the greatest excesses were observed in support staff in small dry cleaning shops (120%) and owners of combined laundry and dry cleaning shops (92%). A significant increase (of 159%) in mortality from bladder and other urinary cancer was observed in workers from four U.S. cities (San Francisco/Oakland, California; Chicago, Illinois; Detroit, Michigan; and New York City, New York), but only in subjects who had worked in multiple shops, some of which had predominant use of solvents other than tetrachloroethylene (Calvert et al., 2011). In the third cohort, composed of dry cleaners from St. Louis, Missouri, only slight increases (of 30%) in mortality from bladder tumours were observed over the entire follow-up period, and these increases were not significant (Blair et al., 2003). No cases of bladder cancer were observed in the smaller, predominantly tetrachloroethylene-exposed cohort. Risk of bladder tumours did not increase with longer exposure periods in one of the studies (Lynge et al., 2006) and actually appeared to decrease with duration of exposure in another (Blair et al., 2003), but the majority of cases were observed in workers exposed for longer than 5 years in the third cohort (Calvert et al., 2011). Other occupational studies encompassing exposure to various substances but investigating the specific effects of tetrachloroethylene exposure have reported no effects on renal cell carcinoma (Dosemeci et al., 1999), lymphoma (Mester et al., 2006; Miligi et al., 2006; Seidler et al., 2007) or cancer in general (Anttila et al., 1995), as well as no increase in childhood leukemia from maternal exposure (Shu et al., 1999; Infante-Rivard et al., 2005). However, effects on multiple myeloma (Spirtas et al., 1991) and renal cell carcinoma (Schlehofer et al., 1995) were noted.

Thus, the current epidemiological data suggest that tetrachloroethylene may be involved in carcinogenic outcomes. However, the number of studies that also demonstrate a lack of association between tetrachloroethylene exposure and cancer, coupled with limitations such as exposure to other potentially carcinogenic solvents and absence of exposure analysis for tetrachloroethylene, make it difficult to draw firm conclusions. As cancers were inconsistently observed at various sites, it is difficult to draw any conclusions on the carcinogenicity of tetrachloroethylene in humans based on these studies. The cancer endpoint that has been considered to be the most consistent across studies is bladder cancer (Guha et al., 2012); however, the lack of thorough exposure assessments in these studies combined with an absence of bladder tumours in experimental animal studies prevents the inclusion of this tumour type in the dose-response assessment.

9.1.3 Reproductive and developmental toxicity

The potential effects of occupational exposure to tetrachloroethylene on reproductive and developmental toxicity have been examined in a number of case-control and cohort studies. Overall, an increased risk of spontaneous abortion was reported for some of the studies (Hemminki et al., 1980; Kyyronen et al., 1989; Olsen et al., 1990; Windham et al., 1991; Doyle et al., 1997), although others have reported negative findings for tetrachloroethylene-induced pregnancy loss (Ahlborg, 1990; Lindbohm et al., 1990; Eskenazi et al., 1991; Aschengrau et al., 2009a), including one study that investigated paternal exposure (Eskenazi et al., 1991). Exposure to tetrachloroethylene via soil vapours was not associated with lower birth weight or fetal growth restrictions, but a non-significant association related to cardiac defects was noted (Forand et al., 2012).

A limited number of community-based studies have evaluated the potential reproductive and developmental toxicity of tetrachloroethylene through exposure from drinking water. In an area of Massachusetts affected by leaching of tetrachloroethylene from vinyl-lined pipes, pregnancy outcomes were evaluated in a cohort of pregnant women exposed to concentrations ranging from 1.5 to 80 μg/L (most highly contaminated areas were flushed until levels were below 40 μg/L) in comparison with an unexposed control group. These investigations revealed non-significant associations between exposure to tetrachloroethylene and congenital anomalies such as neural tube defects and oral clefts (Aschengrau et al., 2009b). Moreover, no associations between tetrachloroethylene exposure and birth weight (Aschengrau et al., 2008), gestational duration (Aschengrau et al., 2008) or pregnancy loss (Aschengrau et al., 2009a) were observed. Several factors limited the findings from these studies, including the fact that tetrachloroethylene concentrations in water were estimated using a leaching transport model and that details related to water consumption and bathing habits could not be included because of poor recall; moreover, there is a possibility of underreporting of the pregnancy losses. One study in a population in northern New Jersey exposed to various chemicals through drinking water noted increases in oral cleft defects when tetrachloroethylene levels were above 10 ppb (10 µg/L) (Bove et al., 1995); however, the relevance of this is unknown, since it is not possible to elucidate which of the chemicals could have resulted in the effect. Many other chemicals were considered in this study, including trichloroethylene, 1,1,1-trichloroethane, trans-1,2-dichloroethylene, 1,1-dichloroethylene, carbon tetrachloride, 1,2-dichloroethane, benzene, dichloromethane, vinyl chloride, chlorobenzene, dichlorobenzenes, trichlorobenzenes, chlordane and polychlorinated biphenyls.

Other epidemiological studies related to neurodevelopmental endpoints, including those that investigated the effects of tetrachloroethylene in populations exposed via drinking water in utero and until 5 years of age (Janulewicz et al., 2008, 2012; Aschengrau et al., 2011, 2012; Getz et al., 2012) and in children exposed from indoor air in apartment buildings containing a dry cleaning facility (Schreiber et al., 2002; Storm and Mazor, 2004; New York State Department of Health, 2010; Storm et al., 2011), are summarized in Section That section also discusses a study investigating the effects of tetrachloroethylene on prolactin levels in occupationally exposed women (Ferroni et al., 1992), which could have reproductive repercussions.

9.2 Effects on experimental animals

A large number of studies have investigated the effects of tetrachloroethylene on the health of animal models. Only the most relevant information from these studies is discussed in the following sections.

9.2.1 Acute toxicity

Tetrachloroethylene is of low acute oral toxicity. Oral median lethal doses (LD50 values) in rats range from 2.4 to 13.0 g/kg bw (Pozzani et al., 1959; Smyth et al., 1969; Withey and Hall, 1975; Hayes et al., 1986), and values for mice range from 6.4 to 9.6 g/kg bw (Dybing and Dybing, 1946; Wenzel and Gibson, 1951; Bandman, 1990). Tetrachloroethylene-related deaths occurred at most exposure levels in rats exposed to 2445-5163 ppm and mice exposed to 2328-3786 ppm for 4 hours via inhalation, with 100% mortality in the 5163 ppm rat group and 2971 and 3786 ppm mouse groups (NTP, 1986). Symptoms preceding death after high exposures included tremors, ataxia and central nervous system depression in rats (Hayes et al., 1986) and hypoactivity and anaesthesia in mice (NTP, 1986).

The nervous system was the most commonly researched organ system in acute toxicity studies with tetrachloroethylene. Neurobehavioural effects were the neurological effects observed at the lowest concentrations in oral and inhalation studies, with the lowest gavage effect (anticonvulsion) occurring at ≥ 50 mg/kg bw (Chen et al., 2002). Changes in motor activity occurred at ≥ 150 mg/kg bw by ingestion and at ≥ 90 ppm by inhalation (Kjellstrand et al., 1985; Moser et al., 1995), with a tendency for increased arousal at lower doses and lower activity at higher doses in one of the studies (Moser et al., 1995). The circadian rhythm of motor activity was also increased in rats intraperitoneally injected with higher doses (Motohashi et al., 1993). Other neurobehavioural effects observed at higher doses included decreased righting ability (Moser et al., 1995); abnormal gait, increased landing foot splay and inability to grasp grip strength devices (Moser et al., 1995); decreased equilibrium and coordination (Umezu et al., 1997); and suppressed learned behaviour (Warren et al., 1996). Of the non-neurobehavioural effects, autonomic effects—including altered pupil response, lacrimation and salivation—were those observed at gavage doses of ≥ 150 mg/kg bw (Moser et al., 1995). At the lowest inhalation exposure level of 200 ppm (6 hours/day for 4 days), neurochemical measures of toxicity, including reduction in brain ribonucleic acid (RNA) and glutathione content and increased cholinesterase activity, were observed (Savolainen et al., 1977). Other neurological effects observed at higher doses included changes to the shape and amplitude of waveforms for visually evoked potentials, cerebellar flash-evoked potentials and somatosensory evoked potentials (Mattsson et al., 1998), electroencephalogram-confirmed anaesthesia (Mattsson et al., 1998) and decreased sensory perception (Chen et al., 2002).

Only a few studies investigated acute effects on other organ systems after tetrachloroethylene exposure. Renal effects observed after intraperitoneal injection included hydropic degeneration and necrosis in tubules of mice (Klaassen and Plaa, 1966) and calcification in tubules in dogs (Klaassen and Plaa, 1967), with the lowest median effective dose (ED50) of 14 mmol/kg bw, equivalent to 2322 mg/kg bw, as measured by phenolsulphonphthalein retention, in the dogs. The same studies reported hepatic effects of vacuolation of hepatocytes in both mice and dogs and liver and hepatocyte enlargement, cellular infiltration and slight necrosis in mice; again, the dogs were more sensitive, with ED50- values of 7.2 mmol/kg bw (1194 mg/kg bw), as measured by serum alanine transaminase (ALT) levels. Immunotoxicity was also observed, which manifested as decreased bactericidal activity and increased mortality from streptococcal pneumonia in mice exposed by inhalation to 50 ppm for 3 hours (Aranyi et al., 1986) and as enhanced allergic reactions (measured by increases in passive cutaneous anaphylaxis after antigen challenge) in mice exposed to ≥ 0.01 mg/kg bw via intraperitoneal injection (Seo et al., 2012). Cardiotoxic effects observed in rabbits, cats and dogs injected with tetrachloroethylene included ventricular arrhythmias, increased myocardial vulnerability to premature ventricular contractions, bigeminy and tachycardia; rabbits had the lowest threshold for effects, which was 10 mg/kg bw (Kobayashi et al., 1982).

9.2.2 Short-term exposure

Many studies investigated the effects of tetrachloroethylene on various organ systems over subacute and subchronic durations. Exposures in these studies were mainly via oral gavage and inhalation, with a few studies of intravenous or intraperitoneal injection. No studies exposed experimental animals to tetrachloroethylene by drinking water. Clinical and serological indicators of health

Adverse clinical outcomes were observed after high subacute and subchronic exposures to tetrachloroethylene. Increased mortality in rats and mice occurred at 1000 mg/kg bw in a gavage study (Philip et al., 2007) and at ≥ 1600 ppm in inhalation studies (NTP, 1986). Decreases in body weight gain were consistently observed in both sexes of various species of mice and rats at doses of ≥ 5 mg/kg bw via gavage (Hayes et al., 1986; Chen et al., 2002) and at inhalation exposures of ≥ 75 ppm (Kjellstrand et al., 1984; NTP, 1986; JISA, 1993; Wang et al., 1993). Hematology and clinical chemistry were not explored in most subacute or subchronic studies, but exposure of mice to 3000 mg/kg bw per day by gavage for 15 days resulted in lower levels of hemoglobin, red blood cells, hematocrit (Ebrahim et al., 2001), blood glucose and gluconeogenic enzymes and increases in hexokinase, aldolase and phosphoglucoisomerase enzymes (Ebrahim et al., 1996); these parameters returned to normal with antioxidant administration. Hematological and clinical chemistry changes were also reported in mice and rats after inhalation of ≥ 609 ppm, but further details on the effects were not provided by the authors (JISA, 1993). Neurological effects

Neurological effects were investigated in several subacute and subchronic studies of exposure to tetrachloroethylene. Animals were exposed via gavage in oil, by inhalation and by intraperitoneal injection.

Whole brain weight (absolute and/or relative) was decreased in female (but not male) rats gavaged with 1400 mg/kg bw per day (Hayes et al., 1986) and in rats and gerbils exposed to 320 ppm via inhalation (Kyrklund et al., 1988), with no effects observed at lower levels (Hayes et al., 1986). At inhalation concentrations of ≥ 320 ppm, the weights of certain brain sections (frontal cerebral cortex and brain stem) were also decreased in gerbils and rats (Rosengren et al., 1986).

Studies also looked at the impact of tetrachloroethylene on brain biochemistry—including protein, amino acid, deoxyribonucleic acid (DNA), phospholipid, cholesterol and fatty acid content, and levels of neurotransmitters and cell marker proteins—in rodents exposed by inhalation only. Of these effects, altered brain amino acid content was observed at the lowest level, with effects on glutamine and taurine levels observed at ≥ 120 ppm in gerbils and rats (Briving et al., 1986). Other effects were observed at higher exposure levels. Decreased DNA and protein concentrations were found in specific brain sections (cerebral cortex, cerebellum and brain stem) of rats and gerbils (Rosengren et al., 1986; Karlsson et al., 1987). Cholesterol and phospholipid levels, as well as a change in their ratio, tended to be decreased in gerbils and male rats (Kyrklund et al., 1987, 1990), but not at tetrachloroethylene levels as low as 120 ppm (Kyrklund et al., 1984). Fatty acid patterns in the cerebral cortex and hippocampus were also changed in gerbils and rats, with a tendency towards elongation and desaturation of fatty acids, resulting in greater levels of very long chain polyunsaturated fatty acids (Kyrklund et al., 1987). Results for cell marker proteins were variable, with increases in S-100 in several brain lobes (hippocampus, anterior and posterior cerebellum, and occipital cerebral cortex) and a decrease in the frontal cerebral cortex in gerbils (Rosengren et al., 1986), but decreases in S-100 and glial fibrillary acid in the cerebral cortex, hippocampus and brain stem of rats (Wang et al., 1993). Finally, a significant decrease in acetylcholine level and a decrease in dopamine level that was dose related but not significant at any individual tetrachloroethylene doses were also observed in male rats exposed by inhalation (Honma et al., 1980).

The only subacute or subchronic study to investigate neurophysiological effects noted changes to midlatency components of visual cortex flash-evoked potentials in rats exposed by inhalation to 800 ppm (Mattsson et al., 1998).

The neurobehavioural effects that were observed at the lowest level in gavage studies were antinociception and anticonvulsive effects (for myoclonic twitch and forearm clonus) in male rats exposed to ≥ 5 mg/kg bw per day by gavage for 8 weeks (Chen et al., 2002). Although the adversity of these effects is unclear, the high dose in this study (50 mg/kg bw per day) demonstrated more clearly undesirable outcomes, with a lower level of motor activity and rearing. The most sensitive effect in inhalation studies was also related to motor activity, but with an increase in motor activity over controls in mice exposed to 225 ppm for 16 hours/day (and other higher exposure levels, but for shorter time periods, resulting in the same 24-hour time-weighted average concentrationFootnote 2) for 30 days (Kjellstrand et al., 1984). Because lower activity was observed at lower doses and higher activity was observed at higher doses,Footnote 3 the effect on activity was opposite to that observed in acute studies. The circadian rhythm of motor activity was also increased at higher doses in the Kjellstrand et al. (1984) study. Some evidence of weak neuromuscular effects was demonstrated in rats, as significant increases were observed only in an overall neuromuscular domain score and not in the results of individual tests (for gait score, righting reflex, forelimb and hindlimb grip strength, and landing foot splay) (Moser et al., 1995), and grip strength was unchanged in another study (Mattsson et al., 1998).

Neurological effects also occurred in rodents that were exposed for short durations as developing fetuses or pups. Mice exposed to tetrachloroethylene at oral gavage doses as low as 5 mg/kg bw per day for 7 days, as of postnatal day 10, exhibited hyperactive behaviour later in development (i.e., at 60 days of age, but not 17 days) (Fredriksson et al., 1993). Altered behaviour was also reported in rats exposed in utero (Nelson et al., 1980; Szakmáry et al., 1997). Moreover, a study with mice exposed to tetrachloroethylene, chloroform and bromoform via drinking water throughout pregnancy, weaning and development noted behavioural changes that are relevant to autism in male (but not female) offspring (Guariglia et al., 2011). As a result of in utero exposure, no changes in brain fatty acids were observed (Kyrklund and Haglid, 1991); however, decreases in acetylcholine and dopamine levels have been reported (Nelson et al., 1980). These studies suggest that tetrachloroethylene may target the nervous system, leading to altered behaviour later in development. Hepatic effects

Increases in absolute and relative liver weights were observed in many studies, with more consistent effects in mice, particularly at lower doses. Exposures of mice resulted in consistent increases at doses as low as 100 mg/kg bw per day in oral gavage studies (Schumann et al., 1980; Ebrahim et al., 1996) and 9 ppm by inhalation (Kjellstrand et al., 1985; Odum et al., 1988; Kyrklund et al., 1990). In rats, liver weight increases were observed at ≥ 1000 mg/kg bw per day by gavage in two studies (Hayes et al., 1986; Goldsworthy and Popp, 1987), but another study did not demonstrate effects at 1000 mg/kg bw per day (Schumann et al., 1980); similarly, inhalation studies demonstrated effects at 320 ppm (Kyrklund et al., 1988), but not at 200 or 400 ppm (Odum et al., 1988). Liver weight effects were somewhat, but not completely, reversible after cessation of exposure (Kjellstrand et al., 1984; Kyrklund et al., 1990).

Biochemical indicators of hepatic effects were also observed in several studies. The most sensitive effects were increases in liver triglycerides at ≥ 100 mg/kg bw per day by gavage (Buben and O'Flaherty, 1985) and a reversible increase in plasma butyrylcholinesterase activity in mice exposed to ≥ 37 ppm by inhalation (Kjellstrand et al., 1984). Other effects at higher levels included an increase in enzymes associated with glycolysis and a decrease in those associated with gluconeogenesis (Ebrahim et al., 1996) and increases in ALT (referred to in older studies as serum glutamate-pyruvate transaminase) (Philip et al., 2007) and glucose-6-phosphatase activity (Buben and O'Flaherty, 1985).

Various morphological and histological changes were observed in animals exposed to tetrachloroethylene via gavage or inhalation for up to 90 days. The lowest effect in animals exposed via gavage was hepatocyte hypertrophy at ≥ 100 mg/kg bw per day (Buben and O'Flaherty, 1985), which was also observed in an inhalation study (Odum et al., 1988). In one inhalation study with the lowest exposure levels (≥ 9 ppm continuous exposure), liver discolouration, hepatocyte hypertrophy, vacuolization and accumulation of inflammatory cells were observedFootnote 4 (Kjellstrand et al., 1984). These effects were also observed in other studies, albeit at higher exposure levels (liver discoloration: Buben and O'Flaherty, 1985; JISA, 1993; Philip et al., 2007; vacuolization and infiltration of inflammatory and immune cells: NTP, 1986; Odum et al., 1988). Effects observed at higher levels included hepatocyte degeneration (Buben and O'Flaherty, 1985; Ebrahim et al., 1996; Philip et al., 2007) and necrosis (Ebrahim et al., 1996); central enlargement (JISA, 1993); peroxisome proliferation, lipid accumulation and changes in mitochondria levels (Odum et al., 1988); clumped cytoplasm and karyorrhexis (disintegrating nucleus) (Buben and O'Flaherty, 1985); and bile stasis (NTP, 1986). Of the aforementioned effects, only hepatocyte hypertrophy was observed in rats, and it occurred at higher doses than in mice. Additional effects that occurred in rats included hepatic congestion (NTP, 1986), hepatic foci (Story et al., 1986) and altered staining affinity of hepatocytes (Schumann et al., 1980). Renal effects

Tetrachloroethylene increased absolute and relative kidney weights in some gavage studies at ≥ 400 mg/kg bw per day, with male rats appearing more sensitive than female rats and with effects occurring at lower doses in rats than in mice (Hayes et al., 1986; Green et al., 1990; Ebrahim et al., 1996; Jonker et al., 1996). Conversely, no increase in relative or absolute kidney weight was observed in another gavage study (Goldsworthy and Popp, 1987) or in any inhalation studies (Odum et al., 1988; Green et al., 1990).

Biochemical changes related to kidney effects have been investigated in gavage studies only, and primarily in rats, with the most sensitive effect of proteinuria (increases in urinary excretion of albumin, RBP and α2-microglobulin and transient increase in N-acetylglutamate [NAG]) at ≥ 500 mg/kg bw per day (Bergamaschi et al., 1992; Jonker et al., 1996). At higher doses, increased urinary volume and alterations in levels of urinary ALP, NAG, glucose, gamma-glutamyl transpeptidase, lactate dehydrogenase (LDH), aspartate transaminase (AST) and ALT (Green et al., 1990; Jonker et al., 1996) were observed. In mice, an increase in renal glycolytic enzymes and decrease in renal gluconeogenic enzymes were observed (Ebrahim et al., 1996), but tetrachloroethylene did not appear to have any effect on levels of blood urea nitrogen (BUN) (Philip et al., 2007) or renal catalase (Odum et al., 1988).

Renal morphological and histological changes occurring after subacute and subchronic exposure to tetrachloroethylene primarily affected the proximal tubule, with the majority of effects noted in rats. Protein (hyaline) droplet accumulation was the effect observed at gavage doses of ≥ 500 mg/kg bw per day (Green et al., 1990; Bergamaschi et al., 1992), but not in an inhalation study (Green et al., 1990); the effects were sometimes observed in females, but they occurred earlier and more severely in males (Bergamaschi et al., 1992). Renal peroxisome proliferation was observed in rats and male mice at ≥ 200 ppm (Odum et al., 1988) and in mice exposed to 1000 mg/kg bw per day by gavage (Goldsworthy and Popp, 1987). Other renal tubular effects observed at higher exposure levels, and predominantly in male rats, included necrosis (Goldsworthy et al., 1988), regeneration (Goldsworthy et al., 1988), cellular proliferation (Goldsworthy et al., 1988), tubular casts (Green et al., 1990) and vacuolation and karyomegaly (enlarged nuclei) in female rats (Jonker et al., 1996). In mice, reported changes included karyomegaly in the renal tubule epithelium (NTP, 1986), unspecified changes to proximal tubules (JISA, 1993) and hypercellular effects in the glomerulus (Ebrahim et al., 1996). Effects on other organs and systems

Two subacute studies investigated the immunotoxicity of tetrachloroethylene. Tetrachloroethylene was considered to enhance allergic reactions due to dose-dependent increases in passive cutaneous anaphylaxis in mice with mean intakes of 0.07 µg/day and 7.2 µg/dayFootnote 5, via drinking water prior to an antigen challenge (Seo et al., 2012). No increase in susceptibility to respiratory infection was observed in mice exposed to 25 ppm via inhalation (Aranyi et al., 1986).

Lung congestion was observed in male and female rats exposed to 1600 ppm, but not lower levels, via inhalation for 13 weeks (NTP, 1986).

9.2.3 Long-term exposure and carcinogenicity

No long-term studies have investigated the adverse effects of tetrachloroethylene from exposure in drinking water. Although a few experiments have exposed animals via gavage, the vast majority of studies of tetrachloroethylene in animals have relied on inhalation exposure.

The three major lifetime studies in rodents exposed rats and mice (50 of each sex per group) to tetrachloroethylene at various concentrations. The first inhalation study exposed F344 rats to 0, 200 or 400 ppm and B6C3F1 mice to 0, 100 or 200 ppm, both for 6 hours/day, 5 days/week, for 103 weeks (NTP, 1986), and the other exposed F344 rats to 0, 50, 200 or 600 ppm and Crj:BDF1 mice to 0, 10, 50 or 250 ppm for 6 hours/day, 5 days/week, for 104 weeks (JISA, 1993). The ingestion study exposed animals by gavage in corn oil, for 71 weeks (with observation until 110 weeks) for Osborne-Mendel rats and for 78 weeks (with observation until 90 weeks) for B6C3F1 mice (NCI, 1977). Dose levels in all groups were changed throughout the study, leading to respective time-weighted average doses of 471 and 474 mg/kg bw per day in low-dose male and female rats, 941 and 949 mg/kg bw per day in high-dose male and female rats, 536 and 386 mg/kg bw per day in low-dose male and female mice and 1072 and 772 mg/kg bw per day in high-dose male and female mice, respectively. This ingestion study is considered inconclusive, because high incidences of respiratory disease and mortality were observed in exposed groups, leading to the early termination of the study; therefore, the NTP (1986) and JISA (1993) studies hold more weight in the dose-response assessment. Clinical and serological indicators of health

General health effects were observed after lifetime tetrachloroethylene exposure in all three studies. Body weight gains were inhibited in both rats and mice in two of the lifetime studies (NCI, 1977; JISA, 1993), but not in the NTP (1986) study; furthermore, feed consumption was decreased in both rats and mice in the JISA (1993) study. Animals exposed to tetrachloroethylene tended to have decreased survival when compared with controls in all three studies. The effects appeared to be dose related, with more highly dosed animals having higher mortality levels and earlier deaths (NCI, 1977; NTP, 1986; JISA, 1993). Clinical signs in animals exposed via gavage included hunched appearance and stained fur in both rats and mice, body sores and irritation, alopecia and hunched appearance in mice, and salivation in rats (NCI, 1977).

Haematological and biochemical effects were also measured in the JISA (1993) study. Mean corpuscular haemoglobin levels were increased in high-dose female rats and mice, but decreased in high-dose male mice. Mean cell volume and platelet counts were also decreased in male mice exposed to 250 ppm. Red blood cell count and hematocrit levels were increased in the high-dose group in both male and female mice. In low-dose female mice, the segmented neutrophil ratio was increased and lymphocyte ratio was decreased. Triglyceride levels were decreased in high-dose female rats and high-dose male mice, and glucose level reduction was observed in high-dose male mice; however, total protein and total cholesterol levels were increased in the high-dose male mice. LDH activity was increased in male and female mice. Although potassium concentrations were increased in high-dose female rats, they were decreased in high-dose female mice; calcium concentrations were increased in high-dose female mice, and chloride concentrations were decreased in high-dose male and female mice (JISA, 1993). Additional biochemical effects that were measured were indicators of hepatic and renal effects; therefore, these results will be described in the sections related to those organ systems. Neurological effects

Although the three lifetime studies did not investigate more subtle neurological effects in study animals, they did examine brain tissues for gross and histological abnormalities. No adverse effects were noted in the NCI (1977) study. Relative brain weights were increased in high-dose male and female mice in the JISA (1993) study. Furthermore, vitreous deposits in the brain were significantly increased in 50 ppm male mice, but not in the high dose males (JISA, 1993). The NTP (1986) study identified a significantly positive dose-related trend in the increase in gliomas in male rats exposed to 200 and 400 ppm; however, the incidence of gliomas in each dose group was not elevated over control (NTP, 1986). Hepatic effects

The only study to report changes in liver weight was the JISA (1993) study. Both absolute and relative liver weights were increased in female rats exposed to 50 and 600 ppm (but not 200 ppm, and no effects were observed in male rats). Increases in both absolute and relative liver weights were observed in male mice exposed to 250 ppm, but only relative liver weight was increased in female mice exposed to 250 ppm (JISA, 1993).

Among the lifetime studies, only JISA (1993) considered biochemical effects. Effects on blood levels of hepatic indicators were observed, with an increase in ALT and levels in high-dose male rats and female mice and mid- and high-dose female rats and male mice; increases in AST in mid-dose males and high-dose females, and total bilirubin and ALP in high-dose males and females were also observed in mice (JISA, 1993).

Hepatic effects were more prevalent in mice than in rats. No adverse effects were observed in rats in two of the studies (NCI, 1977; NTP, 1986). In the third study, the NOAEL for hepatic effects was 50 ppm, with increases in spongiosis hepatitis in males at ≥ 200 ppm and decreases in granulation in females at ≥ 200 ppm that were dead or dying (JISA, 1993). In 600 ppm male rats, hyperplasia was also increased, but not significantly. Although there were no significant increases in tumours, two cases of hepatocellular adenoma were observed (compared with zero in all other groups).

Non-neoplastic histological observations in the livers of mice were made in all three of the lifetime studies. Angiectasis and central degeneration were observed in males and females exposed to 250 ppm (as well as males scheduled for dissection only) (JISA, 1993). Central degeneration was also significantly increased in males and females exposed to 250 ppm. Necrosis occurred in males exposed to 250 ppm (JISA, 1993), males exposed to ≥ 100 ppm and females exposed to 200 ppm (NTP, 1986). Cell proliferation, in the form of foci of altered cells, was observed only in three males in the high-dose gavage group (NCI, 1977), and not in the other studies. The lowest-observed-adverse-effect level (LOAEL) for non-neoplastic effects, based on necrosis in male mice in the NTP (1986) study, was 100 ppm.

Tumours were also observed in mice in the three lifetime studies. Hepatocellular adenomas and carcinomas were increased in mice in all of the studies, with a significantly positive trend. Hepatocellular carcinomas had also metastasized to other organs in one study (NTP, 1986). A positive trend in haemangioendothelioma was also observed in male mice in one study (JISA, 1993). Renal effects

The effect of tetrachloroethylene on kidney weights was inconsistent. Both absolute and relative kidney weights were increased in male rats exposed to ≥ 200 ppm, and relative kidney weight was increased in female rats exposed to ≥ 200 ppm and in male and female mice exposed to 250 ppm, but a decrease in absolute kidney weight was observed in male mice exposed to 250 ppm (JISA, 1993).

Biochemical effects were measured only in the JISA (1993) study. In this study, inconsistent kidney-relevant effects were observed in the blood, with BUN level increases in mid-dose female mice and decreases in high-dose male rats (JISA, 1993).

Non-neoplastic effects were observed in the kidneys of rats and mice, with the majority of effects observed in the proximal tubule. Nuclear enlargement was observed primarily in the proximal tubule in all rat and mouse exposure groups in the NTP (1986) study and in male rats exposed to ≥ 200 ppm, female rats exposed to 600 ppm and male and female mice exposed to 250 ppm (as well as low-dose males scheduled for dissection) in the JISA (1993) study. Other proximal tubule effects included atypical dilation in all high-dose groups (JISA, 1993); degeneration, necrosis and regeneration in epithelium, inflammatory cell infiltration, fibrosis and focal mineralization in male and female rats (NCI, 1977); and hyperplasia in a small number of rats (NTP, 1986). Nephropathy was observed in the majority of male and female rats and mice exposed by gavage (NCI, 1977), and nephrosis occurred in female mice exposed by inhalation in the NTP (1986) study. Casts were also observed in the kidneys of dosed male mice and high-dose female mice in the NTP (1986) study. The NOAEL for renal effects in the JISA (1993) study was 50 ppm, with nuclear enlargement occurring at 250 ppm; the LOAEL in the NTP (1986) study was 100 ppm for the same effect.

A slight increase in renal neoplasms (no statistical significance reported) was observed in male rats in one of the studies. In the NTP (1986) study, tubular cell adenomas and adenocarcinomas (combined) were observed in 2% of controls, 6% of low-dose rats and 8% of high-dose rats. Effects on other organs and systems

Evidence of immunotoxicity was observed in the JISA (1993) study. Splenic enlargement occurred in male and female rats of all exposure groups, and absolute and relative spleen weights were increased in 250 ppm male mice. Extramedullary hematopoiesis was decreased in male rats exposed to 200 ppm (but not higher) that were scheduled for dissection; conversely, it was increased in male and female mice exposed to the highest dose, as well as in low-dose female mice.

The relative weight of adrenal glands was increased in high-dose male mice in the JISA (1993) study. Moreover, hyperplasia was increased in the adrenal medulla in dosed male rats and in the adrenal cortex of high-dose female rats in the NTP (1986) study.

Effects on the gastrointestinal tract were observed in animals exposed via inhalation. These effects included forestomach ulcers in male rats exposed to 400 ppm (NTP, 1986) and decreased stomach ventriculus hyperplasia in male and female mice exposed to 250 ppm that were scheduled for dissection, as well as the 10 ppm male mice scheduled for dissection (JISA, 1993).

Tetrachloroethylene adversely affected teeth and the skeletal system in mice in the JISA (1993) study. Effects observed were dysplasia of teeth in ≥ 50 ppm male mice and increased osteosclerosis in 250 ppm female mice.

Non-neoplastic respiratory tract effects were observed, but only in the inhalation studies. Some of these effects included increased lung weight in female rats and male and female mice (JISA, 1993); reduction in cellular changes in respiratory epithelium in female rats and male and female mice and in olfactory epithelium in male and female mice (JISA, 1993); nasal cavity thrombosis in rats (NTP, 1986) and unspecified blood clots in male rats (JISA, 1993); and lung congestion in male mice (NTP, 1986). Cardiovascular effects were also observed, but were limited to an increase in relative heart weight in female mice exposed to 200 ppm (JISA, 1993).

Tetrachloroethylene exposure was also associated with neoplastic effects in various organ systems that have not been described in the previous sections. An increase in the incidence of mononuclear cell leukaemia (MCL) was observed in male and female rats in the NTP (1986) study, with significance measured both in increases over control in all exposed groups and in a positive trend. In earlier stages of the disease, infiltration was primarily limited to the liver sinusoids and interfollicular pulp of the spleen, but as the disease advanced, infiltration affected most organs and tissues (NTP, 1986). MCL also showed a significant trend in male and female mice and a significant increase in high-dose male mice in the JISA (1993) study. Increases in mammary fibroadenomas were observed in one study in female rats in the lowest exposure group (50 ppm), but not in higher dose groups (JISA, 1993). A significant positive trend was observed in pheochromocytomas of the adrenal gland in male rats (NTP, 1986), Harderian (lacrimal) gland adenomas in male mice and hemangioendothelioma in all organs—including the spleen—in male and female mice (JISA, 1993). Nasal squamous cell metaplasia was also observed in dosed male rats in one study (NTP, 1986).

9.2.4 Reproductive and developmental toxicity

Animal studies of reproductive and developmental effects have focused primarily on the inhalation route of exposure, with limited evidence of effects via oral exposure. Three studies have addressed reproductive and/or developmental effects of tetrachloroethylene exposure via the oral route. These studies had few doses and endpoints; thus, only limited conclusions can be drawn. Exposure of pregnant rats to ≥ 900 mg/kg bw per day by oral gavage on gestation days 6-19 resulted in increased resorption and postnatal death, as well as increased cases of microphthalmia/anophthalmia in offspring (Narotsky and Kavlock, 1995). The two other ingestion studies (Fredriksson et al., 1993; Guariglia et al., 2011) identified neurobehavioural effects in mice that were exposed prenatally to ≥ 5 mg/kg bw per day; these studies were discussed in Section

Studies of exposure to tetrachloroethylene by inhalation throughout gestation do not suggest teratogenicity at doses as high as 500 ppm in rabbits and 900 ppm in rats (Nelson et al., 1980; Hardin et al., 1981; Carney et al., 2006). However, several studies have reported incomplete or delayed ossification of rats and mice exposed in utero (Schwetz et al., 1975; Szakmáry et al., 1997; Carney et al., 2006). Adverse neurological effects have been observed after intrauterine exposure, with dams exposed by inhalation (Nelson et al., 1980; Kyrklund and Haglid, 1991; Szakmáry et al., 1997), as discussed in Section Inhalation studies suggest that tetrachloroethylene exposure during gestation may increase cases of resorption as well as decrease fetal weight (Schwetz et al., 1975; Szakmáry et al., 1997). A multigenerational study in rats exposed to 1000 ppm resulted in reduced live births, increased pup mortality during lactation, as well as reduced growth with and without pup exposure, thus confirming in utero effects (Tinston, 1995). When unexposed females were mated with exposed males, no effects were observed, thus suggesting no paternal effects of exposure in offspring (Tinston, 1995). However, one study noted increased sperm abnormalities in mice following exposure to 500 ppm tetrachloroethylene for 7 hours/day over 5 consecutive days, although this effect was not observed in rats (Beliles, 2002).

Overall, studies suggest that exposure to tetrachloroethylene during gestation can affect survival of offspring and may be involved in alterations in ossification. Inhalation exposure through gestation or early in development may also be linked to neurological effects in adulthood.

9.3 Genotoxicity

9.3.1 In vitro findings

The genotoxicity of tetrachloroethylene and most of its metabolites has been investigated in vitro. Overall, studies do not suggest that tetrachloroethylene is genotoxic. However, metabolites of tetrachloroethylene may be involved in genotoxic outcomes.

Tetrachloroethylene did not increase mutation rates in the thymidine kinase locus of L5178Y/TK+/- mouse lymphoma cells at any dose tested (6.25, 12.50, 25, 50 and 100 nL/mL) with and without metabolic activation with S9, despite increased mutations observed in positive controls (NTP, 1986). Gene mutations were also not observed in bacterial systems, including Escherichia coli and Salmonella typhimurium, in the absence of metabolic activation or in the presence of standard S9 (Greim et al., 1975; Bartsch et al., 1979; Hardin et al., 1981; Kringstad et al., 1981; Haworth et al., 1983; Connor et al., 1985; Shimada et al., 1985; NTP, 1986; Milman et al., 1988; Warner et al., 1988; Roldan-Arjona et al., 1991; DeMarini et al., 1994; Watanabe et al., 1998; Emmert et al., 2006). However, exposure of bacterial cells with enhanced metabolic activity resulted in a clear dose-response relationship, suggesting the role of metabolic activation in tetrachloroethylene genotoxicity (Vamvakas et al., 1989a).

Studies of additional genotoxic endpoints also do not suggest tetrachloroethylene genotoxicity. Chromosomal aberrations were not observed in Chinese hamster ovary cells exposed to tetrachloroethylene at concentrations of 17, 34.1, 68.1 and 136.3 μg/mL with and without S9 metabolic activation (the highest concentration was not tested in the presence of standard S9) (NTP, 1986). Micronuclei induction was not affected by tetrachloroethylene exposure (125-250 μg/mL) in a Chinese hamster lung cell line in the absence of metabolic activation. However, a marginal increase (non-statistically significant) in micronuclei was observed at a lower dose of 75 μg/mL in the presence of S9 (Matsushima et al., 1999). Experiments in various metabolically competent human lymphoblastoid cell lines resulted in increased micronuclei induction at concentrations as low as 5, 1 and 1 mM for the AHH-1, H2E1 and MCL-5 cell lines, respectively (Doherty et al., 1996). Exposure of the MCL-5 cell line at tetrachloroethylene concentrations of 0, 0.01, 0.05, 0.1, 0.25, 0.5, 1.0 and 2.0 mM over 24 hours resulted in a dose-dependent induction of micronuclei (White et al., 2001). Exposure to tetrachloroethylene via vapours dissolved in culture medium (with a peak concentration in the culture medium of approximately 63 ppm) resulted in dose-dependent increases in micronuclei induction in Chinese hamster ovary cells (Wang et al., 2001). Tetrachloroethylene did not induce sister chromatid exchanges or DNA strand breaks in cultured human blood cells exposed to up to 5 mM (approximately 830 mg/L) tetrachloroethylene, a concentration that induced 40% cell death (Hartmann and Speit, 1995). Analysis of sister chromatid exchanges and chromosomal aberrations in Chinese hamster ovary cells gave negative results for both assays (Sofuni et al., 1985; Galloway et al., 1987). An additional study in Chinese hamster ovary cells using concentrations of up to 164 µg/mL showed no increase in sister chromatid exchanges in the presence and absence of standard S9 (NTP, 1986). Tetrachloroethylene exposure was not associated with unscheduled DNA synthesis, as measured in cultured human lymphocytes, human fibroblasts and rat/mouse hepatocytes. One study that examined unscheduled DNA synthesis in human fibroblasts exposed to concentrations ranging from 0.1 to 5.0 µL/mL reported increased DNA synthesis at low doses that was similar to that observed in positive controls, whereas highly exposed cells were not affected (although cytotoxicity was high) (Beliles et al., 1980).

Studies have investigated the genotoxicity of tetrachloroethylene metabolites, including TCA, DCA and chloral hydrate; limited information is available on other metabolites. Some of these metabolites were shown to exhibit genotoxic activity. Overall, TCA did not induce mutation in S. typhimurium with and without metabolic activation (Shirasu et al., 1976; Waskell, 1978; Nestmann et al., 1980; Rapson et al., 1980; Moriya et al., 1983; DeMarini et al., 1994; Nelson et al., 2001; Kargalioglu et al., 2002). Studies in cultured mammalian cells revealed no induction of DNA strand breaks in mouse and rat hepatocytes, Chinese hamster ovary cells or human lymphoblastic cell lines (Chang et al., 1992; Plewa et al., 2002). Weak mutagenic effects of TCA were observed at a cytotoxic concentration of 3000 µg/mL in mouse lymphoma cells (Harrington-Brock et al., 1998). The genotoxicity of DCA varied across studies. Studies in mammalian systems have revealed increased mutation rates and chromosomal aberrations in a mouse lymphoma cell line, whereas studies in bacterial systems have shown increased mutation frequency in S. typhimurium (DeMarini et al., 1994; Giller et al., 1997; NTP, 2007). Overall, chloral hydrate tested positive in bacterial mutation tests (Haworth et al., 1983; Ni et al., 1994; Beland, 1999) with and without metabolic activation, as well as in the mouse lymphoma assay for mutagenicity at the Tk locus (Harrington-Brock et al., 1998). Clastogenic effects, including induction of micronuclei in Chinese hamster cells (Degrassi and Tanzarella, 1988; Parry et al., 1990; Lynch and Parry, 1993; Seelbach et al., 1993) and mouse lymphoma cells (Nesslany and Marzin, 1999) and aneuploidy in human lymphocytes (Vagnarelli et al., 1990; Sbrana et al., 1993) and various other mammalian cell systems (Furnus et al., 1990; Natarajan et al., 1993; Warr et al., 1993; Harrington-Brock et al., 1998) were observed. Few studies examined additional metabolites, including trichloroacetyl chloride, tetrachloroethylene epoxide and trichloroethanol. Trichloroacetyl chloride was mutagenic in S. typhimurium when exposed in the vapour phase (DeMarini et al., 1994), but not in the liquid phase (Reichert et al., 1983). Tetrachloroethylene epoxide was mutagenic in S. typhimurium, but not in E. coli. Finally, no evidence of mutagenicity was found for trichloroethanol (Waskell, 1978; Bignami et al., 1980; DeMarini et al., 1994). Additional metabolites, including TCVC, TCVG and NAcTCVC, observed upon glutathione conjugation were positive in bacterial screening assays with S. typhimurium (Vamvakas et al., 1987, 1989a; Dreessen et al., 2003).

9.3.2 In vivo findings

In vivo investigations of tetrachloroethylene provide some support for genotoxicity, although the majority of studies reported negative findings. Positive results for frameshift mutagenicity were observed using the S. typhimurium strain TA98 implanted into the peritoneal cavity of male and female CD-1 mice exposed to tetrachloroethylene by inhalation at 100 and 500 ppm (7 hours/day for 5 days) (Beliles et al., 1980). However, positive results were found at the low dose only in males and high dose only in females. No significant effects on chromosomal aberrations were observed in rats exposed to 100 or 500 ppm (7 hours/day for 5 days) (Beliles et al., 1980). Oral and inhalation exposures to radiolabelled tetrachloroethylene in mice (10 or 600 ppm by inhalation for 6 hours measured at 6, 24, 48 and 72 hours post-exposure and 500 mg/kg bw by oral gavage measured at 1, 6, 12, 24 and 48 hours post-exposure) did not induce DNA binding, whereas binding to protein and RNA increased (Schumann et al., 1980). Low levels of tetrachloroethylene binding to DNA were detected in mouse liver 22 hours after intraperitoneal injection of 1.4 mg/kg bw (Mazzullo et al., 1987). Exposure of ddY mice to a single intraperitoneal injection of tetrachloroethylene at 1000 or 2000 mg/kg bw resulted in increased micronucleus induction in hepatocytes that was exclusively observed after partial hepatectomy (Murakami and Horikawa, 1995). Micronucleus induction was not observed in peripheral blood reticulocytes in this study. Tetrachloroethylene did not induce DNA strand breaks in kidney of male F344 rats following a 7-day oral gavage exposure to tetrachloroethylene at 1000 mg/kg bw (Potter et al., 1996). One study reported marginal increases in DNA strand breaks in liver, but not in kidney, of CD-1 mice exposed to two doses (24 hours apart) of 1000 or 2000 mg/kg bw by gavage, 3 hours after the second dose (Cederberg et al., 2010). However, the significance of these results has been disputed after reanalyses of the data using more conservative statistical approaches, which indicated that these results may not represent significant increases (Lillford et al., 2010; Lovell, 2010). Exposure of NMRI mice to tetrachloroethylene at 4-8 mmol/kg bw by intraperitoneal injection increased DNA single strand breaks in liver and kidney that were reversible within 24 hours post-exposure, but no effects were observed in lung (Walles, 1986). Single intraperitoneal injections of up to 1000 mg/kg bw in male Fischer rats caused no increase in 8-hydroxydeoxyguanosine (8-OH-dG) levels within liver, lymphocytes or urine; however, this exposure level caused high toxicity (Toraason et al., 1999). Only negative findings for induction of sex-linked recessive lethal mutations and chromosomal aberrations through exposure by feeding, inhalation and injection have been reported in Drosophila melanogaster (Beliles et al., 1980; Valencia et al., 1985; NTP, 1986).

A limited number of studies have evaluated the genotoxicity of tetrachloroethylene metabolites in vivo. Genotoxic outcomes of TCA exposure varied across studies. Some studies of oral exposure demonstrated genotoxicity in mouse liver (Nelson and Bull, 1988; Nelson et al., 1989; Hassoun and Dey, 2008) and induction of micronuclei and chromosomal aberrations in bone marrow of mice (Bhunya and Behera, 1987) at doses as low as 22 mmol/kg bw. However, negative findings for liver, stomach and duodenum epithelial DNA strand breaks (Nelson et al., 1989; Chang et al., 1992) in mice upon oral exposure at a dose as high as 1630 mg/kg bw have also been reported. Oral exposure of mice to DCA has been shown to result in liver DNA strand breaks (Nelson and Bull, 1988; Nelson et al., 1989; Hassoun and Dey, 2008) in most of the studies that have investigated these endpoints and micronuclei in erythrocytes (Fuscoe et al., 1996). Experiments with chloral hydrate suggest that it may be involved in aneuploidy through spindle poisoning. Several studies indicate that chloral hydrate can induce DNA strand breaks in liver of mice and rats upon oral exposure (Nelson and Bull, 1988), micronuclei in mouse bone marrow and spermatids upon intraperitoneal injection (Russo and Levis, 1992; Russo et al., 1992; Allen et al., 1994; Marrazzini et al., 1994; Nutley et al., 1996; Beland, 1999) and aneuploidy in mouse secondary spermatocytes and bone marrow erythrocytes upon intraperitoneal injection (Miller and Adler, 1992; Marrazzini et al., 1994). Increased frequency of micronucleus formation and sister chromatid exchange in lymphocytes was observed in infants after a single exposure to chloral hydrate as a sedative at 50 mg/kg bw (Ikbal et al., 2004). In vivo experiments on additional metabolites were not available.

Tetrachloroethylene exposure was not clastogenic in human lymphocytes of exposed factory workers according to sister chromatid exchange tests and examination for chromosomal aberrations at concentrations ranging from 70 to 1500 mg/m3 (Ikeda et al., 1980). This finding is further supported by an additional study that reported negative results for sister chromatid exchanges in peripheral lymphocytes of occupationally exposed individuals (Seiji et al., 1990). No evidence of oxidative DNA damage, as measured by 8-OH-dG in leukocytes and urinary 8-OH-dG was found in female dry cleaners relative to unexposed launderers (Toraason et al., 2003).

9.4 Mode of action

An in-depth mode of action (MOA) analysis was performed (Health Canada, 2013) according to guidelines set out in the International Life Sciences Institute/International Programme on Chemical Safety conceptual frameworks for the evaluation of cancer and non-cancer endpoints (IPCS, 2007). Results of the assessment for MOAs potentially relevant to the major endpoints considered in Section 10 are summarized here.

9.4.1 Hepatocellular adenomas and carcinomas

Mutagenicity and other genotoxicity MOAs were considered for tetrachloroethylene. As described in Section 9.3, the weight of evidence from in vivo and in vitro studies demonstrates that tetrachloroethylene and its major metabolite, TCA, are not genotoxic. The genotoxicity of other metabolites of tetrachloroethylene was considered in the assessment. Data were limited for other metabolites acting via oxidative metabolism. Increased DNA strand breaks were observed in a dose-related manner in a single in vivo study of chloral hydrate (Nelson and Bull, 1988); however, the increase was significant only in the highest dose group, and not at lower doses, at which hepatocellular adenomas and carcinomas from chloral hydrate exposure were observed. Hepatocellular tumours from chloral hydrate exposure are therefore not considered to be due to mutagenicity. Oxidative metabolism of tetrachloroethylene is expected to produce a small amount of an epoxide (tetrachloroethylene oxide), which would be suspected to be reactive due to the electrophilicity of chloroalkene epoxides. Studies of this epoxide are limited to in vitro exposure; although the metabolite is genotoxic in bacterial systems (Kline et al., 1982), it is difficult to identify the relevance of this effect at the low levels that would be produced by metabolism of tetrachloroethylene. Metabolites via the GST pathway were also considered for their genotoxicity. Studies of DCA identified DNA strand breaks (Nelson et al., 1989; Hassoun and Dey, 2008) and positive gene mutation (Leavitt et al., 1997) observed at doses lower than those causing hepatocellular tumours. For other GST metabolites, only in vitro studies were performed, and these identified the mutagenicity of TCVG (Vamvakas et al., 1989a; Dreessen et al., 2003), TCVC (Green and Odum, 1985; Dekant et al., 1986; Vamvakas et al., 1989b; Dreessen et al., 2003) and NAcTCVC (Vamvakas et al., 1987). These data suggest that genotoxicity from tetrachloroethylene and its oxidative metabolites is not likely to be a relevant MOA, but that the mutagenicity of GST metabolites is a plausible MOA for hepatic tumours. Further support for this is provided by quantitative results of the PBPK model, as the pattern of increase of GST—and not oxidative—metabolites matches the pattern of significant increases in hepatocellular tumours. Although this MOA would not be presumed to be relevant to humans, as the GST metabolism of tetrachloroethylene in humans is considered to be minor, it cannot be completely ruled out, as even low exposures to the genotoxic metabolites could play a role in tumour development.

Epigenetic change, specifically changes in DNA methylation, was considered as a potential MOA for carcinogenicity. The MOA involves the decrease in global DNA methylation, which leads to an increase in the messenger RNA and protein synthesis of proto-oncogenes (e.g., c-myc, c-jun and/or the growth promoting hormone IGF-II), resulting in cell proliferation. Data specific to tetrachloroethylene are lacking, but DCA and TCA have been shown to induce hypomethylation of DNA and proto-oncogenes in mouse liver (Tao et al., 2000, 2004; Ge et al., 2001; Pereira et al., 2001, 2004). The exposure levels in these few studies, however, have been higher than those that are associated with hepatocellular tumours in mice. This endpoint is therefore plausible, but further research is required to identify whether the MOA is quantitatively relevant to hepatocellular tumours induced by tetrachloroethylene or its metabolites, particularly at levels that are the same as or lower than those in the carcinogenicity bioassays.

Another MOA that was considered for hepatocellular tumours was peroxisome proliferation. Peroxisome proliferation is involved in fatty acid metabolism (Feige et al., 2006). Various chemicals can activate peroxisome proliferator-activated receptor alpha (PPARα), but once this activation occurs, the pathway is independent of the chemical. As a by-product of the fatty acid metabolism, oxidative stress occurs, which results in indirect DNA damage. Oxidative stress then results in the activation of NF-κB, a negative regulator of apoptosis, leading to a perturbation of hepatocyte growth. This can result in selective clonal expansion and tumour development (Corton, 2010).

Although increases in markers of PPARα activation were observed in tetrachloroethylene-exposed mice (Goldsworthy and Popp, 1987; Odum et al., 1988), the exposure levels at which these effects were observed were higher than those at which hepatocellular tumours were observed. This discrepancy is potentially related to study design in the PPARα activation marker studies, as no studies have been performed at levels lower than those used in the carcinogenicity studies. Philip et al. (2007) noted an additional weakness in this endpoint, which is that CYP4A expression was not sustained for as long as 2 weeks; however, palmitoyl coenzyme A (CoA) oxidase activity increases were sustained for as long as 28 days (Odum et al., 1988), and no similar studies of longer duration have been performed. A lack of data on oxidative stress and hepatocyte growth perturbation also decreases the ability to identify whether peroxisome proliferation is a main MOA for tetrachloroethylene.

The major metabolite of tetrachloroethylene, TCA, has been considered to be a peroxisome proliferator. The most commonly used marker for PPARα in animal (primarily mouse) studies was palmitoyl CoA oxidase activity. TCA exposure consistently resulted in increased activity in this marker at exposure levels as low as 19 mg/kg bw per day (Austin et al., 1995; Parrish et al., 1996; DeAngelo et al., 2008), except in a study of short duration (1 week), in which 400 mg/kg bw per day was required for this activity (Laughter et al., 2004); these levels are lower than those at which hepatocellular adenomas and carcinomas were observed. The other markers of this effect were less studied and occurred at higher exposure levels. Markers of oxidative stress (8-OH-dG and thiobarbituric acid reactive substances [TBARS] activity) and hepatocyte growth perturbation (hepatocyte proliferation and centrilobular hepatocyte hypertrophy) were not as well studied and tended to occur at higher levels than tumours. The TCA data therefore provide a clear improvement in certainty of the MOA when compared with tetrachloroethylene, as PPARα activation is measured at lower levels than tumour induction; confidence in this MOA for TCA could be strengthened if measures of oxidative stress and hepatocyte growth perturbation were measured at levels lower than those associated with hepatocellular tumours.

The observation of peroxisome proliferation in rats can also provide some additional support for the peroxisome proliferation MOA. Rats, in which no increased incidence of hepatocellular tumours was observed, had lower levels of activation of PPARα for both tetrachloroethylene (Odum et al., 1988) and TCA (Mather et al., 1990; DeAngelo et al., 1997).

These analyses indicate that a peroxisome proliferation MOA is potentially plausible for tetrachloroethylene-induced hepatocellular tumours; however, there are key data gaps, which prevent the confident identification of peroxisome proliferation as a main MOA. There is stronger evidence that TCA exposure leads to hepatocellular tumours via a peroxisome proliferation MOA, but there are again limitations in the existing data. However, the weight of evidence supporting tetrachloroethylene's activity via this MOA, with TCA being a toxic moiety, is stronger than for other MOAs. If the data are considered sufficiently supportive of tetrachloroethylene acting by a peroxisome proliferation MOA, a threshold approach could be used for risk assessment. This is because a peroxisome proliferation MOA is most likely a threshold effect, since low concentrations of tetrachloroethylene—and therefore low internal doses of TCA—activate the PPARα receptor only at low levels, leading to inadequate signalling to generate increased peroxisome levels. If tetrachloroethylene were considered to act via a peroxisome proliferation MOA, the human relevance of hepatocellular tumours could also be considered to be limited. Lower levels of TCA are generated in humans than in animals; therefore, higher exposures to tetrachloroethylene would be required to generate sufficient levels of TCA. Moreover, quantitative and qualitative toxicodynamic differences in PPARα activation appear to occur between experimental animals and humans. Humans exhibit ≥ 10-fold lower PPARα expression than mice and rats, and activation of the receptor triggers different gene targets, with an absence of induction of liver cell proliferation and suppression of apoptosis (Corton, 2010). PPARα activators are used as a human therapeutic agent for hyperlipidemia; the therapeutic effect in humans appears to be separate from the proliferating effect (Fidaleo, 2009). Peroxisome proliferation does not occur and liver tumours are not developed in mice transplanted with human PPARα and exposed to peroxisome proliferators (Corton, 2010). In support of this, only one clinical trial has ever identified a potential for an increase in cancer deaths (associated with clofibrate), but these observations diminished with follow-up (Klaunig et al., 2003); however, these trials have low power to identify hepatic tumours, and treatment was halted in participants demonstrating liver toxicity, which precludes the ability to make conclusions about the carcinogenicity of these therapeutic agents (Guyton et al., 2009). In general, the adverse effects of peroxisome proliferators are considered to be of limited relevance to humans.

Recent studies of known peroxisome proliferators, however, suggest that activation of PPARα might be a necessary but insufficient step in tumorigenesis. A study in PPARα-null mice exposed to di(2-ethylhexyl) phthalate, a peroxisome proliferator, identified rates of liver tumour development that were similar to those for wild-type mice with active PPARα receptors, indicating that the tumours occurred independently of PPARα activation (Ito et al., 2007; Guyton et al., 2009). Moreover, activation of PPARα via constitutive expression in transgenic mice resulted in similar hepatic responses in rodents activated with the model peroxisome proliferator Wy-14,643 (including activation of acyl CoA, palmitoyl CoA, CYP4A and cell cycle genes and decreases in triglycerides and fatty acids that were quantitatively similar to those observed with the ligand-induced peroxisome proliferation), but did not develop the expected preneoplastic lesions (Yang et al., 2007). Guyton et al. (2009) also demonstrated that there is a lack of correlation between the PPARα activation potency of a compound and its potency in hepatic tumorigenesis. These observations suggest that PPARα activation and other key events discussed in this section are not completely sufficient for liver tumour development, and additional unknown key events might play a role in the hepatocarcinogenesis of peroxisome proliferators. The previous conclusion that peroxisome proliferators are of limited human relevance is weakened considering the absence of knowledge of these additional key events (Guyton et al., 2009).

A final MOA that was considered for hepatocellular tumours was cytotoxicity resulting from oxidative stress, which can lead to regenerative proliferation. Few data exist for this MOA, but chronic studies have identified hepatotoxicity, including hepatocellular degeneration and necrosis, resulting from tetrachloroethylene exposure (NTP, 1986; JISA, 1993). Mice administered high doses (3 g/kg bw per day) of tetrachloroethylene for 15 days demonstrated these effects, which were reduced in animals that were subsequently exposed to antioxidants (Ebrahim et al., 1996), providing support for the effects being due to oxidative stress. In an in vitro study, similar effects were observed, with tetrachloroethylene exposure causing significant elevations in TBARS production in rat hepatocytes, which was modulated when co-exposures with vitamin E occurred (Costa et al., 2004). Although these studies indicate that tetrachloroethylene exposure likely results in oxidative stress and that neutralizing the oxidant activity reduces hepatocyte necrosis and degeneration, there are no studies demonstrating that the hepatocellular tumours produced as a result of tetrachloroethylene exposure progress from the hepatic necrosis and degeneration. If a longer-term study with co-exposures to tetrachloroethylene and antioxidants were performed, this might give a better indication as to whether the antioxidants would reduce or prevent the occurrence of hepatocellular tumours, which could lend support to or allow for the rejection of oxidative stress-related cytotoxicity as a relevant MOA. Studies on TCA and DCA indicate that oxidative stress occurs in the liver in mice at ≥ 7 mg/kg bw per day (Larson and Bull, 1992; Austin et al., 1996; Parrish et al., 1996; Hassoun and Dey, 2008; Hassoun et al., 2010; Hassoun and Cearfoss, 2011; Cearfoss and Hassoun, 2012; Hassoun and Al-Dieri, 2012), which tends to be below levels causing cell proliferation (Carter et al., 1995; Snyder et al., 1995; Ge et al., 2001; DeAngelo et al., 2008).

9.4.2 Mononuclear cell leukaemia

Limited data exist on the MOA of MCL. Although several different MOAs (increased growth hormones leading to proliferation of splenocytes, inhibition of apoptosis leading to cell cycle changes in natural killer cells, immunosuppression leading to decreased elimination of tumour cells and genotoxicity leading to increased mutation of splenocytes) were considered in the analysis, none had adequate data to allow for proper assessment of the plausibility of each of the MOAs or their human relevance. No data could be found related to the effects of tetrachloroethylene or its metabolites on large granular lymphocytes or their precursors. However, high levels of chloral hydrate, a minor intermediate metabolite of tetrachloroethylene, were found to induce micronuclei and sister chromatid exchange in peripheral blood lymphocytes (further delineation of cell types was not mentioned) in infants (Ikbal et al., 2004). Fischer 344 rats are prone to MCL and have high background rates of the disease; however, the incidence of the disease was clearly above historical control rates in tetrachloroethylene-exposed rats (NTP, 1986; JISA, 1993), and exposure to tetrachloroethylene also resulted in a dose-related increase in severity of MCL in female rats and in decreased tumour development latency in male and female rats (NTP, 1986).

Without knowledge of potential MOAs for tetrachloroethylene, it is difficult to assess the human relevance of rat MCL. In rats, MCL is first observed in the spleen; therefore, this is considered to be the organ of origin (Stromberg and Vogtsberger, 1983; Stromberg et al., 1990). The evidence that MCL development is drastically reduced in splenectomized F344 rats (Moloney and King, 1973) provides further support to this hypothesis. The bone marrow was affected by MCL only late in the progression of the disease in rats, and a smaller percentage of rats have bone marrow involvement (Stromberg and Vogtsberger, 1983). Haematopoiesis is different between rats and humans, as extramedullary haematopoiesis occurs in the spleen in rats (Stefanski et al., 1990). In humans, the spleen is a site of haematopoiesis in foetuses and neonates, but bone marrow becomes the primary site for haematopoiesis very early in life. Although haematopoietic stem cells can migrate to the spleen when increased blood cells are required (e.g., in infection, inflammation and diseases affecting the bone marrow), this is not a typical site for haematopoiesis (Kim, 2010). Because of these haematopoietic differences between rats and humans, MCL in rats might not be relevant to humans if the MOA for MCL is found to be related to haematopoiesis; however, because there is no indication as to whether this might be a relevant MOA, the human relevance of MCL in rats cannot currently be discounted.

Many similarities (in the cellular phenotype, clinical, haematological and serum chemistry and pathological characteristics) exist for MCL in rats and the large granular lymphocyte leukaemia associated with natural killer cells in humans (Thomas et al., 2007). Although some epidemiological studies have considered associations between tetrachloroethylene exposure and leukaemia, it is difficult to assess the associations. Participants in the community-based studies were exposed to a variety of chemicals, and the occupational studies considering leukaemia were based on census data and cancer registries; therefore, exposure assessment is limited to the identification of individuals who have reported working in the dry cleaning industry. Moreover, the assessments do not typically distinguish between different types of leukaemia.

9.4.3 Neurotoxicity

Limited data exist on the potential MOAs for tetrachloroethylene-associated neurotoxicity; however, because the effects have been clearly identified in environmental and occupational epidemiological studies, the effects can be considered relevant to humans, despite a lack of MOA data. The MOA of tetrachloroethylene is presumably most closely related to parent compound effects, rather than effects of metabolites (Bale et al., 2011); therefore, decreased tetrachloroethylene metabolism in humans compared with experimental animals does not lessen the likelihood of an effect in humans.

Some evidence exists to indicate that tetrachloroethylene might have effects on dopamine activity, which might have an effect on colour discrimination and contrast sensitivity, effects that were observed in epidemiological and acute controlled dosing studies. Decreased release of dopamine can result in decreases in transmission of nerve signals to the retina, which can result in decrements in visual abilities (Gralewicz and Dyzma, 2005). Although there have not been any direct measures of altered dopamine levels after tetrachloroethylene exposure, other data indirectly suggest that this might occur. Increases in dopamine levels can lead to inhibition of glutamate release, resulting in the activation of gamma-aminobutyric acid (GABA)-producing neurons, which can cause tremors (Gralewicz and Dyzma, 2005). No changes were measured in glutamate or GABA uptake in gerbils exposed to a single exposure level of tetrachloroethylene (Briving et al., 1986), but the anti-convulsive effects observed in rats (Chen et al., 2002) might suggest a decrease in dopamine levels. Dopamine also inhibits the release of prolactin; therefore, the increases in plasma prolactin levels in female dry cleaners (Ferroni et al., 1992) could be an indication of decreased dopamine levels. However, more conclusive evidence of the effects of tetrachloroethylene on dopamine levels is required before this MOA can be confirmed. There is also some evidence that tetrachloroethylene might disrupt the function of voltage- and ligand-gated ion channels. In general, tetrachloroethylene and other related solvents tend to increase inhibitory receptor function and decrease excitatory receptor function, resulting in the inhibition of current (Bale et al., 2011). These changes are expected to affect neurobehavioural responses in the cognitive domain—particularly if the changes occur in the hippocampus—and visual function (Bale et al., 2011). Changes to cholinergic system function are suspected to affect memory and cognition, as demonstrated in Alzheimer's disease research (Mufson et al., 2008). The specific receptors that affect visual function have not been completely elucidated, but one hypothesis is that increased activation of the N-methyl-D-aspartate-glutamate receptor system, likely via increase of agonist neurotransmitters such as glutamate, results in changes to visually evoked potentials (Bale et al., 2011). This hypothesis is supported by evidence that the agonist N-methyl-D-aspartate decreases the F2 amplitude in rats, similar to tetrachloroethylene, but that N-methyl-D-aspartate antagonists do not change visual evoked potentials (Bale et al., 2005a). Several different ion channels have been found to be disrupted by tetrachloroethylene in in vitro studies. Tetrachloroethylene increases the activation and decreases the inactivation of voltage-sensitive calcium channels (Bushnell et al., 2005; Shafer et al., 2005), resulting in shifts in a hyperpolarizing direction. A concentration-dependent inhibition of human and rat neuronal nicotinic acetylcholine receptors has also been observed for tetrachloroethylene (Bale et al., 2005b). Although there are no data for the effects of tetrachloroethylene on N-methyl-D-aspartate, GABA or glycine receptors, data for compounds that potentially act by similar modes of action (toluene, trichloroethylene and ethanol) can be used to predict tetrachloroethylene activity on these receptors. In general, these compounds tend to inhibit excitatory (e.g., N-methyl-D-aspartate) receptors and potentiate inhibitory (e.g., GABA and glycine) receptors (Bushnell et al., 2005).

As mentioned above, the hippocampus controls cognition and memory; therefore, changes directly to this area of the brain could impact the neurobehavioural domains that were affected in exposed humans. One hypothesized MOA that could affect cognition is demyelination of the hippocampus (Bale et al., 2011). Fatty acid changes in the brains of gerbils, including in the hippocampal region, were potentially indicative of changes to the myelin sheath (Kyrklund et al., 1984, 1987).

9.5 Dose-response assessment

An extended summary of the cancer and non-cancer dose-response assessments is presented in Sections 9.5.1 and 9.5.2, respectively, as a supplement to the discussions in Section 10. Several endpoints were considered for each assessment (Health Canada, 2013), but those that were most relevant to the assessment (based on consistency of observations among studies, increased incidence of effects at lower doses and demonstrated relevance to humans) are discussed in detail in these sections.

As described in Section 8.5, a PBPK model was developed for use in both the cancer and non-cancer risk assessments for tetrachloroethylene. This is because quantitative differences in metabolism exist when comparing humans with rodents, and saturation of the oxidative pathway in high-dose studies results in a plateau in the generation of oxidative metabolites while driving the metabolism to the GST pathway. Linear extrapolation from high-dose studies in rodents to low-dose human exposures might not be representative of the actual risk of exposure to drinking water containing low concentrations of tetrachloroethylene. For this reason, internal dose metrics are more appropriate for use in risk assessments of tetrachloroethylene.

The species-relevant models were applied to studies that were deemed to be relevant for the dose-response assessment to allow for the estimation of various internal dose metrics at each of the exposure levels reported in the studies. Wherever possible, benchmark dose values representing a 10% increase in adverse effect over background rates (BMD10) and their lower 95% confidence limits (BMDL10) were obtained for relevant internal dose metrics using the U.S. EPA Benchmark Dose Software (BMDS Version 2.2 R67) (U.S. EPA, 2011a). BMDS was also used to calculate cancer slope factors for each of the internal dose metrics, which allowed for the estimation of excess cancer risks of 10-4, 10-5 and 10-6 for these dose metrics. To extrapolate from these internal dose metrics to external oral doses of concern in humans, the human PBPK model was applied.

9.5.1 Dose-response assessment for carcinogenic endpoints

As described in Section 9.2.3, four cancer types have been identified in chronic bioassays in mice and rats. Two of the tumour types, splenic haemangioendothelioma in male mice (observed in JISA, 1993) and renal tubular adenomas and adenocarcinomas in male rats (observed in NTP, 1986) were initially considered, but were eventually excluded from the dose-response assessment for two reasons: (1) the observations were not replicated in other studies or in other sexes and species within the same study; and (2) the incidence of tumours was low in the studies and only slightly above background, with no significant increase in any dose over controls when using pair-wise comparisons. The two remaining cancer types, hepatocellular adenoma and carcinoma (combined) and MCL, were the main focus of the dose-response assessment, which is described below.

Hepatocellular adenomas and carcinomas were observed in mice in both chronic inhalation studies. In both males and females, significant increases were observed at high doses in the JISA (1993) study and at both doses in the NTP (1986) study, and the trends were significantly positive for both studies. The NCI (1977) study, in which significant increases in the tumours were observed in both dose groups, can also be used as qualitative support of the effect via the oral route, but the study was not considered quantitatively in the dose-response assessment due to the shortened duration of the study and dose levels that varied over time. As discussed in Section 9.4, although multiple MOAs were considered, none was found to be the sole causative factor for hepatocellular tumours. MOAs of peroxisome proliferation from the TCA metabolite and cytotoxicity are suspected to be predominant, with the greatest weight of evidence for peroxisome proliferation. If tumours in mice were due to peroxisome proliferation or cytotoxicity, a threshold dose-response assessment approach could be considered relevant, because preventing the PPARα activity or the development of hepatocellular degeneration and necrosis could be expected to prevent the occurrence of later key events in the MOA pathway, including the development of hepatocellular tumours. The more conservative linear low-dose extrapolation is presented together with the threshold approach, because although the weight of evidence most strongly supports a threshold approach and refutes mutagenicity, there is inadequate evidence to firmly conclude whether either peroxisome proliferation or cytotoxicity is the MOA responsible for these tumours. Moreover, the role of mutagenicity in these tumours cannot be completely ignored, because minor metabolites from the GST pathway have displayed genotoxicity in assays. Although there is not absolute certainty regarding the MOA of tetrachloroethylene for hepatocellular tumours, there is a much stronger weight of evidence to support the threshold approach than for the use of linear low-dose extrapolation for the assessment of hepatocellular tumours.

Elevated levels of MCL were observed in male and female rats in both chronic inhalation studies. These increases were significant for both dose groups in the NTP (1986) study and in high-dose males in the JISA (1993) study; although no significant increase was observed through pair-wise comparisons for female rats in the latter study, a significantly positive exposure-response trend was reported. F344 rats, the strain in which the MCL occurred, have high background rates of the disease. The increase was still considered to be an increase over background rates, because MCL levels in exposed rats were clearly higher than those in both concurrent and historical controls, shorter latency periods for MCL development occurred in exposed rats and dose-related increases in disease severity were observed. This conclusion is further supported by other judgements of the relevance of tetrachloroethylene to MCL. Pathologists from the NTP (1986) study concluded that there was "clear evidence of carcinogenicity" for tetrachloroethylene in male F344 rats, based on MCL and renal tubular neoplasms, and "some evidence of carcinogenicity" in female F344 rats, based on MCL alone. Moreover, using a weight-of-evidence approach to analyze the incidence of MCL in rats from the NTP (1986) study, Thomas et al. (2007) concluded that MCL was definitively associated with tetrachloroethylene exposure, a classification that was given to only 5 of the 34 substances studied by NTP (1986) that had increased MCL. The relevance of rat MCL to humans has been questioned (Caldwell, 1999; Ishmael and Dugard, 2006). Although rat MCL is potentially of limited relevance to humans, the disease is similar to the rare human natural killer large granular lymphocyte leukaemia (Thomas et al., 2007). Very little is known about the MOA of MCL in rats; however, the weight of evidence related to physiological differences between humans and rats and the weak associations between tetrachloroethylene exposure and leukaemia in humans support the lack of human relevance for this endpoint. For this reason, the dose-response assessment for MCL was performed, but is considered to hold a lessened weight in the overall risk assessment, because the dose-response relationship in genetically predisposed rats might not be quantitatively relevant to humans without this genetic predisposition (Ishmael and Dugard, 2006).

Results for the cancer dose-response modelling for tetrachloroethylene exposure are presented in Table 3. Hepatic oxidative metabolism rate was selected as the most relevant dose metric for hepatocellular tumours, because the generation of oxidative metabolites in the liver is likely most closely associated with the tumours. Because the MOA for MCL is unknown, the most appropriate dose metric cannot be selected, so the most conservative dose metric (TCA concentration in blood) was selected. Although none of the models had adequate fit with the data for hepatocellular tumours from the JISA (1993) study, the value calculated for the NTP (1986) study is considered to be appropriate, as external points of departure (PODs) for other internal dose metrics were similar for the two studies, with the NTP (1986) results being slightly more conservative.

Table 3. Summary of dose-response modelling results for cancer endpoints
Endpoint, species (reference) POD Dose metric Dose-response model External POD (mg/kg bw per day)
Hepatocellular tumours, male mice (JISA, 1993) 10-5 Hepatic oxidative metabolism rate All models were rejected based on lack of fit; no external POD
Hepatocellular tumours, male mice (NTP, 1986) 10-5 Multistage 0.000 2
10-6 0.000 02
BMD10 6.2
BMDL10 1.7
MCL, male rats (JISA, 1993) 10-5 TCA concentration in blood Multistage 0.000 4
10-6 0.000 04
Stage 3 MCL, male rats (NTP, 1986) 10-5 Multistage 0.000 4
10-6 0.000 04

These results indicate that a "de minimis" (essentially negligible) excess cancer risk of 10-5 is 0.0002 mg/kg bw per day for hepatocellular tumours and 0.0004 mg/kg bw per day for MCL. Because the weight of evidence for tetrachloroethylene tends to suggest that a non-mutagenic mode of action is predominant for hepatocellular tumours, the threshold approach can also be used for the risk assessment. Using this approach, the tolerable daily intake (TDI) is calculated by identifying the most appropriate POD and dividing by relevant uncertainty factors. Relevant PODs for the threshold approach for tetrachloroethylene are the BMD10 of 6.2 mg/kg bw per day and the BMDL10 of 1.7 mg/kg bw per day.

9.5.2 Dose-response assessment for non-cancer endpoints

For effects other than cancer, a TDI can be derived by considering all studies and selecting the critical effect that is most relevant or occurs at the lowest dose, selecting a dose (or POD) at which the critical effect either is not observed or would occur at a relatively low incidence (e.g., 10%) and reducing this dose by an uncertainty factor to reflect the differences between study conditions and conditions of human environmental exposure.

The non-cancer risk assessment is based on neurotoxicity. Neurotoxic endpoints have been demonstrated to be relevant to humans in epidemiological studies performed in environmental and occupational settings and in acute controlled dosing studies; furthermore, the effects are more conservative than those for other non-cancer endpoints (hepatic, renal and reproductive/developmental effects). Neurotoxicity was observed in experimental animal studies, and these were initially considered in the dose-response assessment, but the epidemiological studies are more relevant to humans and demonstrate neurotoxicity at lower levels than in the animal studies. The non-cancer risk assessment is thus based on human studies. Most of the studies were excluded from the dose-response assessment process, for the following reasons:

  • characteristics of control groups, including socioeconomic status, were different from exposed (Schreiber et al., 2002; New York State Department of Health, 2010; Storm et al., 2011);
  • no control group was included (Gobba et al., 1998);
  • only a single exposure group was included, precluding measurement of dose-response relationships (Ferroni et al., 1992; Altmann et al., 1995);
  • exposure estimates were limited to concentrations in water, with no estimates of water intake for the participants (Janulewicz et al., 2008, 2012; Aschengrau et al., 2011, 2012; Getz et al., 2012);
  • exposures were acute and higher than in other studies (Stewart et al., 1961, 1970, 1977, 1981; Altmann et al., 1990, 1992); or
  • adverse effects were absent (Nakatsuka et al., 1992), limited to subjective symptoms (Lauwerys et al., 1983; Cai et al., 1991) or occurred in groups with low exposure but not high exposure (Seeber, 1989).

After excluding these studies, only two remained for consideration in the dose-response assessment. The first study (Cavalleri et al., 1994) demonstrated a NOAEL of 4.8 ppm for colour confusion—considered to be an adverse effect—as decreases in CCI scores were not significant in ironers (the lower exposure group, with an average exposure level of 4.8 ppm). A multivariate analysis confirmed that the effect was most likely due to tetrachloroethylene and not other characteristics that can affect CCI score (e.g., age, alcohol, smoking). Health Canada calculated BMDL10 values using the summary statistics for CCI scores (average and standard deviation for the CCI score and exposure for controls, ironers and dry cleaners), which resulted in a BMD10 of 7.2 ppm and a BMDL10 (using a 95% lower confidence limit) of 6.6 ppm using the power model (it should also be noted that using a benchmark response level of 1 standard deviation resulted in a similar BMDL of 6.7 ppm using this approach). One weakness of the study was that some overlap in exposure occurred between the two exposure groups (range of exposures, dry cleaners: 0.38-31.19 ppm; ironers: 0.52-11.28 ppm); however, the authors also found a significant correlation between workers' time-weighted average exposure levels and their CCI scores, further supporting a dose-response relationship. Using these regression data, Health Canada calculated a BMD10 of 10.1 ppm and a BMDL10 of 8.1 ppm using a linear model. However, for the latter approach, no controls were plotted in the graph presented in the original paper. Because of this weakness, the BMDL10 from the former modelling approach was used as the relevant POD for neurological endpoints. The regression approach also identified that the association was driven primarily by three workers with exposures > 12.5 ppm, presenting a weakness in the use of the study in the dose-response assessment, as the association between the CCI score and tetrachloroethylene exposure levels was no longer significant when these subjects were removed from the analysis (Cavalleri et al., 1994; Benignus et al., 2009).

The other study that was not excluded was the neurobehavioural study of dry cleaners performed by Echeverria et al. (1995). Although this study can be used to qualitatively support the Cavalleri et al. (1994) study, it was not included in the quantitative dose-response assessment because associations between exposure and neurobehavioural (visuospatial and visuomotor) effects were observed only for the 3-year or lifetime indices and not for current exposure levels. The only exposure data provided in the study were for current exposures—which were also limited, because they were based on 15-minute samples from a subsample of participants—and no associations were observed between current exposure levels and neurobehavioural effects. Because details of participants' exposure were provided only for the current exposures, for which there were no significant changes, the exposure indices from the Echeverria et al. (1995) study cannot be used for performing quantitative dose-response assessments. This would not impact the dose-response assessment, as the current exposure levels in the Echeverria et al. (1995) study were higher than those in the Cavalleri et al. (1994) study; therefore, the Cavalleri et al. (1994) study would still have driven the risk assessment, even if the Echeverria et al. (1995) study were considered quantitatively.

PBPK modelling was performed using the NOAEL and BMDL10 values from Cavalleri et al. (1994) to extrapolate from inhalation exposures to equivalent oral doses. Because the parent compound is suspected to be the toxic moiety for neurological effects and peak exposure levels for solvents are often more relevant for neurological effects, the relevant dose metrics were the daily peak and average concentrations of tetrachloroethylene. No brain compartment was included in the PBPK model, but partitioning coefficients for tetrachloroethylene in brain are similar to those in kidney (Dallas et al., 1994); therefore, the daily peak and average concentrations of tetrachloroethylene in the kidney were used as proxies for brain values. Peak and average concentrations of tetrachloroethylene in blood were also considered. Results of this modelling are summarized in Table 4. Peak exposures to tetrachloroethylene are thought to be most closely associated with adverse neurological effects; therefore, a POD of 4.7 mg/kg bw per day was selected.

Table 4. Dose-response modelling for neurological effects from Cavalleri et al. (1994)
Point of departure Measure Peak in blood Peak in kidneyTable 4 footnote a Average in blood Average in kidney

Table 4 footnotes

Table 4 footnote a

In the absence of a brain compartment in the PBPK model, kidney was used as a proxy for brain exposure, as described above.

Return to Table 4 footnote a referrer

NOAEL Internal dose 0.23 mg/L 1.18 mg/L 0.087 mg/L 0.44 mg/L
External dose 3.4 mg/kg bw per day 3.4 mg/kg bw per day 1.3 mg/kg bw per day 1.3 mg/kg bw per day
BMDL10 Internal dose 0.32 mg/L 1.62 mg/L 0.119 mg/L 0.602 mg/L
External dose 4.7 mg/kg bw per day 4.7 mg/kg bw per day 1.7 mg/kg bw per day 1.7 mg/kg bw per day

The TDI derived for neurological endpoints is also considered to be protective of hepatic, renal and reproductive/developmental non-cancer effects. These endpoints were also considered in a detailed dose-response assessment, using the most conservative results from animal studies as the PODs and applying the PBPK model to extrapolate between species and routes of exposure (Health Canada, 2013). Using BMDL10 values (using a 95% lower confidence limit) for liver necrosis and liver degeneration in mice and kidney tubule nuclear enlargement in rats (NTP, 1986; JISA, 1993), the TDIs for liver and kidney effects were calculated as 176 µg/kg bw per day and 52 µg/kg bw per day, respectively. For reproductive and developmental endpoints, a PBPK model to estimate rat foetal exposures to tetrachloroethylene was developed; however, due to a lack of data on the pharmacokinetics of the compound in pregnant women, the model could not be used for extrapolation to humans. Using default assumptions to extrapolate from experimental animals to humans and from inhalation to ingestion, the TDI for reproductive and developmental effects based on decreased foetal weight in rats (Carney et al., 2006) was 210 mg/kg bw per day. The PBPK model was applied to estimate foetal exposure, and comparisons with other non-cancer endpoints were made using internal dose metrics; foetal tetrachloroethylene exposure levels at the POD were estimated to be on the same order of magnitude as those for kidney effects. The TDI for neurological effects was more conservative than the other non-cancer endpoints, and the effects have been directly observed at relevant concentrations in the human population; therefore, neurological effects were considered as the most relevant non-cancer endpoint for the tetrachloroethylene risk assessment.

Immunological effects were not considered in the dose-response assessment. Although enhanced allergic reactions were observed in mice exposed to doses as low as 2 µg/kg bw per day (Seo et al., 2012), the study involved passively sensitizing mice prior to exposure, which leads to difficulties in including the study in a quantitative dose-response assessment. This factor, combined with some study limitations, led to the exclusion of this study for further consideration in the non-cancer risk assessment.

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