Guidelines for Canadian Drinking Water Quality: Supporting Documentation - Trichloroethylene

5.0 Analytical Methods

Several methods can be used to measure TCE in drinking water. Due to TCE's volatility, analytical methods for the chemical are based on purge and trap or head space gas chromatography using photoionization or mass spectrometric detection. TCE can also be captured by liquid-liquid extraction followed by gas chromatography with electrolytic conductivity detection.

Four methods for measuring TCE in drinking water have been approved by the U.S. Environmental Protection Agency (EPA). EPA Method 502.2, which employs purge and trap capillary gas chromatography with photoionization detectors and electrolytic conductivity detectors in series, has a detection limit in the range 0.01-3.0 µg/L (U.S. EPA, 1999b). EPA Method 524.2, which uses purge and trap capillary gas chromatography with mass spectrometric detectors in series, has a detection limit of 0.5 µg/L (U.S. EPA, 1999b). EPA Method 503.1 employs purge and trap capillary gas chromatography with photoionization conductivity detectors and has a detection limit of 0.01-3.0 µg/L (U.S. EPA, 1999b). EPA Method 551.1 uses liquid-liquid extraction and gas chromatography with electron capture detectors; this method has a method detection limit of 0.01 µg/L (U.S. EPA, 1999b).

For the determination of TCE in water, the practical quantitation limit (PQL) considered to be achievable by most good laboratories is 5 µg/L.

6.0 Treatment Technology

6.1 Municipal

Municipal water filtration plants that rely on conventional water treatment techniques (coagulation, sedimentation, precipitative softening, filtration and chlorination) have been found to be ineffective in reducing concentrations of TCE in drinking water (Robeck & Love, 1983). Two common water treatment technologies that, when combined, are effective in removing TCE are air stripping and activated carbon.

Air stripping has been found to be effective in removing VOCs such as TCE from groundwater. Air stripping is effective in stripping large quantities of TCE from water, but at low rates of removal (Russell et al., 1992).

Adsorption onto activated carbon is widely used to remove synthetic organic compounds such as TCE from drinking water, if the activated carbon filter bed is deep enough (Russell et al., 1992).

Combining air stripping and activated carbon into a two-step treatment train has been shown to improve TCE removal. In a municipal-scale treatment plant combining these processes, air stripping is used in the first step for bulk removal of the majority of the TCE from the water, and activated carbon is used in the second step to remove most residual TCE from the water. TCE levels below 1 µg/L can be achieved in municipal drinking water supplies using these methods (U.S. EPA, 1985b).

6.2 Residential

Generally, it is not recommended that drinking water treatment devices be used to provide additional treatment to municipally treated water. In cases where individual households obtain drinking water from private wells or the drinking water is contaminated by low concentrations of TCE, private residential water treatment devices may be an option for removing TCE from drinking water.

A number of residential treatment devices from various manufacturers are available that are affordable and can remove TCE from drinking water to make it compliant with the applicable guidelines or regulations. Filtration systems may be installed at the faucet (point of use) or where water enters the home (point of entry). Point-of-entry systems are preferred for VOCs such as TCE because they provide treated water for bathing and laundry as well as for cooking and drinking. Certified point-of-use treatment devices are currently available for the reduction of VOCs, including TCE. Although certified point-of-entry treatment devices cannot be purchased off the shelf, systems can be designed and constructed with certified materials. Periodic laboratory testing should be conducted on both the water entering a treatment device and the water it produces to verify that the treatment device is effective.

Health Canada does not recommend specific brands of drinking water treatment devices, but it strongly recommends that consumers look for a mark or label indicating that the device has been certified by an accredited certification body as meeting the appropriate NSF International (NSF)/American National Standards Institute (ANSI) standards. These standards have been designed to safeguard drinking water by helping to ensure the material safety and performance of products that come into contact with drinking water. Certification organizations provide assurance that a product or service conforms to applicable standards. In Canada, the following organizations have been accredited by the Standards Council of Canada (www.scc.ca) to certify drinking water devices and materials as meeting NSF/ANSI standards:

• Canadian Standards Association International (www.csa-international.org);
• NSF International (www.nsf.org);
• Water Quality Association (www.wqa.org);
• Underwriters Laboratories Inc. (www.ul.com);
• Quality Auditing Institute (www.qai.org); and
• International Association of Plumbing and Mechanical Officials (www.iapmo.org).

Treatment devices to remove TCE from untreated water (such as from a private well) should be certified for the removal of either TCE or VOCs. In the case of TCE, these treatment devices are certified to reduce TCE levels from an average influent (challenge) concentration of 0.3 mg/L to a maximum finished effluent concentration of 0.005 mg/L or less (NSF International, 2005). Treatment devices that are certified to remove TCE or VOCs incorporate some type of adsorption technology, usually activated carbon, or utilize reverse osmosis, usually in combination with one or more adsorption-type filters.

7.0 Kinetics and Metabolism

7.1 Absorption

TCE is readily absorbed following both oral and inhalation exposures. Dermal absorption is also possible, but information pertaining to this route of exposure is limited. Significant inter- and intraspecies variabilities in TCE absorption following all routes of exposure have been well documented.

TCE is rapidly and extensively absorbed from the gastrointestinal tract into the systemic circulation in animals. Mass balance studies using radiolabelled TCE indicated that mice and rats metabolized TCE at 38-100% and 15-100%, respectively, following oral administration in corn oil vehicle. For both species, the lower values were obtained following treatment with large doses in excess of 1000 mg/kg bw, implying that the rate of absorption was higher at low doses than at high doses in both species (Daniel, 1963; Parchman and Magee, 1982; Dekant and Henschler, 1983; Dekant et al., 1984; Buben and O'Flaherty, 1985; Mitoma et al., 1985; Prout et al., 1985; Rouisse and Chakrabarti, 1986). Different vehicles affect the rate of absorption, with the rate being almost 15 times greater following dosing in water than following administration in corn oil. Overall, absorption of TCE through the gastrointestinal tract is considerable and, at very low concentration, nearly complete. Although human exposure studies investigating oral absorption of TCE were not identified, numerous case studies of accidental or intentional ingestion of TCE suggest that absorption of TCE from the gastrointestinal tract in humans is likely to be extensive (Kleinfeld and Tabershaw, 1954; DeFalque, 1961; Bruning et al., 1998).

Pulmonary uptake of TCE into the systemic circulation is rapid in animals, but blood:gas partition coefficients in rodents vary across species, strain and gender (Lash et al., 2000). After inhalation exposure to radiolabelled TCE at 10 or 600 ppmv over a 6-hour period, net pulmonary uptake was 10 times greater at the higher concentration than at the lower concentration in rats, whereas it was similar at both exposure concentrations in mice (Stott et al., 1982). In humans, TCE is rapidly and extensively absorbed by the lungs and into the alveolar capillaries. The blood:air partition coefficient of TCE has been estimated to be approximately 1.5- to 2.5-fold lower in humans than in rodents (Sato et al., 1977; Monster, 1979; Clewell et al., 1995). Under non-steady-state conditions, TCE pulmonary uptake is rapid during the first 30-60 minutes of exposure, decreasing significantly as TCE concentrations in tissues approach steady state (Fernandez et al., 1977; Monster et al., 1979).

Dermal absorption has been demonstrated in mice (Tsuruta, 1978) and guinea pigs (Jakobson et al., 1982). Dermal absorption has also been demonstrated in human volunteers (Stewart and Dodd, 1964; Sato and Nakajima, 1978); however, variability between individuals precludes any meaningful interpretation of these data.

7.2 Distribution

Once absorbed, TCE diffuses readily across biological membranes and is widely distributed to tissues and organs via the circulatory system. Studies in animals (e.g., Fernandez et al., 1977; Dallas et al., 1991; Fisher et al., 1991) and humans (De Baere et al., 1997) have found TCE or its metabolites in most major organs and tissues. Primary sites of distribution include the lungs, liver, kidneys and central nervous system. TCE may accumulate in adipose tissue because of its lipid solubility. Consequently, slow release of TCE from adipose stores might act as an internal source of exposure, ultimately resulting in longer mean residence times and bioavailability of TCE (Fernandez et al., 1977; Dallas et al., 1991; Fisher et al., 1991). Age-dependent factors may influence TCE distribution in humans, suggesting greater susceptibility to TCE in children than in adults (Pastino et al., 2000).

7.3 Metabolism

TCE metabolism occurs primarily in the liver, although it may also occur in other tissues, particularly the kidney. There are two main pathways responsible for TCE metabolism: oxidation by cytochrome P-450 and conjugation with glutathione (GSH) by glutathione-S-transferases (GSTs) (OEHHA, 1999; Lash et al.., 2000). In the liver, TCE is metabolized by cytochrome P-450 enzymes to an epoxide intermediate, which spontaneously rearranges to chloral. Chloral is further metabolized to trichloroethanol (TCOH), trichloroethanol glucuronide (TCOG) and trichloroacetic acid (TCA) as the principal metabolites. Under certain conditions, TCE-epoxide forms dichloroacetyl chloride, which rearranges to dichloroacetic acid (DCA). Other minor metabolites include carbon dioxide, N-(hydroxyacetyl)aminoethanol and oxalic acid, all believed to be products of hydrolysis of a TCE-epoxide intermediate (Goeptar et al., 1995).

In the conjugation pathway, the reactive electrophilic species produced through the oxidation are deactivated by conjugation to the nucleophilic sulphur atom of GSH. This may be catalysed by various cytosolic and microsomal GSTs or may occur spontaneously via a non-enzymatic addition/elimination reaction. The resulting conjugates undergo further metabolism to yield various metabolites, the most important of which are mercapturic acids, which are rapidly excreted in urine (Goeptar et al., 1995).

The oxidative metabolism of TCE takes place primarily in the liver, although it may occur to some extent in various other tissues, such as the lung (Lash et al., 2000). Four isozymes of cytochrome P-450 (primarily CYP2E1) oxidize TCE (OEHHA, 1999; Lash et al., 2000). An intermediate electrophilic epoxide (2,2,3-trichlorooxirane, or TCE-oxide) is suspected to form during oxidative metabolism, although it is not known whether TCE-oxide exists in free form (Lash et al., 2000). TCE-oxide may be metabolized by several pathways, the predominant pathway being spontaneous rearrangement to chloral, which is then hydrated to chloral hydrate (CH) (OEHHA, 1999). CH is metabolized to TCA, which is the main TCE metabolite in the blood, and TCOH. TCA and TCOH may be further metabolized to DCA and TCOG, respectively.

The GSH conjugation by GST also occurs primarily in the liver, although several other tissues (kidney, biliary tract and intestines) are involved (Lash et al., 2000). The GSH conjugation reactions occur more slowly than the cytochrome P-450-catalysed oxidation reactions. TCE is converted by GST to S-(1,2-dichlorovinyl) glutathione (DCVG), which is excreted into the bile, then reabsorbed through enterohepatic circulation and converted to the cysteine conjugates S-(1,1-dichlorovinyl)-L-cysteine (1,1-DCVC) and S-(1,2-dichlorovinyl)-L-cysteine (1,2-DCVC) (Lash et al., 2000; Clewell et al., 2001). 1,1-DCVC may undergo N-acetylation and be excreted in the urine or metabolized by a lyase enzyme to reactive metabolites, including a thioacetaldehyde, whereas 1,2-DCVC may be metabolized by N-acetyltransferase and excreted in the urine or converted by β-lyase to reactive metabolites, including a thioketene (Clewell et al., 2001). Therefore, exposure to TCE clearly results in exposure of tissues to a complex mixture of metabolites (OEHHA, 1999; U.S. EPA, 2001a).

Enterohepatic circulation of TCOG is believed to play a very important role in maintaining levels of TCA, which has a major impact on dosimetry and the very high clearance of TCE seen at low doses by first-pass metabolism in the liver (Stenner et al., 1997, 1998; Barton et al., 1999). This appears to control the low-dose behaviour of the metabolites, essentially favouring the oxidative metabolites. It is one of the reasons why the GSH pathway does not seem to contribute much to the clearance of TCE at low doses. Since the oxidative metabolites are clearly responsible for the effects on the liver (both cancer and non-cancer), this implies that the oral route is most importantly related to liver effects, whereas other routes may preferentially affect other organs (e.g., kidney) (discussed in a later section).

There are several interspecies differences in TCE metabolism. For example, human hepatic microsomes possess less activity towards TCE than rat or mouse hepatic microsomes (Nakajima et al., 1993), and humans are less efficient at metabolizing TCE than rodents. Furthermore, a comparison of renal β-lyase activities in the kidney indicates that rats are more efficient than humans at metabolizing DCVC to reactive metabolites (Clewell et al., 2000). There are also intraspecies differences. In humans, interindividual variations in enzyme expression and activity, such as individual variation in activities of CYP1A2 and CYP2E1, for example, have been observed. As well, males generally have higher GSH conjugation rates than females, and genetic polymorphisms may influence GSH conjugation rates in humans (Lash et al., 2000).

The major metabolism of DCA occurs through GSH transferase (zeta), a family of cytosolic enzymes. DCA's rates of metabolism are very high compared with those of TCA and TCE, explaining why it is difficult to generate sufficient concentrations in vivo to measure. However, TCA is unlikely to be responsible for human liver cancer at the levels that are encountered in the environment, based on its mode of action as a peroxisome proliferator and because it produced liver tumours only in mice despite being adequately tested in rats (DeAngelo et al., 1997). One of the issues of most concern with TCE is its conversion to DCA. The relative contributions of DCA and TCA to liver tumours in mice have recently been discussed (Chen, 2000). A recent paper (Bull et al., 2002) strongly suggests that DCA does contribute to the liver cancer response in mice. DCA is clearly carcinogenic in both mice and rats, and its mode of action is clearly different from that of TCA. Therefore, DCA cannot be dismissed as a potential human carcinogen. It is, however, apparent that while DCA may be formed during the metabolism of TCE, it is very unlikely to be produced in significant amounts at environmental levels of exposure to TCE.

7.4 Excretion

The database pertaining to the elimination of TCE is large, and TCE clearance is well characterized in both animals and humans. Although the elimination kinetics of TCE and its metabolites vary by route of exposure, elimination pathways appear to be similar for ingestion and inhalation. No data regarding the elimination of TCE and its metabolites following dermal exposures were found.

TCE is eliminated either unchanged in expired air or by metabolic transformation with subsequent excretion, primarily in urine, as TCA, TCOH or TCOG (following oxidative metabolism) or as DCVG or the cysteine conjugate N-acetyl-S-dichlorovinyl-L-cysteine (DCVCNac) (following GSH conjugation). Studies in human volunteers have shown that following TCE exposure, urinary TCOH is first produced more quickly and in larger amounts than urinary TCA. However, over time, TCA production eventually exceeds that of TCOH (Nomiyama and Nomiyama, 1971; Muller et al., 1974; Fernandez et al., 1975; Sato et al., 1977; Monster and Houtkooper, 1979; Monster et al., 1979). Small amounts of metabolized TCE are excreted in the bile or as TCOH in exhaled air. TCE may also be excreted in breast milk (Pellizzari et al., 1982; Fisher et al., 1987, 1989).

Comparative studies have found that elimination is more rapid in mice than in rats (Lash et al., 2000). However, the formation of the more toxic metabolite TCA is also approximately 10 times faster in mice than in rats. Therefore, differential elimination kinetics help explain interspecies differences in toxicity and toxicokinetics associated with TCE, given that the toxicity of TCE is linked to the formation of its metabolites (Parchman and Magee, 1982; Stott et al., 1982; Dekant et al., 1984; Buben and O'Flaherty, 1985; Mitoma et al., 1985; Prout et al., 1985; Rouisse and Chakrabarti, 1986). In humans, interindividual heterogeneity was seen in the metabolism and elimination of TCE (Nomiyama and Nomiyama, 1971; Fernandez et al., 1975; Monster et al., 1976).