a Henry’s law constant is defined as the concentration in air divided by the concentration in water at equilibrium; unit of Henry’s law constant: dimensionless (Moore et al., 1995).

Exposure to methyl chloride can also be monitored in air by a direct-reading infrared analyser, at minimum detectable concentrations of 800–3100 µg/m³ (390– 1500 ppb) (IARC, 1985).

Stratospheric air samples are often concentrated by a cryogenic procedure, at liquid nitrogen or argon tem perature, followed by GC analysis employing electron capture detection (Rasmussen et al., 1980; Singh et al., 1983, 1992; Rudolph et al., 1992, 1995; Khalil & Rasmussen, 1993; Fabian et al., 1996) or with GC/mass spectrometry (MS) (Schauffler et al., 1993). The GC may be equipped with a flame ionization detector (Evans et al., 1992) or a mass selective detector (Atlas et al., 1993). Almasi et al. (1993) described a modified version of a method commonly used by the US Environmental Protection Agency (EPA) to analyse low levels of volatile organic compounds in air (EPA Method TO-14). It includes sample concentrations on glass beads at -160 °C, thermal desorption, separation on a GC capil lary column, and detection with ion trap MS. The detec tion limit is about 0.06 µg/m³ (0.03 ppb) for methyl chloride.

In water, methyl chloride can be analysed by EPA Method 502.2 at a detection limit of 0.1 µg/litre (US EPA, 1986b). Other methods for the detection of volatile organic substances in water are EPA Method 502.1, which has a detection limit of 0.01 µg/litre (US EPA, 1986b), and EPA Method 524.2, with a detection limit of 0.05 µg/litre (US EPA, 1986b). Another method involv ing a solid-phase micro-extraction technique has a detec tion limit of <25 µg/litre (Shirey, 1995).

EPA Method 601 (purgeable halocarbons) is suit able for measuring methyl chloride in wastewater. The detection limit is 0.06 µg/litre (US EPA, 1982; CFR, 1990). A similar method is EPA Method 624 (purgeables), with a detection limit of 2.8 µg/litre (US EPA, 1982; CFR, 1991). A third method used for analysis of wastewater is EPA Method 1624, with a minimum detection level of 50 µg/litre (CFR, 1991).

In soil and solid waste, EPA Method 5030 (US EPA, 1986a) may be used to analyse methyl chloride. Analysis is performed by different EPA methods. In Method 8010B, the limit of detection is 12.5 µg/kg for high-concentration soils and sludges (US EPA, 1986a). Gomes et al. (1994) describes another method to collect and analyse methyl chloride, which can also be used to analyse contaminants in groundwater.

Natural sources of methyl chloride dominate over anthropogenic sources. The major source appears to be the marine/aquatic environment, likely associated with algal growth. Other sources are biomass burning (forest fires), degradation of wood by fungi, and direct and indirect anthropogenic sources.

Methyl chloride is produced industrially by reac tion of methanol and hydrogen chloride or by chlorina tion of methane (Key et al., 1980; Edwards et al., 1982a; Holbrook, 1992). In almost all of the commercial uses, methyl chloride is reacted to form another product (ATSDR, 1998). The current principal uses are in the production of silicones and also as a general methylating agent. The use of methyl chloride in the manufacture of synthetic rubber, its refrigerant and extractant applica tions, and its use as a tetramethyllead intermediate now have secondary importance (Holbrook, 1992).

Indirect sources of methyl chloride are tobacco smoke, turbine exhaust (Wynder & Hoffmann, 1967; Graedel, 1978; Häsänen et al., 1990), incineration of municipal and industrial waste (Graedel & Keene, 1995), chlorination of drinking-water, and sewage effluent (Abrams et al., 1975).

The current production capacity of methyl chloride in the USA has been estimated to be about 0.417 × 10⁶ tonnes per year (CMR, 1995). The production in Japan in 1996 was 0.13 × 10⁶ tonnes (Chemical Daily Co. Ltd., 1998).

It has been concluded that well over 90%, and perhaps as much as 99%, of ambient air concentrations on a global scale appear to originate from natural sources rather than from anthropogenic sources (ATSDR, 1998). Edwards et al. (1982b) estimated the emissions of methyl chloride during production, transport, storage, and use to be approximately 0.02 × 10⁶ tonnes per year, corres ponding to nearly 6% of the amount produced. Accord ing to this estimate, anthropogenic sources would account for 1–2% of the total release, including natural sources. Other estimates of the global yearly release from anthropogenic sources are in the range of 0.024– 0.6 × 10⁶ tonnes (Watson et al., 1980; Gribble, 1992; Dowdell et al., 1994), the higher estimate including indirect anthropogenic sources and possibly also biomass burning.

Methyl chloride, which is the most prevalent halo genated methane in the atmosphere, is present in the troposphere at a concentration of about 1.2 µg/m³ (0.6 ppb) (WMO, 1994). It has been calculated that at a production rate of about 3.5 × 10⁶ tonnes per year, the steady-state mixing ratio of 1.2 µg/m³ (0.6 ppb) is maintained given an atmospheric lifetime in the order of 2 years (WMO, 1994). Estimates of the total global annual release of methyl chloride from all sources are around 5 × 10⁶ tonnes (Rasmussen et al., 1980; Logan et al., 1981; Edwards et al., 1982b; Dowdell et al., 1994; WMO, 1994; Fabian et al., 1996). According to ATSDR (1998), the total release from all sources amounts to approximately 3.2–8.2 × 10⁶ tonnes per year.

Over the Pacific Ocean, the concentration of methyl chloride is higher in the lower troposphere than in the higher layers. However, over the continents, the concentration is independent of the altitude. Thus, the ocean seems to be a source of methyl chloride (Geckeler & Eberhardt, 1995). In the oceans, algae, especially planktonic algae, are considered to be responsible for most of the methyl chloride production. However, this has not been fully proven. Phytoplankton have been shown to produce methyl chloride in laboratory studies (Tait & Moore, 1995). An alternative model is that methyl chloride is formed as a result of exchange processes between methyl iodide and chlorine ions in seawater (Isidorov, 1990). Estimates of the global yearly release of methyl chloride from marine sources are in the range 1–8 × 10⁶ tonnes (Watson et al., 1980; Singh et al., 1983; Isidorov, 1990).

Terrestrial species also produce methyl chloride. The activity of methyltransferases, believed to be responsible for methyl chloride production, has been observed in several herbaceous species (Saini, 1995). According to Harper et al. (1988), 34 species of fungi are also known to biosynthesize methyl chloride.

Estimates of the global annual release of methyl chloride from biomass burning are in the range 0.4–1.8 × 10⁶ tonnes (Watson et al., 1980; Andreae, 1991, 1993; Lobert et al., 1991; Rudolph et al., 1994, 1995). The major part of the methyl chloride released from biomass burning originates from forest fires in the tropics (Andreae et al., 1994). The estimated global release of methyl chloride from temperate and boreal biomass fires has been calculated to be 0.012 × 10⁶ tonnes per year (Laursen et al., 1992). Low-intensity, inefficient combus tion and high chlorine content of the biomass promote methyl chloride formation (Reinhardt & Ward, 1995).

Most methyl chloride discharged to the environ ment will be released to air. The principal sink for methyl chloride in the troposphere is chemical reaction with hydroxyl radicals (ASTDR, 1990; Graedel & Keene, 1995; Fabian et al., 1996). The rate constant for this reaction is approximately 4.3 × 10^–14 cm³/s per molecule at 25 °C (NASA, 1981; Atkinson, 1985). Atmospheric lifetime estimates range from 1 to 3 years (Atkinson, 1985; BUA, 1986; Warneck, 1988; ATSDR, 1990; WMO, 1990, 1994; Fabian et al., 1996; Houghton et al., 1996). Surface deposition, rainout, and washout are unimportant sinks for methyl chloride (Graedel & Keene, 1995).

Estimates of the amount of methyl chloride reaching the stratosphere vary considerably. Borchers et al. (1994) claim that the contribution of methyl chloride to the stratospheric chlorine budget is significant. Crutzen & Gidel (1983) estimated the flux of methyl chloride to the stratosphere to be about 2 × 10⁶ tonnes per year or 20–25% of the total annual stratospheric chlorine input. According to Fabian et al. (1996), only a fraction (less than 10%) of the amount of methyl chlor ide emitted reaches the stratosphere. Edwards et al. (1982b) claim that about 6% of the methyl chloride released reaches the stratosphere (corresponding to 0.3 × 10⁶ tonnes per year). According to Graedel & Crutzen (1993) and Graedel & Keene (1995), only 0.8% (corres ponding to 0.03 × 10⁶ tonnes chlorine per year) is expected to reach the stratosphere. A global budget for tropospheric methyl chloride is shown in Figure 1.

The ability of the ozone layer to absorb ultraviolet radiation shorter than 290 nm should exclude direct photolysis in the troposphere, because methyl chloride does not absorb any radiation above 290 nm (BUA, 1986). In the stratosphere, photodissociation will occur at a rate approximately equal to its reaction with hydroxyl radicals (Robbins, 1976). The chlorine radicals that are generated contribute to ozone depletion. Methyl chloride has been shown to photochemically decompose at 185 nm. Photooxidation products in the gas phase were carbon dioxide, carbon monoxide, formic acid, formyl chloride, water, and hydrogen chloride (Gürtler & Kleinermanns, 1994).

The stratospheric steady-state ODP of methyl chloride has been determined to be 0.02 relative to CFC- 11 (ODP = 1) (Solomon et al., 1992; WMO, 1994; Fabian et al., 1996). Estimates of the amount of methyl chloride reaching the stratosphere, and thereby also its contribution to ozone depletion, vary considerably. However, as estimated from figures presented by WMO (1994), methyl chloride contributes approximately 15% (0.5 ppb) of the total (3.3 ppb) equivalent effective stratospheric chlorine. The term "equivalent effective stratospheric chlorine" includes both stratospheric chlorine and bromine (alpha¹ = 40) and also considers the dissociation rate of each compound involved in ozone depletion.

A radiative forcing value of 0.0053 W/m² per part per billion has been determined for methyl chloride. This value is about 2% of the forcing of CFC-11 and about 300 times the forcing of carbon dioxide, on a per molecule basis (Grossman et al., 1994). Houghton et al. (1996) gave the radiative forcing value as 0 W/m² for methyl chloride. The global warming potential (GWP) has been calculated to be about 25, relative to carbon dioxide (GWP = 1), at a time scale of 20 years (Grossman et al., 1994).

As the current concentration of methyl chloride in the atmosphere is relatively low, approximately 1.2 µg/m³ (0.6 ppb), the contribution of this substance to the greenhouse effect will not become a problem unless large releases of this gas occur (Grossman et al., 1994). WMO (1994) also considers that the contribution of methyl chloride to climate forcing is minimal.

The contribution of methyl chloride to the creation of photochemical air pollution is not significant because of its relatively low reactivity and low amounts emitted. The photochemical ozone creation potential (POCP) of methyl chloride has been determined to be 3.5 (integrated ozone formation over 5 days) relative to that of ethylene (POCP = 100) (Derwent et al., 1996).

Reactive chlorine in the lower atmosphere (as distinguished from chlorofluorocarbon-derived chlorine in the stratosphere) is supposed to be important in considerations of precipitation acidity, corrosion of metals and alloys, foliar damage, and chemistry of the marine boundary layer. The tropospheric reactive chlorine burden of approximately 8.3 × 10⁶ tonnes chlorine is dominated by methyl chloride (~45%) and trichloroethane (~25%) (Graedel & Keene, 1995).

If methyl chloride is released into water, it will be lost primarily by volatilization. The volatilization half- life has been calculated to be 2.1 h in a model river (Lyman et al., 1982). The volatilization half-lives of methyl chloride in a pond and in a lake have been estimated to be 25 h and 18 days, respectively, using the model EXAMS (ATSDR, 1990). The low log K_ow (0.91) of methyl chloride indicates that the substance does not tend to concentrate in sediments.

The transformation of methyl chloride by hydrol ysis is probably negligible under acid and neutral condi tions. Under basic conditions, slow hydrolysis takes place, yielding methanol as a transformation product (Simon, 1989). Hydrolytic half-lives range from 31 days (pH 11) to 2.5 years (pH not given) at 20–25 °C (Zafiriou, 1975; Mabey & Mill, 1976, 1978; Simon, 1989). The hydrolytic half-life of methyl chloride in seawater varies with temperature (0–30 °C) from 0.5 to 77 years (Elliott & Rowland, 1995). Laboratory data indicate that the photochemical transformation of methyl chloride in water is negligible (Mabey & Mill, 1976).

Methyl chloride was not readily biodegraded in a standardized "closed bottle test" (MITI, 1992). How ever, several isolated bacterial strains have been shown to degrade methyl chloride under both aerobic (Stirling & Dalton, 1979; Hartmans et al., 1986; Bartnicki & Castro, 1994; Chang & Alvarez-Cohen, 1996) and anaerobic conditions (Traunecker et al., 1991; Braus- Stromeyer et al., 1993; Dolfing et al., 1993; Leisinger & Braus-Stromeyer, 1995). A half-life of less than 11 days was estimated for the anaerobic biodegradation of methyl chloride in groundwater, based on laboratory data obtained under conditions favourable for anaerobic biodegradation (Wood et al., 1985).

The very low log K_ow (0.91) of methyl chloride indicates that it will not tend to adsorb to soil (Lyman et al., 1982). The adsorption coefficient, K_oc, has been calculated to be 5, based on physical/chemical data (ATSDR, 1990). The very high vapour pressure and low adsorption to soil suggest that methyl chloride present near the soil surface will rapidly be lost by volatilization (ATSDR, 1990; HSDB, 1996). As it is not expected to adsorb to soil, methyl chloride present in deeper soil layers may to some extent leach into the groundwater, as well as diffuse to the surface and volatilize (ATSDR, 1990; HSDB, 1996). In groundwater, methyl chloride is expected to biodegrade or hydrolyse very slowly (ATSDR, 1990; HSDB, 1996). The cumulative volatili zation loss of methyl chloride, from a depth of 1 m beneath ground, has been calculated to be at least 70% and 22% in 1 year for a sandy soil and a clay soil, respectively (Jury et al., 1990).

There are no experimental studies on bioaccumu lation. However, only a minor accumulation in biota would be expected on the basis of the low log K_ow. A bioconcentration factor of 2.9 has been calculated based on the log K_ow (ATSDR, 1990).

Background concentrations of methyl chloride in the troposphere are around 1.2 µg/m³ (0.6 ppb), ranging from about 1.0 to 1.4 µg/m³ (from 0.5 to 0.7 ppb) (Cox et al., 1976; Cronn et al., 1976, 1977; Pierotti & Rasmussen, 1976; Singh et al., 1977, 1979, 1983; Graedel, 1978; Khalil & Rasmussen, 1981, 1993; Guicherit & Schulting, 1985; Gregory et al., 1986; Warneck, 1988; Rudolph et al., 1992; Singh et al., 1992; Atlas et al., 1993; WMO, 1994; Graedel & Keene, 1995; Fabian et al., 1996). In the stratosphere, the concentra tion of methyl chloride decreases with altitude. Concen trations in the Arctic stratosphere in March 1992 ranged from 0.60 to 0.082 µg/m³ (from 0.29 to 0.04 ppb) at altitudes of 11–22 km (von Clarmann et al., 1995). In May 1985, at a latitude of 26–30 °N, Zander et al. (1992) found methyl chloride concentrations ranging from 0.12 to 0.050 µg/m³ (from 0.058 to 0.024 ppb) at altitudes of 12–22 km. Near the tropical tropopause (23.8–25.3 °N, 15–17 km altitude), the mean methyl chloride concentration was measured to be 1.1 µg/m³ (0.531 ppb) during January–March 1992 (Schauffler et al., 1993).

Numerous measurements of methyl chloride levels in air have been performed, especially in the USA. The mean or median concentrations of methyl chloride measured in the air of rural/remote sites in the USA were about 1.0–2.7 µg/m³ (0.5–1.3 ppb), with the majority of the values below 2.1 µg/m³ (1.0 ppb); the maximum concentration measured was 4.3 µg/m³ (2.1 ppb) (Grimsrud & Rasmussen, 1975; Robinson et al., 1977; Singh et al., 1977, 1981b; Brodzinsky & Singh, 1983; Rasmussen & Khalil, 1983; Shah & Singh, 1988). In samples from urban/suburban areas in the USA, the mean/median concentrations were in the range of 0.27– 6.2 µg/m³ (0.13–3.0 ppb), with the majority of the values in the range 1.0–2.3 µg/m³ (0.5–1.1 ppb); the highest concentration found was 25.0 µg/m³ (12.1 ppb) (Singh et al., 1977, 1979, 1981a, 1982, 1992; Brodzinsky & Singh, 1983; Edgerton et al., 1984; Shah & Singh, 1988; Rice et al., 1990; US EPA, 1991a, 1991b; Evans et al., 1992; Kelly et al., 1994; Spicer et al., 1996). Methyl chloride concentrations in three Japanese cities ranged from 4.5 to 35 µg/m³ (from 2.2 to 17 ppb) (Furutani, 1979). In Delft, the Netherlands, and Lisbon, Portugal, concentrations of 6.2 µg/m³ (3.0 ppb) (Guicherit & Schulting, 1985) and 4.5 µg/m³ (2.2 ppb) (Singh et al., 1979), respectively, were found.

From these data, it appears that the concentrations of methyl chloride are slightly higher in the air of urban/ suburban sites than at rural/remote sites. However, a direct comparison is difficult, because samples in urban/ suburban areas were probably often taken at ground level, while several measurements of rural/remote areas were made at higher altitudes.

Methyl chloride has also been occasionally detected in water, soil, and biota. A few studies on measurements of methyl chloride in drinking-water were identified, most of them performed in the USA and Canada (Abrams et al., 1975; Coleman et al., 1976; Burmaster, 1982; Mariich et al., 1982; Otson et al., 1982; Otson, 1987). A maximum concentration of 44 µg/litre was measured in a drinking-water well (Burmaster, 1982).

In measurements of groundwaters in the USA, concentrations of methyl chloride ranged from not detectable up to 100 µg/litre, found at a former waste site of a chemical factory (Page, 1981; Burmaster, 1982; Sabele & Clark, 1984; Lesage et al., 1990; Plumb, 1991; Rosenfeld & Plumb, 1991). The substance was detected in groundwaters at 20 of 479 waste disposal sites in 1991 (Plumb, 1991).

In surface water samples in North America, the concentrations ranged from not detectable up to 224 µg/litre, the highest value reported from New Jersey, USA, in the 1970s (Page, 1981; Otson et al., 1982; Great Lakes Water Quality Board, 1983; Granstrom et al., 1984; Staples et al., 1985; Otson, 1987). In the only European study found (Hendriks & Stouten, 1993), a maximum concentration of 12 µg/litre in the river Rhine was reported. In seawater samples collected near the surface, methyl chloride was mostly found at 0.01– 0.05 µg/litre (Lovelock, 1975; Pearson & McConnell, 1975; NAS, 1978; Singh et al., 1979, 1983; Edwards et al, 1982b); however, a higher concentration of 1.2 µg/li tre was reported from a measurement near the shore of California, USA (Singh et al., 1979).

Methyl chloride was detected in soils at 34 waste sites and in sediments at 13 waste sites in the USA (HazDat, 1998) and in 1 of 345 sampling stations of the US EPA STORET database, at a concentration of <5 µg/kg (Staples et al., 1985). Methyl chloride was also detected in soil at an electronic industrial site in São Paulo, Brazil (Gomes et al., 1994). No data were found on methyl chloride levels in sediment. According to the US EPA STORET database, methyl chloride was detected in 1% of analysed samples of fish and seafood (Staples et al., 1985).

Data given in section 6.1 suggest that humans are exposed to methyl chloride in ambient air. Background concentrations are around 1.2 µg/m³ (0.6 ppb). In urban areas, mean and median concentrations generally seem to be slightly higher, 1.0–2.3 µg/m³ (0.5–1.1 ppb). How ever, individual measurements may be much higher. The highest value found in the literature was 35 µg/m³ (17 ppb).

Workplace concentrations have been measured in four US chemical plants (NIOSH, 1980). Three of the plants produced methyl chloride. The personal 8-h time- weighted average concentrations in the three plants ranged from 18.4 to 25.6 mg/m³ (from 8.9 to 12.4 ppm), from <0.4 to 15.5 mg/m³ (from <0.2 to 7.5 ppm), and from <0.2 to 26.2 mg/m³ (from <0.1 to 12.7 ppm), respectively. In the fourth plant, where methyl chloride was used as a blowing agent in the production of poly styrene foam, the personal exposures ranged from 6.2 to 44.2 mg/m³ (from 3.0 to 21.4 ppm). In a Dutch methyl chloride production plant, individual 8-h time-weighted averages of methyl chloride exposure in the air ranged from 62 to 186 mg/m³ (from 30 to 90 ppm) (van Doorn et al., 1980).

The most important route of exposure of methyl chloride in humans is via the respiratory pathway. Data on methyl chloride toxicokinetics cover only inhalation; no relevant information on other routes of administration was located in the literature.

In humans as well as in experimental animals, methyl chloride is readily absorbed through the lungs following inhalation (Andersen et al., 1980; Stewart et al., 1980; Landry et al., 1983; Nolan et al., 1985; Löf et al., 2000). In human volunteers exposed to 21 or 103 mg methyl chloride/m³ (10 or 50 ppm) for 6 h or to 21 mg/m³ (10 ppm) for 2 h, steady state was reached during the first exposure hour (Nolan et al., 1985; Löf et al., 2000). In rats, equilibrium between uptake and elimination was also obtained within 1 h (Landry et al., 1983).

After rats were exposed to ¹⁴C-labelled methyl chloride by inhalation, radioactivity was found to the largest extent in the liver, kidneys, and testes and to a smaller extent in the brain and lungs (Redford-Ellis & Gowenlock, 1971; Kornbrust et al., 1982; Landry et al., 1983). The presence of residues was, however, attributed to the metabolism of methyl chloride to formaldehyde and formate and subsequent incorporation of the radio labelled carbon atom into tissue macromolecules through single-carbon anabolic pathways (Kornbrust & Bus, 1982; Kornbrust et al., 1982). Methyl chloride may also bind to macromolecules, especially protein, and to a minimal extent probably also to DNA (Kornbrust et al., 1982; Vaughan et al., 1993).

In humans as well as in animals, methyl chloride is mainly metabolized by conjugation with glutathione. S-Methylglutathione can then be further metabolized to S-methylcysteine and methanethiol (Redford-Ellis & Gowenlock, 1971; Bus, 1982; Landry et al., 1983). To a lesser extent, methyl chloride is also metabolized micro somally via cytochrome P-450 in rat liver, resulting in the formation of formaldehyde and formate (Kornbrust & Bus, 1983). Formaldehyde and formate may also be formed via the glutathione pathway (Kornbrust & Bus, 1983).

Inhalation of methyl chloride by male B6C3F1 mice resulted in a concentration-dependent depletion of glutathione in liver, kidney, and brain. The depletion was most pronounced in the liver, where a 6-h inhalation exposure to 206 mg/m³ (100 ppm) decreased the gluta thione level by 45%, and exposure to 5160 mg/m³ (2500 ppm) reduced the glutathione content to 2% of control levels (Kornbrust & Bus, 1984).

Metabolites from methyl chloride are excreted in the urine and in the expired air. S-Methylcysteine has been identified in the urine of occupationally exposed humans and rats (van Doorn et al., 1980; Landry et al., 1983), and formic acid has been found in rat urine (Kornbrust & Bus, 1983). Further, carbon dioxide has been shown to be the major final metabolite of methyl chloride, accounting for nearly 50% of the radiolabelled material recovered after a 6-h exposure of rats to methyl chloride (Kornbrust & Bus, 1983). Methyl chloride is also excreted unmetabolized via the lungs, as seen in studies in volunteers (Stewart et al., 1980; Nolan et al., 1985; Löf et al., 2000).

The plausible metabolic pathways of methyl chloride in mammals are shown in Figure 2.

In several studies in volunteers, large inter individual differences in concentrations of methyl chloride in breath and blood and amounts of excreted urinary metabolites have been observed (Stewart et al., 1980; van Doorn et al., 1980; Putz-Anderson et al., 1981a; Nolan et al., 1985; Löf et al., 2000).

One explanation for the large interindividual differences in uptake and elimination of methyl chloride in humans is the presence or absence of the enzyme GSTT1 (Coles & Ketterer, 1990). The presence of the GSTT1 gene leads to conjugation between glutathione and methyl chloride (GSTT1+), and the absence of the gene leads to no conjugation (GSTT1-) (Pemble et al., 1994).

About 60% of blood samples from a German population showed a significant metabolic elimination of methyl chloride, whereas 40% did not (Peter et al., 1989). In a Swedish population, Warholm et al. (1994) found a large interindividual variation in the glutathione transferase activity in erythrocytes treated with methyl chloride, as 43% had a high activity, 46% had a medium activity, and 11% lacked activity. Nelson and co-workers (1995) mapped the ethnic differences in the prevalence of the null genotype (GSTT1-) and found the highest prevalence among Chinese (64%), followed by Koreans (60%), African-Americans (22%), and Caucasians (20%), and the lowest among Mexican-Americans (10%). Warholm et al. (1994) concluded that the GSTT1 polymorphism leads to three different phenotypes in humans — namely, non-conjugators (NC), low conjuga tors (LC), and high conjugators (HC). In a comparison involving the three human phenotypes and experimental animals, Thier et al. (1998) established that GSTT1 activity towards methyl chloride in human erythrocytes (HC, LC, or NC) and in liver and kidney cytosol in experimental animals decreased in the following order: female mouse (B6C3F1) > male mouse (B6C3F1) > HC > rat (Fischer 344) > LC > hamster (Syrian golden) > NC. In animals, GSTT1 activity towards methyl chloride was 2–7 times higher in liver cytosol than in kidney cytosol (Thier et al., 1998).

The human GSTT1 polymorphism was illustrated in a study on the toxicokinetics of methyl chloride in volunteers phenotyped for GSTT1 activity (HC, LC, and NC) (Löf et al., 2000). It was seen that conjugators with the fast GSTT1 activity (HC) had the highest respiratory net uptake (respiratory net uptake equals the difference between the amount of methyl chloride in inhaled and exhaled air during exposure) of methyl chloride, whereas subjects with no GSTT1 activity (NC) had a smaller respiratory net uptake. At the end of the exposure, the concentration of methyl chloride in blood declined more rapidly among volunteers with high (HC) and intermedi ate (LC) GSTT1 activity than in those with no activity (NC). The area under the curve for NC was higher than those for HC and LC, and the area under the curve for LC was higher than that for HC. Further, the clearance of methyl chloride by metabolism was high in fast conjuga tors (HC) and close to zero in subjects with no GSTT1 activity (NC).

In an investigation by Dekant et al. (1995), sex-, strain-, and species-specific bioactivation of methyl chloride by cytochrome P-450 2E1 (CYP2E1) was seen in the liver and kidneys of rats and mice. In kidney microsomes, the rate of oxidation of methyl chloride was significantly higher in male mice than in female mice and in rats of both sexes than in mice. It was also observed that the rate of oxidation in kidney microsomes was faster in CD-1 mice and NMRI mice than in C3H/He and C57BL/6J mice. In erythrocytes from other species — rats, mice, cows, pigs, sheep, and rhesus monkeys — no conversion of methyl chloride was seen in erythrocyte cytoplasm (Peter et al., 1989).

In B6C3F1 mice, the LC₅₀ of methyl chloride via inhalation for 6 h was reported to be 4644 mg/m³ (2250 ppm) in males and 17 544 mg/m³ (8500 ppm) in females (White et al., 1982). The data on lethal doses were obtained from an abstract without any further details. In another experimental series, where no clinical acute toxicity symptoms except for lethality were reported, five male B6C3F1 mice were exposed (whole body) to methyl chloride at concentrations of 1032– 5160 mg/m³ (500–2500 ppm) in increments of 1032 mg/m³ (500 ppm) (Chellman et al., 1986a). The 6-h LC₅₀ value was determined to be 4540 mg/m³ (2200 ppm). In this study, lethality as well as hepato toxicity, renal toxicity, and cerebellar degeneration were prevented in mice exposed to 5160 mg/m³ (2500 ppm) for 6 h by pretreatment with the glutathione synthesis inhibitor L-buthionine-S,R-sulfoximine, indicating that metabolism of methyl chloride by glutathione conjuga tion increases the toxicity.

Although other single-exposure inhalation toxicity studies in small rodents exist, they are very old (published before 1950) and do not meet current standards, and they are therefore not included in the present evaluation on methyl chloride. In any case, the results are similar to those reported here.

No single-exposure studies on methyl chloride toxicity following other routes of administration were located in the literature.

In conclusion, based on scarce data, the acute inhalation toxicity in male mice seems to be fairly low, with an LC₅₀ value above 4128 mg/m³ (2000 ppm). In mice, a sex difference in susceptibility to methyl chloride was indicated.

No data on irritation or sensitization were available.

The toxic response to methyl chloride was studied in Fischer 344 rats (10 animals per sex per group) exposed by inhalation to 0, 4128, 7224, or 10 320 mg methyl chloride/m³ (0, 2000, 3500, or 5000 ppm) for 6 h/day for 9 days (5 days of exposure followed by a 2- day break in exposure, then a further 4 days’ exposure) and in C3H, C57BL/6, or B6C3F1 mice (5 animals per strain per sex per group) exposed by inhalation to 0, 1032, 2064, or 4128 mg methyl chloride/m³ (0, 500, 1000, or 2000 ppm) for 6 h/day for 12 days (Morgan et al., 1982). The animals were sacrificed 18 h after their last exposure or immediately after the day’s exposure if they were found to be moribund. Clinical observations and histopathological investigations of the brain, liver, kidneys, and adrenal glands in both species and of testes and epididymis of rats were reported. As a result of high intoxication, some rats from the two highest dose groups were sacrificed in extremis (6 males, 5 females: 10 320 mg/m³ [5000 ppm]; 2 females: 7224 mg/m³ [3500 ppm]). No information was given on whether effects were seen in animals with a fulfilled exposure scheme or with an interrupted scheme.

Clinically, especially in the higher dose groups, the rats were seriously affected by the exposure, and symptoms such as lack of coordination of the forelimbs, paralysis of the hindlimbs, convulsive seizures, perineal urine staining, and diarrhoea were recorded. In the kidneys, concentration-related degeneration and necrosis of the proximal convoluted tubules could be seen (lowest- observed-adverse-effect level or LOAEL [males] = 4128 mg/m³ [2000 ppm]; LOAEL [females] = 7224 mg/m³ [3500 ppm]). Testicular degeneration in the seminiferous tubules (LOAEL = 4128 mg/m³ [2000 ppm]) and adrenal fatty degeneration (LOAEL [males and females] = 7224 mg/m³ [3500 ppm]) were also concentration related. Most animals showed minimal hepatocellular response, including loss of normal areas of cytoplasmic basophilia and variable degeneration. Rats in the 10 320 mg/m³ (5000 ppm) group showed degeneration of the cerebellar granular layer.

All mice in the highest dose group died before or were moribund at exposure day 5. No apparent strain differences could be seen from available mortality data. Prior to death, some of the animals developed moderate to severe ataxia, and all females developed haematurea. In the 2064 mg/m³ (1000 ppm) group, females developed haematurea to a much larger extent than males. Cerebellar degeneration of the same type as in rats was seen at the 2064 and 4128 mg/m³ (1000 and 2000 ppm) concentration levels in female C57BL/6 mice only. The other two mouse strains did not develop brain lesions. On the contrary, in all three mouse strains, degeneration in the kidneys was found at 4128 mg/m³ (2000 ppm), and basophilic renal tubules were observed at 2064 mg/m³ (1000 ppm). Hepatocellular necrosis was confined to the 4128 mg/m³ (2000 ppm) group in male C57BL/6 and B6C3F1 mice. Hepatocellular degeneration was seen in lower dose groups, mainly in the 1032 and 2064 mg/m³ (500 and 1000 ppm) groups of male and female C57BL/6 mice. Liver damage in the low dose groups was considered mild and consisted of, for example, variable degrees of glycogen depletion and cytoplasmic vacuolization.

From these studies, which highlight species and sex differences in methyl chloride-induced toxicity, a rat LOAEL of 4128 mg/m³ (2000 ppm) could be derived from the testicular, epididymal, renal, and to some extent hepatocellular findings, and a mouse LOAEL of 1032 mg/m³ (500 ppm) could be derived from the hepatocellular effects. A no-effect level could not be obtained for either species.

The ultrastructure of the methyl chloride-induced cerebellar lesions observed in mice and rats by Morgan and co-workers (1982) and in guinea-pigs (reported under section 8.4) by von Kolkmann & Volk (1975) was further studied by Jiang et al. (1985) in female C57BL/6 mice. The mice were exposed for 6 h/day, 5 days/week, for 2 weeks to 0 or 3096 mg methyl chloride/m³ (0 or 1500 ppm). In all treated mice, degenerative changes of varying severity were observed in the granular cell layer of the cerebellum. The lesions in the granular cells were characterized by nuclear and cytoplasmic condensation of scattered granule cells and also by watery swelling and disruption of granule cell perikarya. From poorly reported clinical observations, neurological deficiency in motor coordination was seen. Few kidney abnormalities were detected, indicating that the cerebellar degenera tions were not secondary to kidney lesions.

In a study primarily designed to investigate the correlation between neurotoxicity and continuous versus intermittent exposure to methyl chloride, Landry et al. (1985) exposed female C57BL/6 mice for 11 days either continuously (22.5 h/day) to 31, 103, 206, 310, or 413 mg/m³ (15, 50, 100, 150, or 200 ppm) or intermittently (5.5 h/day) to 310, 826, 1651, 3302, or 4954 mg/m³ (150, 400, 800, 1600, or 2400 ppm). A quantitative relationship between neurotoxicity and continuous and intermittent exposure was not observed. The lowest effect level for clinical observations, similar to those reported earlier by Dunn & Smith (1947) and later by von Kolkmann & Volk (1975), Morgan et al. (1982), and Jiang et al. (1985), was 206 mg/m³ (100 ppm) for continuous exposure and 3302 mg/m³ (1600 ppm) for intermittent exposure. Cerebellar lesions were recorded at 206 and 826 mg/m³ (100 and 400 ppm) for continuous and intermittent exposure, respectively. A statistically significant decrease was observed in relative and absolute thymus weights at the 31, 103, and 310 mg/m³ (15, 50, and 150 ppm) exposure levels (con tinuous exposure) and also at the 3302 and 4954 mg/m³ (1600 and 2400 ppm) levels (intermittent exposure). Although the decrease in thymus weight at 31 mg/m³ (15 ppm) suggests that this level might be a LOAEL, the absence of a methyl chloride-induced effect on the thy mus or its function in long-term studies indicates that these weight decreases are of uncertain significance. From the results, a LOAEL of 826 mg/m³ (400 ppm) for intermittent exposure (cerebellar lesions) and of 206 mg/m³ (100 ppm) for continuous exposure can be concluded.

To assess the role of inflammation in the toxicity of methyl chloride, especially to sperm, Chellman et al. (1986b) exposed male Fischer 344 rats for 5 days, 6 h/day, to 0 or 10 320 mg methyl chloride/m³ (0 or 5000 ppm) with and without the presence of the anti- inflammatory agent 3-amino-1-(m-[trifluoromethyl] phenyl)-2-pyrazoline (BW755C), an inhibitor of leuco triene and prostaglandin synthesis. Lesions that were induced by exposure to methyl chloride alone were epididymal sperm granulomas, degeneration of cere bellar granule cells, necrosis of renal proximal tubules, cloudy swelling of hepatocytes, and vacuolization of cell cytoplasm in the outer region of zona fasciculata in the adrenal glands. Virtually none of these effects was seen when BW755C was given in parallel with methyl chlor ide, strongly suggesting an inflammatory response.

The question as to whether methyl chloride- induced renal tumours in male mice are evoked by the metabolic intermediate formaldehyde was studied by Jäger et al. (1988). Fischer 344 rats and B6C3F1 mice of both sexes in groups of five were exposed to 0 or 2064 mg methyl chloride/m³ (0 or 1000 ppm) for 6 days, and DNA lesions (cross-links and single-strand breaks), glutathione transferase (GST) activity, and formaldehyde dehydrogenase (FDH) activity were measured. It was shown that the tumour formation in male mice is not based on any obvious biochemical sex differences in enzymatic transformation with respect to FDH. Neither is the metabolically formed formaldehyde likely to be the effective carcinogen, as the characteristic formaldehyde- induced genetic damage is absent. However, the signifi cant species difference between mice and rats — in that mice, due to higher GST activity, especially in the kid neys, seem to be more susceptible to methyl chloride treatment — could not be ruled out. It was, for example, not shown if toxicity caused by the glutathione conjuga tion pathway was due to a metabolite formed or to gluta thione depletion, as suggested by Jäger et al. (1988).

In conclusion, after short-term exposure, the target organ in both rats and mice is the nervous system, with functional disturbances and cerebellar degeneration. The LOAELs in mice are 206 and 826 mg/m³ (100 and 400 ppm) upon continuous and intermittent exposure, respectively. Higher levels of exposure caused toxicity in the kidney and liver in mice and in the testes, epididy mis, and kidney in rats. A mouse LOAEL of 1032 mg/m³ (500 ppm) could be derived from liver toxicity data. The decrease in thymus weight, unaccompanied by histopath ological changes, in mice exposed to 31 mg/m³ (15 ppm) was not corroborated by either a 90-day (CIIT, 1979) or a 2-year study (CIIT, 1981) (reported in sections 8.4 and 8.5, respectively). Because of this lack of corroboration, the decrease in thymus weight will not be forwarded to the sample risk characterization.

No toxicity data for short-term exposures other than those obtained from administration via the respira tory pathway were located in the literature.

To investigate methyl chloride-induced neurotox icity, von Kolkmann & Volk (1975) exposed 19 guinea- pigs to 41 280 mg methyl chloride/m³ (20 000 ppm; 2 vol% in a pressurized vessel) by inhalation for 61– 70 days (10 min/day, 6 times/week). Clinically, in approximately half of the animals in the treated group, ataxia, paresis of the hind legs, staggering, atactic moving of the head, and retardation in spontaneous reaction and mobility were observed. No animal died during the exposure period. Histopathologically, necroses were seen in the cerebellar cortex in the granular cell layer. Further, Purkinje’s cell necrosis occurred. Considering the extremely high exposure concentrations, the study can be used for descriptive purposes only.

In a subchronic toxicity study, 80 Fischer 344 rats (40 per sex) and 80 B6C3F1 mice (40 per sex) were exposed to methyl chloride by inhalation at concentra tions of 0, 774, 1548, or 3096 mg/m³ (0, 375, 750, or 1500 ppm) for 90 days (CIIT, 1979). Clinical observa tions and data on food consumption, body and organ weights, haematology, clinical chemistry, urinalysis, ophthalmoscopic examination, gross pathology, and histopathology were recorded.

Female mice in the 3096 mg/m³ (1500 ppm) dose group had significantly depressed total body weight at the end of the exposure period. Absolute and/or relative organ weights were increased for heart, brain, spleen, liver, kidneys, and lungs in female mice (mainly in the highest dose group) and in pancreas in male mice. Cyto plasmic vacuolization of hepatocytes occurred in the two highest dose groups and was considered compound related. In the 1548 mg/m³ (750 ppm) dose group, vacuolization was seen 5 times as frequently in females as in males. Exposure-related fluctuations in haema tology and in clinical chemistry were observed but not considered significant, as they were within the control range. Further, the methyl chloride-exposed mice had a high incidence of a mucopurulent conjunctivitis. How ever, this effect was probably not related to methyl chlor ide exposure, as it was seen mainly in the 774 mg/m³ (375 ppm) dose group. The available results indicate that female mice are more sensitive than male mice to methyl chloride exposure.

Male rats in all dose groups and females in the two highest dose groups showed significantly decreased absolute body weight. In rats (males and/or females; mainly in the highest dose group), absolute and/or rela tive organ weights were increased for heart, brain, testes, ovaries, spleen, liver, kidney, pancreas, and adrenals.

In the 1979 CIIT study (which served as a pilot for the 2-year chronic toxicity/carcinogenicity study by CIIT [1981]), no compound-related lesions were reported from gross pathology or histopathology on kidneys, heart, or testes. The absence of recorded organ lesions in mice could be due to a fairly high mortality, especially in the highest dose group, in combination with the histo logical examination applied. In the histopathological examination, as a first step, the highest dose group was compared with the control group. In the case of positive findings, the control animals were thereafter compared with the 1548 mg/m³ (750 ppm) dose group and then the 375 ppm (774 mg/m³) dose group. This procedure might suffer from a high mortality in the 3096 mg/m³ (1500 ppm) dose group and give rise to false negatives. No similar explanation could be given for rats, as the mortality during the exposure period was low.

In a 2-year inhalation study, Fischer 344 rats and B6C3F1 mice (120 animals per sex per group) were exposed to 0, 103, 464, or 2064 mg methyl chloride/m³ (0, 50, 225, or 1000 ppm) for 6 h/day, 5 days/week, with the objective of determining the potential toxicological and oncogenic effects (CIIT, 1981). Planned interim necropsies of the experimental animals were completed at 6, 12, 18, and 24 months following initiation of exposure. As a result of high mortality in the mouse high-dose group, the scheduled 24-month sacrifice was carried out after 21 or 22 months of exposure. After 6 or 12 months, 10 rats per sex per dose group were sched uled to be sacrificed, and after 18 or 24 months, 20 and 80 rats per sex per dose group, respectively. Mice were scheduled for sacrifice in groups of 10 per sex per dose group after 6, 12, or 18 months and in groups of 90 per sex per dose group after 24 (or 21, 22 months). Data on body weights, clinical signs of toxic effects, clinical chemical analyses, gross pathology, and histopathology were recorded.

During the exposure period, rat survival was not affected by methyl chloride exposure. However, mouse survival was low in the 2064 mg/m³ (1000 ppm) dose group compared with the control animals. The high mortality occurred predominantly during the first 6 months and was attributed by CIIT (1981) to fighting for dominance. The number of rats and mice that died during the 2-year study for reasons other than planned sacrifice is given in Table 2.

Total body weight gain was significantly reduced throughout the exposure period for male and female rats in the 2064 mg/m³ (1000 ppm) exposure group. Although female rats in the 464 mg/m³ (225 ppm) group and female mice in the 2064 mg/m³ (1000 ppm) group also had significantly decreased growth rates, these occurred periodically and were not observed at the end of exposure. The relative heart weight was increased in female mice and male and female rats at 2064 mg/m³ (1000 ppm). Otherwise, changes in relative or absolute organ weights were seen at the 2064 mg/m³ (1000 ppm) exposure level for kidney, liver, heart, and brain in both species and in testes in rats. For comparison with the short-term exposure study by Landry et al. (1985), thymus weight was not affected by methyl chloride exposure in the 2-year study.

Clinical observations on toxicity to the central nervous system (hunched posture, tremor, and paralysis) were seen in mice in the highest dose group but not in rats.

In mice, statistically significant hepatocellular changes (vacuolization, karyomegaly, cytomegaly, and degeneration) were seen in male and female mice from the 2064 mg/m³ (1000 ppm) group. In male mice exposed to 2064 mg methyl chloride/m³ (1000 ppm), significantly elevated serum glutamic–pyruvic transami nase (SGPT) values were seen, coupled with histopatho logical findings in the liver. Elevated SGPT values were also observed in the lower dose groups but were not correlated with any histopathological findings.

In the 2064 mg/m³ (1000 ppm) dose group in male mice, a large, statistically significant increase (P > 0.05) in the development of renal tubuloepithelial hyperplasia, hypertrophy, and/or karyomegaly, with onset at 12 months, was observed. Further, in the same group, significant exposure-related increases (P < 0.05) in numbers of observed renal cortical adenomas as well as renal adenocarcinomas (including those designed as renal cortical adenocarcinomas and renal cortical papillary cystadenocarcinomas) were noted in animals sacrificed or dying between 12 and 21 months (inci dences of cortical renal lesions are found in Table 3). Cortical adenomas were also seen in two male mice in the 464 mg/m³ (225 ppm) group. Although this increase was not of statistical significance, the adenomas were similar to those that occurred in the 2064 mg/m³ (1000 ppm) group and were therefore judged to be associated with the methyl chloride exposure. In the 103 mg/m³ (50 ppm) group, there was a slightly increased incidence of renal cortical microcysts in male mice sacrificed at 24 months compared with the control males (6/32 vs. 1/20). Renal microcysts were also observed in the 464 mg/m³ (225 ppm) group; the increase, as compared with the control group, was not, however, significant in either males or females. The incidence of renal cortical microcysts was not reported in animals from the highest dose group. Since the micro cysts appear to be variations of the same lesion observed at higher exposure levels, they should be considered to be related to methyl chloride, although no concentration dependence could be established.

Further, at 2064 mg/m³ (1000 ppm), degeneration and atrophy of the seminiferous tubules were seen, as well as lymphoid depletion and splenic atrophy.

At the 18-month sacrifice, cerebellar lesions (degeneration and atrophy of the cerebellar granular layer) were noted in male and female mice at the 2064 mg/m³ (1000 ppm) level. Mice from the control, low, and intermediate exposure groups did not have lesions in the granular cell layer of the cerebellum. At the 22-month sacrifice, similar but more extensive observations in 17/18 females were reported from the 2064 mg/m³ (1000 ppm) group (the only group examined at this time period).

At the 18-month sacrifice, axonal swelling and degeneration of minor severity were observed in the spinal nerves and cauda equina associated with the lumbar spinal cord. The effects occurred in most treated animals in all dose groups (controls: 1/5 males and 2/10 females; 103 mg/m³ [50 ppm]: 4/5 males and 10/10 females; 464 mg/m³ [225 ppm]: 5/5 males and 5/5 females; 2064 mg/m³ [1000 ppm]: 3/7 males and no data on females). The effects at each exposure level were significantly increased in each dose group as compared with control animals. However, no concentration– response relationship could be established. In the 2064 mg/m³ (1000 ppm) dose group sacrificed at 22 months, minimal to moderate swelling and degener ation of the lumbar spinal nerves were recorded in 13 of 18 female mice. Twelve of 18 females had similar lesions in the thoracic spinal cord, and 6/18 females in the cervical spinal cord. Histopathological examinations of mice in the 2064 mg/m³ (1000 ppm) dose group with unscheduled death showed high incidences of cerebellar lesions and axonal degeneration of lumbar spinal nerves, lesions similar to those found at the 18-month sacrifice.

In rats, exposure to methyl chloride at 2064 mg/m³ (1000 ppm) caused testicular lesions (bilateral, P > 0.05, and unilateral, P > 0.05, diffuse degenerations and atrophies of the seminiferous tubules). These lesions were statistically significant as compared with the control group and were first observed at the 6-month sacrifice. Sperm granulomas were noted in three male rats at 2064 mg/m³ (1000 ppm). No statistically signifi cant changes other than the effects on body weight gain were seen in female rats. This indicates that the maxi mum tolerated dose used in the CIIT (1981) study might have been too low to induce toxicity in female rats.

Significant findings from the long-term studies in mice are disturbances of the nervous system and induction of tumours and microcysts in male mice. Axonal swelling and degeneration of spinal nerves in all exposure groups suggest a LOAEL of 103 mg/m³ (50 ppm). This LOAEL is forwarded to the sample risk characterization. The observation of renal microcysts in the 103 mg/m³ (50 ppm) dose group (although not con centration related) supports the LOAEL of 103 mg/m³ (50 ppm). Further significant findings are the testicular lesions in rats. In male rats, testicular lesions occurred at 2064 mg/m³ (1000 ppm). No toxic effects in females were reported.

No toxicity data from long-term exposure to methyl chloride other than those obtained from adminis tration via the respiratory pathway were located in the literature.

For an overview of, and details on, the geno toxicity studies, see Table 4. In all studies referred to below, methyl chloride was administered by inhalation, unless otherwise noted.

8.6.1 Studies in vitro

Using the Ames assay, methyl chloride was shown to induce gene mutations in Salmonella typhimurium TA100 (Simmon et al., 1977) and in S. typhimurium TA1535 (Andrews et al., 1976) in both the presence and absence of metabolic activation. Further, a concentra tion-related increase in the 8-azaguanine-resistant fraction in S. typhimurium was observed (Fostel et al., 1985).

Methyl chloride induced the adaptive response to alkylation damage in Escherichia coli regulated by the ada protein, which suggests that methyl chloride is a direct DNA-alkylating agent (Vaughan et al., 1993).

Methyl chloride caused gene mutations in vitro in TK6 human lymphoblasts, as shown by a dose-related increase in the mutant fraction (Fostel et al., 1985).

In the study by Fostel and co-workers (1985), no increase in DNA strand breaks as measured by alkaline elution was observed. However, the outcome from the positive control (methyl methane sulfonate [MMS]) was questionable, as unexpectedly high doses were needed for a positive result. Thus, the mutagenic lesions pro duced by methyl chloride might be either different from those produced by MMS or formed at a level below the threshold of detection of alkaline solution.

DNA damage following methyl chloride exposure was shown as a statistically significant enhanced transformation of Syrian hamster embryo cells by SA7 adenovirus (Hatch et al., 1983).

In vitro, 1–10% methyl chloride caused induction of unscheduled DNA synthesis (UDS) in rat spermato cytes and hepatocytes (Working et al., 1986).

Methyl chloride was shown to directly bind to bovine serum albumin. No further details were available (Kornbrust et al., 1982).

In TK/6 human lymphoblasts, methyl chloride induced a statistically significant concentration-related induction of sister chromatid exchange (SCE) frequency as well as significantly declined mitotic index and a significant concentration-related increase in second- division metaphases (Fostel et al., 1985).

8.6.2 Studies in vivo

In vivo, methyl chloride did not cause induction of UDS in rat spermatocytes, hepatocytes, or tracheal epithelial cells at exposure concentrations of 6192– 7224 mg/m³ (3000–3500 ppm) for 6 h/day for 5 days. However, exposure to 30 960 mg/m³ (15 000 ppm) for 3 h did cause a marginal increase in UDS in hepatocytes (Working et al., 1986). The doses used were considered below, but close to, the maximum tolerated dose.

In a macromolecular binding study, male rats were exposed to ¹⁴C-labelled methyl chloride (specific activity 25–70 dpm/nmol = 11.2–31.4 × 10^–3 mCi/mmol), and accumulation of ¹⁴C was measured in lipid, RNA, DNA, and protein from isolated liver, kidneys, lungs, and testes (Kornbrust et al., 1982). Radiolabelled carbon was found in all tissues and fractions studied; however, methylation was not found. Pretreatment with the protein synthesis inhibitor cycloheximide or the folic acid antagonist methotrexate, which interferes with single-carbon metabolism, to a large extent inhibited most of the ¹⁴C- incorporation in proteins and macromolecules, respectively. Further, the extent of incorporation of methyl chloride into proteins and lipids was consistent with the rates of turnovers in these macromolecules. Therefore, the most likely mechanism for uptake of methyl chloride into macromolecules is via the one-carbon pool. How ever, this does not exclude the possibility that methyl chloride might bind directly to macromolecules to a lesser extent.

In a DNA binding assay, in which Peter et al. (1985) exposed rats and mice to ¹⁴C-labelled methyl chloride, no methylation of guanine in DNA at N⁷ and/or O⁶ in liver and kidneys by methyl chloride was found. The specific activity in this study (13 mCi/mmol = 2.9 × 10⁴ dpm/nmol) was about 3 orders of magnitude higher than that used by Kornbrust et al. (1982).

Using the alkaline elution technique, no cross-links in male mouse kidneys could be detected after exposure to 2064 mg methyl chloride/m³ (1000 ppm) for 6 days, but some indications of DNA single-strand breaks were obtained (Jäger et al., 1988). However, when mice were exposed to 2064 mg/m³ (1000 ppm) for 8 h only, DNA cross-links were seen in renal tissue of male mice but not in female mice or in hepatic tissues (Ristau et al., 1989). In an attempt to investigate the time-course of the DNA lesions, Ristau et al. (1990) again exposed male mice to 2064 mg methyl chloride/m³ (1000 ppm) for 8 h. In renal tissue, it was observed that DNA–protein cross-links were removed at a fast rate, whereas single-strand breaks appeared to accumulate. At 48 h postexposure, all lesions had disappeared.

In a dominant lethal assay, performed according to Organisation for Economic Co-operation and Develop ment (OECD) test guidelines, male rats were exposed to methyl chloride (Working et al., 1985a). The numbers of live and total implants were decreased, there was an increase in the percentage of preimplantation loss at weeks 2, 4, 6, and 8 postexposure, and there was an increase in the percentage of postimplantation loss at week 1 postexposure. The changes observed were not concentration related. A true dominant lethal effect of genetic origin could be questioned, as the time-courses of the pre- and postimplantation losses after methyl chloride exposure were not the same as those obtained after administration of the positive control, triethylene melamine (TEM). The development of sperm granulo mas in the epididymis, the effects seen in the dominant lethal assay, seems to be cytotoxic rather than genotoxic in origin. However, a genotoxic effect should not be totally excluded.

The role of epididymal inflammation in the induc tion of lethal mutations was studied by Chellman et al. (1986c) in an assay with a test protocol similar to the OECD test guideline for dominant lethal mutations. Rats were exposed to methyl chloride in the presence or absence of the anti-inflammatory agent BW755C. BW755C was effective against the postimplantation losses induced by methyl chloride, but not against pre- implantation losses. The authors’ conclusions, based on unpublished data mentioned in Chellman et al. (1986c), are that the increase in preimplantation losses might be a consequence of testicular lesions caused by methyl chlor ide and that BW755C is effective against epididymal injuries only, thus indicating that epididymal inflamma tion has a role in the induced infertility.

In conclusion, methyl chloride is clearly genotoxic in in vitro systems, in both bacteria and mammalian cells. Methyl chloride binds to protein. Methyl chloride is possibly an alkylating agent; however, the available studies do not allow any quantification. Although the positive effects seen in a dominant lethal test were most likely cytotoxic rather than genotoxic, methyl chloride might be considered a very weak mutagen in vivo based on some evidence of DNA–protein cross-linking at higher doses.

8.7.1 Effects on fertility

In the 1981 CIIT study, referred to in section 8.5, exposure to methyl chloride at 2064 mg/m³ (1000 ppm) caused testicular lesions in rats. The lesions seen were bilateral and consisted of diffuse degeneration and atrophy of the seminiferous tubules.

Chapin et al. (1984) investigated the development of lesions induced in testes and epididymis and effects on reproductive hormones in F-344 rats after exposure to 0 or 6192 mg methyl chloride/m³ (0 or 3000 ppm) for a total of 9 days (6 h/day; approximately 60 exposed ani mals and 16 control animals). Testicular lesions in the form of delay in spermiation, germinal epithelial vacuo lization, and cellular exfoliation as well as bilateral epi didymal granulomas were seen. The effects were observed in most animals, with the onset at day 9 or 11 after the beginning of the exposure. In general, lesions seen in animals at day 19 were of higher severity than those seen earlier. In rats killed 70 days or more after the onset of the exposure, 70–90% of the seminiferous tubules lacked any germinal cells; in 10–30% of the tubules, varying degrees of recovery of spermiation were observed. The LOAEL in this study must be set at 6192 mg/m³ (3000 ppm).

A two-generation inhalation study in Fischer 344 rats was carried out at methyl chloride concentrations of 0, 310, 980, or 3096 mg/m³ (0, 150, 475, or 1500 ppm) (Hamm et al., 1985). The F0 generation (40 males and 80 females per exposure group) was exposed for 10 weeks and during a 2-week mating period (6 h/day, 5 days/week, and 6 h/day, 7 days/week, respectively). A similar exposure schedule was used for the F1 genera tion, with the exclusion of the 3096 mg/m³ (1500 ppm) exposure level. In the high-dose F0-generation males sacrificed immediately after 12 weeks of exposure, treatment-related lesions were found, consisting of minimal to severe atrophy of the seminiferous tubules (10/10 males examined) and granulomas in the epidid ymis (3/10). Severely affected tubules were lined by Sertoli’s cells and by occasional stem cell spermatogo nia. In the less affected tubules, decreased numbers of spermatogonia, primary spermatocytes, and/or secondary spermatocytes were found.

Further, in the F0 generation, no litters were born when high-dose males were mated to exposed or unex posed females, and significantly fewer litters were born to unexposed females mated to the males in the 980 mg/m³ (475 ppm) dose group. No differences in litter size, sex ratio, pup viability, or pup growth were found among the 980 mg/m³ (475 ppm) and 310 mg/m³ (150 ppm) groups compared with the control F0 group. A trend towards decreased fertility was also found in the 980 mg/m³ (475 ppm) dose group in the F1 generation. A LOAEL of 980 mg/m³ (475 ppm) (infertility) was derived from the two-generation study. Breeding results in rats in the F0 and F2 generations after methyl chloride exposure are shown in Table 5.

In a dominant lethal assay in rats exposed to methyl chloride for 5 days, described in section 8.6, visible sperm granulomas in the epididymis were present in the 6192 mg/m³ (3000 ppm) group 17 weeks post exposure but not in the 2064 mg/m³ (1000 ppm) group or in the control group. After exposure to 6192 mg/m³ (3000 ppm), the number of live and total implants was decreased, and there was an increase in postimplantation loss. In both treated groups, there was an increase in preimplantation losses (Working et al., 1985a). The LOAEL for preimplantation loss was 2064 mg/m³ (1000 ppm).

In a subsequent study, Working et al. (1985b) characterized the effect of methyl chloride exposure on sperm quality and histopathology in rats in more detail. Male Fischer 344 rats (80 animals per group) were exposed to 0, 2064, or 6192 mg methyl chloride/m³ (0, 1000, or 3000 ppm) for 5 days, 6 h/day. Besides significantly decreased testis weights in the high-dose group 3–8 weeks postexposure and the findings that more than 50% of the treated animals showed sperm granulomas in the epididymis in the same dose group, observations indicating cytotoxic effects on sperm quality were made. Observations made at the 6192 mg/m³ (3000 ppm) level included significant decreases in testicular spermatid head counts, delay in spermiation, epithelial vacuoliza tion, luminal exfoliation of spermatogenic cells, and multinucleated giant cells. Further, sperm isolated from the vasa deferentia had significantly depressed numbers and an elevated frequency of abnormal sperm head morphology by week 1 postexposure and significantly depressed sperm motility and increased frequency of headless tails by week 3 postexposure. These changes were all within or close to the normal range by week 16 postexposure. A LOAEL of 6192 mg/m³ (3000 ppm) could be derived based on the histopathological findings.

The cause of preimplantation loss induced by methyl chloride was further investigated in rats by Working & Bus (1986). Fischer 344 rats (10–30 animals per group) inhaled 0, 2064, or 6192 mg methyl chlor ide/m³ (0, 1000, or 3000 ppm) 6 h/day for 5 days or received a single injection of TEM as a positive control for genotoxicity. At weeks 1–3 postexposure in the 6192 mg/m³ (3000 ppm) group, preimplantation losses did not exceed unfertilized ova, which was the case for the positive control. From these data, the authors sug gested that preimplantation losses are due to a failure in fertilization rather than to an increase in embryonal deaths.

In conclusion, testicular lesions and epididymal granulomas followed by reduced sperm quality lead to reduced fertility as well as complete infertility in rats. A LOAEL of 980 mg/m³ (475 ppm) and a no-observed- adverse-effect level (NOAEL) of 310 mg/m³ (150 ppm) were identified from the two-generation study of Hamm et al. (1985).

8.7.2 Developmental toxicity

In a study designed to study structural teratogenic ity, pregnant Fischer 344 rats (25 rats per dose group) were exposed to 0, 206, 1032, or 3096 mg methyl chloride/m³ (0, 100, 500, or 1500 ppm) for 6 h/day through gestation days 7–19 (Wolkowski-Tyl et al., 1983a). In the highest dose group, significant reductions in fetal body weight and female crown–rump length were observed. Further, skeletal immaturities such as reduced ossification (metatarsals and phalanges of the anterior limbs, thoracic vertebral centra, pubis of pelvic girdle, and metatarsals of the hindlimbs) were seen. Although these findings were seen in the presence of significantly decreased maternal food consumption, body weight, and weight gain in the same dose group, they should be considered as serious and exposure related. A fetal LOAEL for skeletal immaturities as well as a maternal LOAEL for effects on body weight and food consump tion of 3096 mg/m³ (1500 ppm) were obtained. No other effects, including heart defects, were reported from the 206 and 1032 mg/m³ (100 and 500 ppm) dose groups.

In parallel with the rat study, Wolkowski-Tyl et al. (1983a) also evaluated teratogenicity in pregnant C57BL/6 mice (33 mice per dose group) carrying B6C3F1 fetuses exposed through gestation days 6–17 following the same exposure schedule as the rats. Dams in the 3096 mg/m³ (1500 ppm) group died or were killed in extremis due to very high toxicity (tremor, hunched appearance, difficulty in righting, vaginal bleeding, bloody urine, cerebellar granular cell necrosis and degeneration, etc.). No other maternal toxicity was observed in the other exposure groups. In the 1032 mg/m³ (500 ppm) group, the fetuses (male and female) had a small but significant increase in heart defects (reduction or absence of the atrioventricular valves, chordae tendineae, and papillary muscles). In both the 1032 and 206 mg/m³ (500 and 100 ppm) groups, a significant increase in degree of ossification in the hindlimbs was seen as compared with control animals. A LOAEL for heart defects in fetuses of 1032 mg/m³ (500 ppm) was obtained.

In a subsequent study, Wolkowski-Tyl et al. (1983b) again exposed pregnant C57BL/6 mice carrying B6C3F1 fetuses with the aim of confirming the earlier findings, elucidating the nature of the heart defects more clearly, and establishing a concentration–effect relation ship. Approximately 75 mice per dose group were exposed to 0, 516, 1032, or 1548 mg methyl chloride/m³ (0, 250, 500, or 750 ppm) for 6 h/day during gestation days 6–18. In the 1032 and 1548 mg/m³ (500 and 750 ppm) groups, an exposure-related increase in heart defects (involving effects on the atrioventricular valves, chordae tendineae, and papillary muscles) was observed. Dams were affected at the 1548 mg/m³ (750 ppm) exposure level (decrease in body weight and body weight gain). No maternal toxicity, embryotoxicity, fetotoxicity, or teratogenicity was associated with exposure to methyl chloride at 516 mg/m³ (250 ppm). In this study, the LOAEL for heart defects was 1032 mg/m³ (500 ppm), the NOAEL was 516 mg/m³ (250 ppm), and the maternal LOAEL was 1548 mg/m³ (750 ppm).

In a number of different experiments on small numbers of animals, pregnant C57BL/6 mice carrying B6C3F1 fetuses were exposed to methyl chloride at concentrations of 516, 619, or 2064 mg/m³ (250, 300, or 1000 ppm) for 12–24 h during gestation day 11.5–12.5 (John-Greene et al., 1985). The exposure time was chosen as a critical period in development of cardiac defects. The authors found heart defects when the test was non-blind but not when the technician was unaware of which fetuses were exposed. Further, John-Greene et al. (1985) had concerns regarding the technique used by Wolkowski-Tyl et al. (1983a). However, the investiga tion by John-Greene et al. (1985) is difficult to evaluate, as a small number of animals were used and as the expo sure period was not similar to those in the Wolkowski- Tyl et al. (1983a, 1983b) studies. No LOAEL could be established.

In conclusion, from the studies by Wolkowski-Tyl and co-workers (1983a, 1983b), it seems that methyl chloride could induce heart defects in mice exposed to 1032 mg/m³ (500 ppm) when dams were exposed through gestation days 6–18. A NOAEL of 206 mg/m³ (100 ppm) was established from the developmental toxicity studies.

No specific reports on immunological or neuro logical effects caused by methyl chloride were found in the literature.

In order to monitor the physiological response to methyl chloride in healthy volunteers (eight men and nine women) with no previous methyl chloride exposure, Stewart et al. (1980) exposed the volunteers to methyl chloride at concentrations of 0, 41, 206, or (men only) 310 mg/m³ (0, 20, 100, or [men only] 150 ppm), 1, 3, and 7.5 h/day, 5 days/week, for 6 weeks in an exposure chamber. Using a wide battery of behavioural, neuro logical, electromyographic, and clinical tests, no significant decrements were found in the exposed volunteers as compared with controls. For interindividual differences in methyl chloride concentrations in blood and expired air, see section 7.

In volunteers, Putz-Anderson and co-workers (1981a, 1981b) found minimal or no effects on perfor mance after exposure to methyl chloride at 206 or 413 mg/m³ (100 or 200 ppm) for 3 h (n = 56) and at 413 mg/m³ (200 ppm) for 3.5 h (n = 84), respectively.

Available information related to the toxic effects on humans exposed to high concentrations of methyl chloride is mainly derived from accidental exposures in connection with the use of methyl chloride in the produc tion of polystyrene foams and also from refrigerator leakages. Among symptoms described in case reports (see, for example, MacDonald, 1946; McNally, 1946; Hansen et al., 1953; Thordarson et al., 1964; Scharn weber et al., 1974; Spevak et al., 1976; Gudmundsson, 1977; Lanham, 1982) are effects on the nervous system, such as dizziness, weakness, blurred vision, muscular incoordination, drowsiness, sleep disturbances, mental confusion, and paraesthesis. Neurotic and depressive symptoms are also described. Further, gastrointestinal symptoms (nausea, vomiting, abdominal pain, etc.) have been observed, as well as jaundice. In general, the symptoms seem to develop soon after the exposure. However, recovery periods vary to a large extent; for example, in seamen highly exposed to methyl chloride, effects on the nervous system were observed 13 years after the accident (Gudmundsson, 1977).

Performance and cognitive functions were adversely affected in workers manufacturing foam products. There was also an increase in the magnitude of finger tremors. The workers were exposed for 2 years to approximately 72 mg methyl chloride/m³ (35 ppm), as well as other chemicals (NIOSH, 1976). However, insufficient information was available on exposure to the other chemicals and lifestyle factors, and no relationship could be established between methyl chloride exposure and the various psychological and personality tests employed.

A mortality follow-up study was conducted of 852 male workers employed for at least 1 month between the years 1943 and 1978 in a butyl rubber manufacturing plant using methyl chloride (Holmes et al., 1986). For each cohort member, complete work history and death information were obtained. No information on lifestyle factors was reported. The exposure to methyl chloride and other compounds used in the butyl rubber manufac turing plant was estimated in three categories (high, medium, and low). No detectable excess mortality from any specific cause of death including all cancers was found in the study population after analysis by level and duration of exposure.

In a 32-year follow-up study by Rafnsson & Gudmundsson (1997), indications of elevated mortality from cardiovascular disease after high accidental methyl chloride exposure were seen in Icelandic seamen (deck hands: relative risk [RR] = 3.9, 95% confidence interval [CI] = 1.0–14.4; officers: RR = 1.7, 95% CI = 0.3–6.4). The small number of observed cancers (all cancers and lung cancers) in the exposed group provides an insuffi cient basis for assessing the cancer risk in humans. The reference group used was controlled for age, occupation, social class, and lifestyle factors.

In conclusion, effects on humans, especially on the central nervous system, can clearly be seen after acciden tal (mostly high) exposure or after normal work exposure levels. A rough estimation of the degree of exposure from case reports might be in the order of approximately 200–2000 mg/m³ (100–1000 ppm). In short-term expo sure of volunteers, no significant effects were seen. There are insufficient data available to assess the risk for humans to develop cancer as a result of methyl chloride exposure.

Few data were found on the short-term toxicity of methyl chloride to aquatic organisms, and no data were found on long-term toxicity. The existing data for a cyanobacterium, a green alga, a protozoan, and two fish species indicate a low acute toxicity to aquatic species (see Table 6). The acute toxicity to the two fish species was determined under static conditions, and the concen tration of the test substance was not measured. This means that the toxicity may have been underestimated by the test, if significant amounts of the test substance volatilized during the test. The 96-h LC₅₀ values for the freshwater species, bluegill sunfish (Lepomis macro chirus), and the saltwater species, tidewater silverside (Menidia beryllina), were determined to be 550 and 270 mg/litre, respectively (Dawson et al., 1977).