Data on sources and emissions from Canada, the source country of the national assessment on which this CICAD is based, are presented here as an example. Sources and patterns of emissions in other countries are expected to be similar, although quantitative values may vary.

Estimates of emissions of butadiene are highly variable, depending on the method of estimation and the quality of the data upon which they are based. Total Canadian emissions for 1994 were estimated to range between 12 917 and 41 622 tonnes (Environment Canada, 1998). Major uncertainties are associated with estimates for combustion sources, notably forest fires.

Butadiene is released from biomass combustion, especially forest fires. Total global emissions of buta diene from biomass combustion were estimated to be 770 000 tonnes per year (Ward & Hao, 1992). Releases from forest fires in Canada were estimated to range between 3607 and 26 966 tonnes, which constituted 49.3% (range of estimates is 28–65%) of the total annual emissions of butadiene in Canada (CPPI, 1997). Although Altshuller et al. (1971) suggested that buta diene can be released from natural gas losses and diffu sion through soil from petroleum deposits, no data were identified on this possible source.

All internal combustion engines may produce butadiene as a result of incomplete combustion. The amount generated and released depends primarily on the composition of fuel, the type of engine, the emission control used (i.e., presence and efficiency of catalytic converter), the operating temperature, and the age and state of repair of the vehicle.² Cyclohexane, 1-hexene, 1-pentene, and cyclohexene have been identified as primary fuel precursors for butadiene (Schuetzle et al., 1994). As well, very low levels of butadiene itself may be present in gasoline and in liquefied petroleum gas.

Butadiene can also enter the environment from any stage in the production, storage, use, transport, or dispo sal of products with residual, free, or unreacted butadi ene. Data on Canadian industrial emissions have been collected for industrial processes, plastic products indus tries, refined petroleum and coal products industries, and chemical and chemical products industries as part of the National Pollutant Release Inventory (NPRI) (Environ ment Canada, 1996a, 1997). Emissions other than those reported to the NPRI may occur, including from combus tion of other fuels (e.g., natural gas, oil, and wood space heating), prescribed forest burning, cigarettes, waste incineration, releases from polymer products, releases from the use and disposal of products containing buta diene, and spillage (Ligocki et al., 1994; Environment Canada, 1996b; OECD, 1996).

The following amounts of butadiene were estimated to have been released into the Canadian environment in 1994 from key transportation and related sources (Environment Canada, 1996a; CPPI, 1997): 3376–7401 tonnes from on-road gasoline- and diesel- powered motor vehicles (with about 45–89% of those releases from gasoline engines and 11–55% from diesel engines); 150–258 tonnes from aircraft; 84–1689 tonnes from off-road motor vehicles; 84 tonnes from lawn mowers; 40 tonnes from the marine sector; and 17 tonnes from the rail sector.

In addition, data from NPRI for 1994 (Environ ment Canada, 1996a) listed a total of 270.4 tonnes released from the chemical and chemical products industries. Of this, 270.3 tonnes were released into air, 0.058 tonnes into water (St. Clair River, Ontario), and 0.002 tonnes onto land. There were releases of 17.5 tonnes into air from the plastic products industries. A total of 22.3 tonnes was released from the refined petroleum and coal products industries, of which 22.2 tonnes were released into air. Off-site transfer of wastes (material sent for final disposal or treatment prior to final disposal) from industrial sites in Canada in 1994 was estimated to include a total of 131.3 tonnes of buta diene, with 128.7 tonnes being sent to incineration, 2.1 tonnes to landfill, and 0.5 tonnes to municipal sewage treatment plants (Environment Canada, 1996a). Based on 1995 NPRI data (Environment Canada, 1997), the amount of butadiene estimated to have been released into the Canadian environment was 225.8 tonnes from industrial on-site uses, with 0.058 tonnes released into water, 0.002 tonnes into land, and 225.4 tonnes into air. Releases into air included air fugitive releases (172.8 tonnes), air stack releases (36.3 tonnes), air stor age releases (4.8 tonnes), air spill releases (1.1 tonnes), and other air releases (10.4 tonnes).

Based on data in NPRI, it was estimated that the total release of butadiene from fuel distribution in 1994 was 24 tonnes (Environment Canada, 1996a), although gasoline and diesel fuel contain little or no butadiene (US EPA, 1989).

CPPI (1997) estimated that releases into the Canadian environment in 1994 were 1191 tonnes from prescribed forest burning, 3706 tonnes from wood space heating, 11 tonnes from natural gas/oil space heating, and 1–9 tonnes from cigarettes.

Butadiene is produced during the combustion of organic matter in both natural processes and human activities. In addition, it is produced commercially for use in the chemical polymer industry.

Butadiene is purified by extraction from a crude petroleum butadiene stream. In 1994, there was one Canadian commercial producer of butadiene (located in Sarnia, Ontario), with a domestic production of 103.7 kilotonnes. Importation into Canada from the USA was 1.7 kilotonnes in 1994. The Canadian domestic use of butadiene in 1994 amounted to 105.4 kilotonnes (98.3 kilotonnes for total domestic demand and 7.1 for export sales) (Camford Information Services, 1995). In the USA, total production in 1993 was 1.4 billion kilo grams.³ According to data summarized in IARC (1999), in 1996, production of butadiene in China (Taiwan), France, Germany, Japan, the Republic of Korea, and the USA was 129, 344, 673, 1025, 601, and 1744 kilo tonnes, respectively.

The largest end use of butadiene in Canada is the production of polybutadiene rubber (51.4 kilotonnes; 52.3% of total Canadian consumption for 1994) (Camford Information Services, 1995). Other derivatives produced include styrene-butadiene lattices (31.0 kilo tonnes; 31.5% of total Canadian consumption for 1994), nitrile-butadiene rubbers (10.0 kilotonnes; 10.2% for 1994), acrylonitrile-butadiene-styrene terpolymer (3.4 kilotonnes; 3.5% for 1994), and specialty styrene- butadiene rubbers (2.5 kilotonnes; 2.5% of total Cana dian consumption for 1994).

Butadiene has a long history of use, notably related to production of polymers. Several industrial and commercial products are manufactured with it or may contain it as a component. Examples include tires, car sealants, plastic bottles and food wrap, epoxy resins, lubricating oils, hoses, drive belts, moulded rubber goods, adhesives, paint, latex foams for carpet backing or underpad, shoe soles, moulded toys/household goods, medical devices, and chewing gum (CEH-SRI Interna tional, 1994; OECD, 1996).

Since butadiene is released primarily to air, its fate in that medium is of primary importance. Butadiene is not expected to persist in air, since it oxidizes rapidly with several oxidant species. Destruction of atmospheric butadiene by the gas-phase reaction with photochemi cally produced hydroxyl radicals is expected to be the dominant photo-initiated pathway. Products that can be formed include formaldehyde, acrolein, and furan. Destruction by nitrate radicals is expected to be a sig nificant nighttime process in urban areas. Acrolein, trans-4-nitroxy-2-butenal, and 1-nitroxy-3-buten-2-one have been identified as products of this reaction. Reaction with ozone is also rapid but less important than reaction with hydroxyl radicals. The products of the reaction of butadiene with ozone are acrolein, formalde hyde, acetylene, ethylene, formic acid, formic anhydride, carbon monoxide, carbon dioxide, hydrogen gas, hydro peroxyl radical, hydroxyl radical, and 3,4-epoxy-1- butene (Atkinson et al., 1990; Howard et al., 1991; McKone et al., 1993; US EPA, 1993).

Average atmospheric half-lives for photo-oxidation of butadiene, based on measured as well as calculated data, range from 0.24 to 1.9 days (Darnell et al., 1976; Lyman et al., 1982; Atkinson et al., 1984; Becker et al., 1984; Klöpffer et al., 1988; Howard et al., 1991; Mackay et al., 1993). However, half-lives for butadiene in air can vary considerably under different conditions. Estimations for atmospheric residence time in several US cities ranged from 0.4 h under clear skies at night in the summer to 2000 h (83 days) under cloudy skies at night in the winter. Daytime residence times for different cities within a given season varied by factors of 2–3. Nighttime residence times varied by larger factors. The differences between summer and winter conditions were large at all sites, with winter residence times 10–30 times greater than summer residence times (US EPA, 1993). Because of the long residence times under some conditions, especially in winter under cloudy conditions, there is a possibility of day-to-day carryover. Nonetheless, given the generally short daytime residence times, the net atmospheric lifetime of butadiene is short, and there is generally limited potential for long-range transport of this compound.

It is predicted from its physical/chemical properties that when butadiene is released into air, almost all of it will exist in the vapour phase in the atmosphere (Eisenreich et al., 1981; Environment Canada, 1998). Wet and dry deposition are not expected to be important as transfer processes. Evaporation from rain may be rapid, and the compound is returned to the atmosphere relatively quickly unless it is leached into the soil.

Volatilization, biodegradation, and oxidation by singlet oxygen are the most prominent processes involved in determining the fate of butadiene in water. The estimated half-lives of butadiene by reaction in water range from 4.2 to 28 days (Howard et al., 1991; Mackay et al., 1993).

The processes that are most prominent in deter mining the environmental fate of butadiene in sediment are biotic and abiotic degradation. The modelled half- lives of butadiene by reaction in sediment range from 41.7 to 125 days (Mackay et al., 1993).

Based on its vapour pressure and its solubility, volatilization of butadiene from soil and other surfaces is expected to be significant. Butadiene’s organic carbon/ water partition coefficient indicates that it should not adsorb to soil particles to a great degree and would be considered moderately mobile (Kenaga, 1980; Swann et al., 1983). However, the rapid rate of volatilization and the potential for degradation in soil suggest that it is unlikely that butadiene will leach into groundwater. Based on modelling predictions, the half-life of butadi ene by reaction, given by Howard et al. (1991) and Mackay et al. (1993), ranges from 7 to 41.7 days.

There are no measured bioconcentration factors. Butadiene is metabolized by the mixed-function oxidase system in higher organisms, which contributes to the expected lack of accumulation by many organisms. Estimated bioconcentration factors for butadiene in fish have been reported to range from 4.6 to 19 (Lyman et al., 1982; OECD, 1996). Even though estimation methods likely overestimate the true bioconcentration potential for a readily metabolized substance, they indicate that butadiene is not expected to bioconcentrate in aquatic organisms or to biomagnify in the aquatic food chain.

There are no reported measurements of plant root bioconcentration in soils. However, McKone et al. (1993) estimated the uptake of butadiene by roots from soil solution to be 1.84 litres/kg, which is the ratio of butadiene concentration in root (mg/kg, fresh mass) to the concentration in soil solution (mg/litre). The partition coefficient of butadiene concentration in roots (mg/kg, fresh mass) to concentration in soil solids (mg/kg) was estimated to range from 0.32 to 15 (dimensionless).

The partition coefficient of butadiene concentration in whole plants (mg/kg, fresh mass) to its concentra tion in soil solids (mg/kg) was estimated to range from 0.1 to 2.9 (dimensionless). The steady-state plant/air partition coefficient for foliar uptake of butadiene in plant leaves was estimated to be 0.63 m³/kg. There are no reported bioaccumulation data for any terrestrial invertebrates.

Fugacity modelling was conducted to provide an overview of key reaction, intercompartment, and advection (movement out of a system) pathways for butadiene and of its overall distribution in the environment. A steady-state, non-equilibrium model (Level III fugacity modelling) was run using the methods developed by Mackay (1991) and Mackay and Paterson (1991). Assumptions, input parameters, and results are presented in Environment Canada (1998). Based on butadiene’s physical/chemical properties, Level III fugacity model ling predicts that:

• when butadiene is released into air, the distribution of mass is almost 100% in air, with very small amounts in soil and water;
• when butadiene is released into water, the distribu tion of mass is 99.0% in water, with small amounts in air;
• when butadiene is released into soil, the distribu tion of mass is 38.6% in soil, 59.3% in air, and 2.1% in water.

Modelling predictions do not purport to reflect actual expected measurements in the environment but rather indicate the broad characteristics of the fate of the substance in the environment and its general distribution between media. Thus, when butadiene is discharged into air or water, most of it is expected to be found in the medium receiving the discharge directly. For example, if butadiene is discharged into air, almost all of it will exist in the atmosphere, where it will react rapidly and will also be transported away. If butadiene is discharged to water, it will react in water, and some will also evaporate into air. If butadiene is discharged to soil, most will be present in air or soil, where it will react (Mackay et al., 1993; Environment Canada, 1998).

Data on environmental levels and human exposure from Canada, the source country of the national assess ment on which this CICAD is based, are presented here as a basis for the sample risk characterization. Patterns of exposure in other countries are expected to be similar, although quantitative values may vary.

6.1.1 Ambient air

Butadiene was detected (detection limit 0.05 µg/m³) in 7314 (or 80%) of 9168 24-h samples collected between 1989 and 1996 from 47 sites across Canada.⁴ The mean concentration in all samples was 0.3 µg/m³ (in the calculation of the mean, a value of one- half the detection limit was assumed for samples in which levels were below the detection limit), and the maximum concentration measured was 14.1 µg/m³. Concentrations of butadiene in ambient air correspond ing to the 50th and 95th percentiles were 0.21 and 1.0 µg/m³, respectively. Concentrations were generally higher in urban areas, with a mean exposure to 0.4 µg/m³ (95th percentile 1.3 µg/m³) estimated as a "reasonable worst-case scenario," based on data from four sites. Similar levels were measured in smaller surveys in Canada (Bell et al., 1991, 1993; Hamilton- Wentworth, 1997; Conor Pacific Environmental, 1998).⁵ In areas influenced by industrial point sources of butadiene, concentrations in air were greater, with maximum and mean levels of 28 and 0.62 µg/m³, respectively (95th percentile 6.4 µg/m³) being measured between 1 and 3 km from the source (MOEE, 1995).

Butadiene has also been detected in air in enclosed structures. Concentrations of butadiene between 4 and 49 µg/m³ were measured during the winter months of 1994–1995 in Canadian underground parking garages (Environment Canada, 1994) because of its presence in vehicle exhaust. Similarly, butadiene was frequently detected in samples from 10 parking structures in Cali fornia, with the maximum concentration being 28 µg/m³ (Wilson et al., 1991). Butadiene has also been detected in urban road tunnels during rush hours in Australia (mean concentration 28 µg/m³; Duffy & Nelson, 1996) and Sweden (mean concentrations 17 µg/m³ and 25 µg/m³ in two tunnels; Barrefors, 1996). Butadiene was measured at concentrations ranging from 0.2 to 28 µg/m³ in 96 of 97 5-min air samples collected from a pumping island at randomly identified self-service filling stations in California (Wilson et al., 1991).

6.1.2 Surface water

No data on concentrations of butadiene in Cana dian lake, river, estuarine, or marine waters were identi fied in the literature. Butadiene is being monitored in effluents discharged into the St. Clair River from the butadiene production plant in Sarnia, Ontario. It was detected only twice, at 2 and 5 µg/litre, in 2103 compo site samples of aqueous effluent taken every 4 h in 1996 (detection limit 1 µg/litre). In daily sampling of effluents from the four individual outfalls (detection limit 1 µg/litre in 736 samples and 50 µg/litre in 789 samples), buta diene was detected in only three samples, at concentra tions of 21, 80, and 130 µg/litre.⁶

6.1.3 Groundwater

Butadiene was detected but not quantified in a groundwater plume near a waste site in Quebec where refinery oil residues and a variety of organic chemicals had been dumped (Pakdel et al., 1992).

6.2.1 Indoor air

In available surveys in Canada, 1,3-butadiene was detected up to 6 times more frequently in indoor air in homes than in corresponding samples of outdoor air, with concentrations being up to 10-fold higher indoors than outdoors (Bell et al., 1993; Hamilton-Wentworth, 1997; Conor Pacific Environmental, 1998).⁷ Concentrations in air of indoor environments are highly variable and depend largely on individual activities and circum stances, including the use of consumer products (e.g., cigarettes), the infiltration of vehicle exhaust from nearby traffic and possibly from attached garages, and cooking activities involving heated fats and oils (see section 6.2.3). While data are inadequate to determine the relative contributions of each of these potential indoor sources, the highest concentrations of butadiene in indoor air in Canada have generally been detected in indoor environments contaminated with environmental tobacco smoke (ETS). In a survey of 94 homes across Canada, the mean level in "non-smoking" homes was <1 µg/m³ (data censored by considering levels to be one- half the detection limit in samples in which butadiene was not detected), compared with a mean of 2.5 µg/m³ (data censored) in homes where smoking was present (Conor Pacific Environmental, 1998). Similarly, mean concentrations in indoor air from "non-smoking" loca tions in Windsor, Ontario, ranged from 0.3 to 1.6 µg/m³, while mean levels in "smoking" locations ranged from 1.3 to 18.9 µg/m³. At non-residential indoor sampling sites in Windsor, the frequency of detection of butadiene was 75–100% where ETS was present (Bell et al., 1993).

6.2.2 Drinking-water

There are no data available concerning the pres ence of butadiene in drinking-water. In an investigation on whether the use of polybutylene pipe in water distri bution systems is likely to result in the contamination of drinking-water with butadiene, Cooper⁸ did not detect the substance in water from these types of pipes (no further information was presented in the secondary account [CARB, 1992] of this study).

6.2.3 Food

There are no data available concerning the presence or concentrations of butadiene in food in Canada. In the USA, the migration of butadiene from rubber-modified plastic containers to food was inves tigated by McNeal & Breder (1987). Butadiene was detected in some of the containers, but was generally not detected in the foods (detection limits 1–5 ng/g). Similarly, in the United Kingdom, butadiene was not detected (detection limit 0.2 ng/g) in five brands of soft margarine, although its presence was demonstrated (at concentrations ranging from <5 to 310 ng/g) in the plastic containers (Startin & Gilbert, 1984). Butadiene has been detected in the emissions from heated cooking oils, including Chinese rapeseed, peanut, soybean, and canola oils, at levels ranging from 23 to 504 µg/m³ (Pellizzari et al., 1995; Shields et al., 1995).

6.2.4 Consumer products

Data on emissions of butadiene from potential indoor sources such as styrene-butadiene rubber were not identified.

Butadiene has been detected in both mainstream smoke and sidestream smoke from cigarettes in Canada and the USA. For 18 brands of Canadian cigarettes, the mean butadiene content ranged from 14.3 to 59.5 µg/cig arette (overall mean concentration 30.0 µg/cigarette) in the mainstream smoke and from 281 to 656 µg/cigarette (overall mean concentration 375 µg/cigarette) in the sidestream smoke, according to "preliminary" data (Labstat, Inc., 1995). The US DHHS (1989) reported that the vapour phase of mainstream smoke of non-filter cigarettes contained butadiene at levels of 25–40 µg/cig arette. Brunnemann et al. (1989) measured butadiene levels ranging from 16 to 75 µg/cigarette in mainstream smoke from seven brands of cigarettes and levels ranging from 205 to 361 µg/cigarette in the sidestream smoke from six types of cigarettes. As discussed in section 6.2.1, the presence of ETS contributes to elevated levels of butadiene in indoor air.

6.2.5 Occupational exposure

Potential occupational exposure to butadiene can occur in petroleum refining and related operations, pro duction of butadiene monomer, production of butadiene- based polymers, or the manufacture of rubber and plas tics products (IARC, 1999). Arithmetic mean concentra tions in petroleum and petrochemical operations in several European countries ranged from 0.1 to 6.4 mg/m³ during 1984–1987 (IARC, 1999; European Chemicals Bureau, 2001). Based on occupational hygiene surveys of butadiene production facilities in the United King dom, personal airborne exposures are generally below a mean concentration of 5 ppm (11 mg/m³), with most below 1 ppm (2.2 mg/m³). In polymer manufacture in the United Kingdom, most time-weighted average exposures are below 2–3 ppm (4.4–6.6 mg/m³). Similar concentra tions were reported in other facilities in the European Union (IARC, 1999). In monomer production facilities in the USA surveyed in 1985, arithmetic mean concen trations ranged from 1 to 277 mg/m³, while those in polymer production industries ranged from 0.04 to 32 mg/m³ (IARC, 1999).

The database on the toxicokinetics and metabolism of butadiene is relatively extensive. The proposed metabolism is outlined in Figure 1, based on the path ways described by Henderson et al. (1993, 1996) and Himmelstein et al. (1997). Available data for the path ways most extensively investigated indicate that metabo lism is qualitatively similar among the various species studied, although there may be quantitative differences in the amount of butadiene absorbed as well as in metabolic rates and the proportion of metabolites generated. These differences appear to be in concordance with the observed variation in sensitivity to butadiene-induced toxic effects of the few strains of rodent species tested to date, in that mice appear to metabolize a greater propor tion of butadiene to active epoxide metabolites than do rats. While less of these metabolites are also formed in samples of human tissues in vitro than in those of mice, available data are insufficient to characterize interindi vidual variability in humans. Although there are known genetic polymorphisms for a number of the enzymes involved in the metabolism of butadiene, information on genotype was not included in most investigations in humans.

Based on the metabolic pathways described in Figure 1, butadiene is first oxidized via cytochrome P- 450 enzymes (primarily P-450 2E1 in humans, although other isoforms may also be involved, the relative contrib utions of which vary between tissues and species) to the monoepoxide 1,2-epoxy-3-butene, or EB, which is sub sequently further oxidized via P-450 enzymes to the diepoxide 1,2,3,4-diepoxybutane, or DEB, or hydrolysed via epoxide hydrolase (EH) to butenediol (1,2-dihy droxy-3-butene). The monoepoxide, the diepoxide, and the butenediol may all be conjugated with glutathione (GSH) to form mercapturic acids, which are eventually eliminated in the urine. Hydrolysis of the diepoxide via epoxide hydrolase or oxidation of the butenediol via cytochrome P-450 will result in the formation of the monoepoxide diol (EBdiol). A small amount of buta diene may be converted to 3-butenal, which is subse quently transformed to crotonaldehyde (about 2–5% of the amount that is oxidized to the monoepoxide in human liver microsomes [Duescher & Elfarra, 1994] or microsomes of kidney, lung, or liver of B6C3F₁ mice [Sharer et al., 1992]). However, this pathway has not been extensively investigated, nor was crotonaldehyde detected in a sensitive analysis (using nuclear magnetic resonance spectroscopy) of urinary metabolites of rats and mice exposed to ¹³C-butadiene (Nauhaus et al., 1996).

Metabolism of butadiene and subsequent conver sion of EB to DEB may also take place to a more limited degree in the bone marrow (e.g., Maniglier-Poulet et al., 1995) by means other than P-450 oxidation (possibly via myeloperoxidase; Elfarra et al., 1996), based on in vitro observations and the detection of the epoxides in the bone marrow of rodents (Thornton-Manning et al., 1995a, 1995b), although this potential pathway has not yet been extensively investigated. EB may also react with both myeloperoxidase and chloride to form a chlorohydrin (1-chloro-2-hydroxy-3-butene) (Duescher & Elfarra, 1992). Metabolites arising from other possible pathways have been identified in the urine of mice exposed to butadiene (including metabolites known to be derived from metabolism of acrolein or acrylic acid) (Nauhaus et al., 1996), but no further research has yet been generated.

There is a substantial amount of evidence from in vitro and in vivo investigations that B6C3F₁ mice oxidize butadiene to the monoepoxide via P-450 in the liver to a greater extent than do Sprague-Dawley rats and humans. Levels of EB in the blood and other tissues of mice were two- to eightfold higher than those in rats exposed to similar levels of butadiene (Bond et al., 1986; Himmel stein et al., 1994, 1995; Bechtold et al., 1995; Thornton- Manning et al., 1997).

Available data also suggest that there are similar species differences in the amount of the diepoxide formed from oxidation of the monoepoxide. Levels of DEB were 40- to 160-fold higher in blood and other tissues of B6C3F₁ mice than in Sprague-Dawley rats exposed to the same concentration of butadiene (Thorn ton-Manning et al., 1995a, 1995b). While concentrations of EB at various sites were similar in male and female rats, levels of DEB were at least fivefold higher in females than in males, which correlates with the greater incidence of tumours in female rats. Although the mam mary gland is a target tissue in rats, extended exposure to butadiene at 8000 ppm (17 696 mg/m³) for 10 days did not result in any accumulation of DEB at this site (Thornton-Manning et al., 1998), which suggests that DEB may not play a significant role in the induction of mammary tumours in rats. Available in vitro data in human liver and lung samples suggest that humans also form less of the active metabolites of butadiene than do mice (although somewhat varying results have been reported with respect to the magnitude of the differences between species) (Csanády et al., 1992; Duescher & Elfarra, 1994; Krause & Elfarra, 1997).

Although epoxide metabolites of butadiene are formed to a greater extent in mice than in rats or humans, they are also cleared via glutathione conjugation more rapidly in mice (Kreuzer et al., 1991; Sharer et al., 1992; Boogaard et al., 1996a, 1996b). Conversely, hydrolysis of EB and DEB is greater in humans than in rats (based on in vitro data, as DEB has not been detected in tissues of exposed humans), and hydrolysis of EB and DEB in rats is in turn greater than that in mice (Csanády et al., 1992; Krause et al., 1997). In both humans and monkeys, removal of EB via hydrolysis appears to predominate over conjugation with glutathione, based on analysis of urinary metabolites (Sabourin et al., 1992; Bechtold et al., 1994). Although hydrolysis of the epoxide metabo lites is generally considered to be a detoxifying mecha nism, it may also lead to the formation of the diolepox ide, EBdiol, which is biologically reactive. However, no data were identified on species differences in the forma tion of EBdiol via metabolism of both epoxide metabo lites.

The formation of stable adducts of both the mono epoxide and monoepoxide diol metabolites of butadiene with the N-terminal valine of haemoglobin has been observed in experimental animals and humans exposed to butadiene (Albrecht et al., 1993; Osterman-Golkar et al., 1993, 1996; Neumann et al., 1995; Sorsa et al., 1996b; Tretyakova et al., 1996; Pérez et al., 1997).⁹ Consistent with the greater formation of epoxide metab olites, greater concentrations of haemoglobin–EB adducts were measured in mice than in rats exposed to the same concentration of butadiene. However, levels of haemoglobin–EB adducts in butadiene-exposed workers, although significantly elevated compared with levels in non-exposed workers, were considerably less than would be expected on the basis of results of studies in mice and rats (Osterman-Golkar et al., 1993). Based on observa tions in rats and humans exposed to butadiene, levels of haemoglobin–EBdiol adducts are substantially greater than levels of haemoglobin–EB adducts (although it is noted that the same adduct can result from binding with DEB). Metabolites of butadiene may also form adducts with DNA (see sections 8.5 and 9.2.3).

In addition to quantitative interspecies differences in the metabolism of butadiene, there is also evidence that there is significant variation within the human popu lation. Indeed, although available data are inadequate to assess interindividual variation in metabolism, which has been observed in in vitro investigations in microsomes from a small number of subjects (Boogaard & Bond, 1996; Krause et al., 1997), there has been significant interindividual variability in the extent of formation of haemoglobin adducts with butadiene metabolites in human populations (Neumann et al., 1995; Osterman- Golkar et al., 1996). Such variability is not unexpected, in view of the complexity of the metabolic pathways involved in the biotransformation of butadiene: i.e., the three principal enzymatic processes that determine the extent of exposure to the putatively toxic epoxide metabolites, namely formation via cytochrome P-450 2E1 and removal via epoxide hydrolase and glutathione conjugation. For example, the inducibility of cytochrome P-450 2E1 by low molecular weight compounds such as ethanol is likely to contribute to interindividual variabil ity in sensitivity. Moreover, genetic polymorphisms for glutathione-S-transferases and epoxide hydrolase might also contribute to considerable variation in sensitivity. While the influence of genotype for epoxide hydrolase has not been well investigated (although data indicate that hydrolysis of EB predominates over oxidation and glutathione conjugation in humans), interindividual sensitivity to the genetic effects of the epoxide metabo lites in in vitro studies has been clearly related to geno type for the glutathione-S-transferases (see section 9.2.3).

Although few data are available, butadiene appears to be of low acute toxicity in experimental animals, with reported LC₅₀ values for rats and mice of >100 000 ppm (>221 000 mg/m³). Lowest LC₅₀ values for butadiene are reported for mice, at 117 000 ppm (256 000 mg/m³) (duration not specified) (Batinka, 1966) and 121 000 ppm (268 000 mg/m³) (2 h) (Shugaev, 1969). The nervous system and the blood appear to be the principal targets; however, in only one study were data sufficient to determine a lowest-observed-effect level (LOEL) of 200 ppm (442 mg/m³) for haematological effects (Leavens et al., 1997). Exposure to butadiene for 7 h caused a concentration-dependent depletion (by as much as 80%) of cellular non-protein sulfhydryl content of liver, lung, or heart in mice, with a LOEL of 100 ppm (221 mg/m³) (Deutschmann & Laib, 1989). Depletion of non-protein sulfhydryl content may inhibit detoxification of epoxide metabolites vie glutathione conjugation.

No investigations in experimental animals on the potential for irritation or sensitization of butadiene have been identified.

The majority of short-term and subchronic studies were designed as either range-finding studies preliminary to chronic bioassays or investigations of potential mech anisms of action for butadiene-induced cancer and are not adequate for determination of critical effect levels. Effects on body weight were observed in B6C3F₁ mice exposed to 625 ppm (1383 mg/m³) butadiene or more for 2 weeks; no histopathological changes were noted at any concentration at or below 8000 ppm (17 696 mg/m³) (NTP, 1984).

Haematological effects consistent with megalo blastic anaemia and effects on bone marrow, including alterations in stem cell development, have been observed in two strains of mice (B6C3F₁ and NIH Swiss) exposed to 1000 or 1250 ppm (2212 or 2765 mg/m³) butadiene for up to 31 weeks (Irons et al., 1986a, 1986b; Leider man et al., 1986; Bevan et al., 1996). Other effects, including decreased survival and body weight gain (with males being more sensitive than females), altered organ weights, and ovarian or testicular atrophy, have also been observed in B6C3F₁ mice exposed subchronically to similar or higher levels of butadiene (NTP, 1984; Bevan et al., 1996). In addition, an increased incidence of a variety of tumours has been observed in B6C3F₁ mice exposed to 625 ppm (1383 mg/m³) butadiene for as little as 13 weeks (NTP, 1993) (see section 8.4). Although histopathological changes and haematological effects were reported in early studies in rats exposed to low concentrations (3 or 10 mg/m³) (Batinka, 1966; Ripp, 1967; Nikiforova et al., 1969), these results were not confirmed in more recent investigations of rats exposed for up to 13 weeks to much higher concentra tions (e.g., 17 600 mg/m³) (e.g., Crouch et al., 1979; Bevan et al., 1996). In view of the limitations of the studies in rats, it is not possible to draw any conclusions regarding species differences in response to subchronic exposure to butadiene.

The carcinogenic potential of inhaled butadiene has been studied in two strains of mice and one strain of rats. Butadiene was a multi-site carcinogen in all identified long-term experiments, inducing common and rare tumours in mice and rats, although there appear to be marked species and strain differences in sensitivity. In an early bioassay by the National Toxicology Program (NTP, 1984) in which male and female B6C3F₁ mice were exposed to 0, 625, or 1250 ppm (0, 1383, or 2765 mg/m³) butadiene for up to 61 weeks, there were exposure-related increases in the incidences of malignant lymphomas, cardiac haemangiosarcomas (an extremely rare tumour in B6C3F₁ mice), and lung tumours in both sexes. There were also increased incidences of papillo mas or carcinomas of the forestomach, hepatocellular adenomas or carcinomas, ovarian granulosa cell tumours, acinar cell carcinomas of the mammary gland, brain gliomas, and Zymbal gland carcinomas (the latter two tumour types have only rarely been observed in this strain of mice at the NTP) in one or both sexes.

Because of the poor survival of mice in the earlier bioassay and to better characterize the exposure– response relationship, the NTP (1993) subsequently exposed B6C3F₁ mice to lower concentrations (0, 6.25, 20, 62.5, 200, or 625 ppm [0, 13.8, 44.2, 138, 442, or 1383 mg/m³]) of butadiene for up to 2 years. Survival was again decreased in most groups (³ 20 ppm [³ 44.2 mg/m³]); at the highest concentration, deaths were principally due to lymphatic lymphomas, which appeared to arise from the thymus and occurred as early as week 23. Non-neoplastic effects observed in exposed mice included a variety of haematological effects, alter ations in organ weights, bone marrow atrophy and hyper plasia, atrophy of the thymus, atrophy and angiectasis of the ovaries, uterine atrophy, mineralization of the cardiac endothelium, liver necrosis, and olfactory epithelial atrophy. There were significant increases in the inci dence of tumours at a variety of sites (incidence data presented in Table 2), including malignant lymphomas (particularly lymphocytic lymphomas), histiocytic sarco mas, cardiac haemangiosarcomas, Harderian gland adenomas and carcinomas, hepatocellular adenomas and carcinomas, alveolar/bronchiolar adenomas and carcino mas, mammary gland adenoacanthomas, carcinomas, and malignant mixed tumours, ovarian granulosa cell tumours, and forestomach squamous cell papillomas and carcinomas, particularly when the incidences were adjusted for survival. The incidence of alveolar/bronchi olar adenomas or carcinomas was significantly increased in females at all concentrations (i.e., ³ 6.25 ppm [³ 13.8 mg/m³]). Low incidences of uncommon tumours, such as preputial gland carcinomas, Zymbal gland car cinomas in males, and renal tubule adenomas in both sexes, were also suspected of being related to exposure. In addition, exposure to butadiene induced malignant tumours at several sites, whereas, in general, tumours at the same sites in control animals were benign.

The NTP also conducted a "stop-exposure" experiment in male B6C3F₁ mice designed to investigate whether tumour induction was associated with the exposure concentration or the duration of exposure. Animals were exposed to 200 ppm (442 mg/m³) for 40 weeks or 625 ppm (1383 mg/m³) for 13 weeks (both equivalent to 8000 ppm-weeks) or to 312 ppm (690 mg/m³) for 52 weeks or 625 ppm (1383 mg/m³) for 26 weeks (both equivalent to 16 000 ppm-weeks); all groups were observed for the remainder of the 2-year study. Again, survival was reduced in all exposed mice, largely due to malignant neoplasms, with significant increases in the incidences of lymphocytic lymphomas, histiocytic sarcomas, cardiac haemangiosarcomas, Harderian gland adenomas or carcinomas, hepatocellular adenomas or carcinomas, alveolar/bronchiolar adenomas or carcinomas, and squamous cell papillomas or carcino mas of the forestomach (even in mice exposed for only 13 weeks) (incidence data presented in Table 2). In addition, low incidences of several uncommon tumour types (preputial gland carcinomas, Zymbal gland car cinomas, malignant gliomas and neuroblastomas of the brain, Harderian gland carcinomas, and renal tubule adenomas) were again observed in one or more of the exposed groups. Concentration may be more important than the duration of exposure in tumour development, as the incidence of malignant lymphomas and squamous cell carcinomas of the forestomach was greater in the groups that had been exposed to 625 ppm (1383 mg/m³) for a shorter period than in those exposed to 200 ppm (442 mg/m³) for a longer period (i.e., similar total cumu lative exposure) (NTP, 1993).

Acute exposure of B6C3F₁ mice for 2 h to up to 10 000 ppm (22 120 mg/m³) butadiene, followed by observation for 2 years, did not induce an increased incidence of tumours at any site (Bucher et al., 1993).

Sensitivity to butadiene-induced thymic lympho ma/leukaemia appears to be enhanced by the presence of an endogenous ecotropic retrovirus in B6C3F₁ mice, as the incidence of this tumour was greater in male B6C3F₁ mice exposed to 1250 ppm (2765 mg/m³) butadiene for 52 weeks than in male Swiss mice, which do not express an endogenous retrovirus (57% versus 14%). Exposed mice of both strains had elevated incidences of thymic lymphoma/leukaemia compared with controls, as did B6C3F₁ mice exposed to 1250 ppm (2765 mg/m³) for 12 weeks and then observed for an additional 40 weeks, although the MuLV env sequence for the retrovirus was detected only in tumours of the B6C3F₁ mice. Other tumours reported in the mice exposed for 52 weeks included haemangiosarcomas of the heart (mainly in B6C3F₁ mice) and lung tumours. Neoplasms of the glandular and non-glandular stomach were observed in the B6C3F₁ mice, whereas adenocarcinomas of the Harderian gland and the thyroglossal duct were observed in the Swiss mice (Irons et al., 1989).

In the only identified long-term bioassay in rats (Hazleton Laboratories Europe Ltd., 1981a; Owen et al., 1987; Owen & Glaister, 1990), male and female Sprague-Dawley rats were exposed to 0, 1000, or 8000 ppm (0, 2212, or 17 696 mg/m³) butadiene for up to 111 weeks. At 8000 ppm (17 696 mg/m³), survival was reduced in both sexes; there were also changes in the relative weights of a number of organs in males at this concentration, along with an increase in the severity of nephrosis of the kidney relative to controls. Relative liver weights were increased in all exposed groups, although there were no exposure-related histopathologi cal effects on the liver. At 8000 ppm (17 696 mg/m³), there were increased incidences of follicular cell adeno mas and carcinomas of the thyroid gland in females and exocrine adenomas of the pancreas in males (with a carcinoma occurring in a rat of either sex) (incidence data presented in Table 3). In females, the incidence of benign or malignant mammary gland tumours, along with the incidence of animals with multiple mammary gland tumours, was increased at both 1000 and 8000 ppm (2212 and 17 696 mg/m³). The incidence of sarcomas of the uterus and carcinomas of the Zymbal gland increased significantly with level of exposure in females; in addi tion, a Zymbal gland carcinoma occurred in one male rat at each exposure level. The incidence of Leydig cell tumours of the testes was increased in both groups of exposed males. The investigators suggested that the occurrence of tumours of the testes and Zymbal gland may have been unrelated to exposure, as the incidences observed were reportedly similar to those in other control rats of the same strain in the study laboratory, although it is noted that Zymbal gland tumours were noted in the chronic bioassays in mice discussed above.

Both the mono- and diepoxide metabolites (EB and DEB) have induced local tumours at the site of applica tion in Swiss mice or Sprague-Dawley rats (Van Duuren et al., 1963, 1965, 1966), although available studies are inadequate to evaluate species differences in sensitivity.

It has been hypothesized that the observed greater sensitivity of B6C3F₁ mice compared with Sprague- Dawley rats to the induction of thymic lymphoma by butadiene may be related to differences in the potential of EB to affect haematopoietic stem cell differentiation observed in in vitro investigations, as suppression of clonogenic response was greater in bone marrow cells from C56BL/6 mice than in those from Sprague-Dawley rats or humans; it was also hypothesized that the sub population of progenitor cells affected in mice is not present in humans (Irons et al., 1995).

The genotoxicity of butadiene has been investigated in a limited range of in vitro assays and a more extensive range of in vivo tests. Butadiene was mutagenic in Salmonella typhimurium strains TA1530 and TA1535 in the presence of metabolic activation with rodent or human S9 preparations (de Meester et al., 1978, 1980; Arce et al., 1990; NTP, 1993; Araki et al., 1994), although it was generally inactive in strains TA97, TA98, and TA100 with or without exogenous activation under similar experimental conditions (Victorin & Ståhlberg, 1988; Arce et al., 1990; NTP, 1993). Results of mouse lymphoma assays have been conflicting, with an increased frequency of mutations at the tk locus in one study at very high concentrations (i.e., 200 000–800 000 ppm [442 400–1 796 600 mg/m³]) in the presence of metabolic activation (Sernau et al., 1986), while there was no convincing activity at concen trations of up to 300 000 ppm (663 600 mg/m³) in another study (although the authors noted that the lack of a positive response may have been due to the low solubility of butadiene in the culture medium; NTP, 1993). Butadiene dissolved in ethanol induced sister chromatid exchanges in cultured mammalian cells (hamsters and humans) (Sasiadek et al., 1991a, 1991b), while in vitro exposure to gaseous butadiene did not induce this effect in preparations from rats, mice, and humans (Arce et al., 1990; Walles et al., 1995).

An overview of the results of available in vivo assays for genotoxicity in germ and somatic cells in mice and rats is presented in Table 4; in general, the data are consistent with species-specific differences in sensitivity to butadiene-induced genetic damage, likely related to the quantitative differences in the formation of active metabolites, although fewer studies have been conducted in rats. Butadiene induced dominant lethal mutations in two strains of mice (CD-1 and (102/E1 × C3H/E1)F₁) following short-term or subchronic exposure of males to concentrations as low as 500 ppm (1106 mg/m³) for 5 days or 65 ppm (144 mg/m³) for 4 weeks; however, exposure to 6250 ppm (13 825 mg/m³) for 6 h did not induce dominant lethal mutations in CD-1 mice. The results of these studies, which depended upon the timing of mating relative to exposure, suggested that the induc tion of dominant lethal mutations in mice was likely caused by effects on mature germ cells. In the only simi lar study in rats identified, there was no evidence of dominant lethal mutations in Sprague-Dawley rats exposed to up to 1250 ppm (2765 mg/m³) butadiene for 10 weeks.

Short-term exposure to 500 or 1300 ppm (1106 or 2876 mg/m³) butadiene also induced an exposure-related increase in the incidence of heritable chromosomal

translocations in mice; an increased incidence of chromosomal aberrations was also noted in zygotes of male mice exposed to ³ 500 ppm (³ 1106 mg/m³) for 5 days. Other butadiene-induced effects observed in male germ cells of mice include sperm head abnormal ities, micronuclei in spermatids, and DNA damage (strand breaks and alkaline-labile sites). Investigations of these end-points in rats have not been identified.

Butadiene was consistently genotoxic in somatic cells of several strains of mice, inducing chromosomal aberrations, sister chromatid exchanges, and micronuclei in numerous assays; micronuclei have been observed following exposure to concentrations as low as 6.25 ppm (13.8 mg/m³) butadiene for 13 weeks or 62.5 ppm (138 mg/m³) for 8 h. Although only few studies were identified, these effects were not observed in rats exposed to much higher concentrations. However, gene mutations at the hprt locus have been induced in both mice and rats, with a four- to sevenfold greater muta genic potency being determined for mice than for rats. Mutagenic activity was also observed in two transgenic mouse systems and in the mouse spot test. Binding to DNA has been observed in all strains of mice and rats tested; following exposure to butadiene, adducts of both guanine and adenine with the monoepoxide as well as the monoepoxide diol metabolites (EB and EBdiol, respec tively) have been observed. The degree of adduct forma tion was generally similar in the two species or, in some studies, up to twofold greater in mice than in rats. Similarly, there was little quantitative difference in the amount of butadiene-induced single strand breaks in DNA of mice and rats. DNA–DNA and DNA–protein cross-links were noted in one of two studies in mice, but not in rats exposed to higher concentrations of butadiene.

Metabolites of butadiene have also been mutagenic and clastogenic in numerous in vitro and in vivo assays (see Table 4 for overview of results of in vivo assays). EB, DEB, and EBdiol all induced mutations in bacteria and yeast in the absence of exogenous metabolic activa tion (IARC, 1992; NTP, 1993; Thier et al., 1994; Adler et al., 1997); mutagenic activity was also observed for all three metabolites at two foci in human TK6 lymphoblas toid cells, with DEB being much more potent (Cochrane & Skopek, 1993, 1994a). Conversely, the monoepoxide was much more potent than the diepoxide in the induc tion of mutations at the lacI transgene of fibroblasts obtained from a transgenic rat strain (Saranko & Recio, 1998; Saranko et al., 1998). Both EB and DEB also induced sister chromatid exchanges, chromosomal aberrations, and micronuclei in cultured mammalian (including human) cells (IARC, 1992; Xi et al., 1997). Aneuploidy in chromosomes 12 and X was also induced in human lymphocytes, which is notable in view of the fact that aneuploidy in these chromosomes is commonly observed in lymphocytic leukaemias (Xi et al., 1997). In vitro exposure to DEB, but not EB or EBdiol, induced micronuclei in spermatids isolated from rats (Sjöblom & Lähdetie, 1996).

The monoepoxide, diepoxide, and monoepoxide diol metabolites all induced micronuclei in germ cells of male mice and rats; in one of these studies, the magni tude of the effect was greater in Lewis rats than in F₁(102 × C3H) mice. There were no consistent patterns in the relative potency of the three metabolites. Chromosomal aberrations in zygotes produced by exposed males and dominant lethal mutations were induced by DEB in mice (strains (C57Bl/Cne × C3H/Cne)F₁ and (102/E1 × C3H/ E1)F₁), respectively), whereas EB and EBdiol did not induce dominant lethal mutations. In the only identified investigation of the potential effects on female germ cells, pre-mating exposure of female B6C3F₁ mice to DEB resulted in an increased frequency of chromosomal aberrations in embryos in the absence of ovarian toxicity.

EB, DEB, and EBdiol were also genotoxic in somatic cells (bone marrow, peripheral blood, lung, and spleen), inducing sister chromatid exchanges, chromo somal aberrations, or micronuclei in several strains of mice, rats, and hamsters, with little consistent evidence of interspecies differences in sensitivity; in general, the diepoxide was more potent than the monoepoxide or the monoepoxide diol. Although negative results were obtained in Lewis rats, both EB and DEB induced an increased frequency of hprt^– mutations in B6C3F₁ mice and F344 rats, with rats being more sensitive than mice, which may be related to slower clearance in rats. EB induced mutations in the bone marrow of lacI transgenic rats, but not in lacI transgenic mice; DEB did not induce lacI mutations in either species. Meng et al. (1999) suggested that the hprt assay is more sensitive to the detection of large deletions induced by DEB than the lacI transgene assay. DNA damage (strand breaks or alkaline-labile sites) was caused by EB and DEB in the bone marrow of rats and mice, with DEB being less potent than EB; the only damage observed in haploid testicular cells was in mice exposed to EB. It was suggested that the apparent greater potency of EB compared with DEB may be due to the bifunctional alkylating ability of DEB, subsequent induction of DNA repair, and the inability of the alkaline Comet assay employed to measure cross-links (Anderson et al., 1997).

8.6.1 Effects on fertility

Few data on the effects of butadiene on reproduc tive ability were identified. Exposure to up to 1300 ppm (2876 mg/m³) for 5 days did not affect the reproductive abilities of male (102/E1 × C3H/E1)F₁ mice, based on percentages of successful pairings with unexposed females and unfertilized metaphase I oocytes (Pacchi erotti et al., 1998a). Similarly, there were no decreases in mating frequency or pregnancy rate in the dominant lethal studies in mice and rats (Anderson et al., 1993, 1998; BIBRA International, 1996a, 1996b; Brinkworth et al., 1998). Documentation of an earlier study in rats, guinea-pigs, and rabbits (Carpenter et al., 1944) is too limited for evaluation.

The reproductive organs have consistently been targets of non-neoplastic effects induced by butadiene in subchronic and long-term bioassays in B6C3F₁ mice but not in Sprague-Dawley rats, although butadiene-induced tumours of the reproductive organs have been observed in both species. Ovarian atrophy and decreased weight were observed in mice exposed to 1000 ppm (2212 mg/ m³, the only concentration tested) for 13 weeks (Bevan et al., 1996). In the 2-year bioassay conducted by the NTP, there was a significant increase in the incidence of ovar ian atrophy in females exposed for up to 2 years to all concentrations tested (i.e., ³ 6.25 ppm [³ 13.8 mg/m³]); both the incidence and the severity of this lesion increased with exposure. Ovarian atrophy was also observed at the interim sacrifices at 9 and 15 months at higher concentrations (³ 200 and ³ 62.5 ppm [³ 442 and ³ 138 mg/m³], respectively). Atrophied ovaries charac teristically had no evidence of oocytes, follicles, or corpora lutea. Angiectasis and germinal epithelial hyperplasia of the ovaries were reported at ³ 62.5 and ³ 200 ppm (³ 138 and ³ 442 mg/m³), respectively, after exposure for 2 years. Uterine atrophy was also noted at concentrations of 200 ppm (442 mg/m³) or greater. Survival was decreased at ³ 20 ppm (³ 44.2 mg/m³), principally due to neoplastic lesions at several sites, including the ovaries (Melnick et al., 1990; NTP, 1993).

Effects on the testes, including reduced weight, degeneration, or atrophy, were observed in B6C3F₁ mice exposed to concentrations at or above 200 ppm (442 mg/m³) for 2 years or to higher levels for shorter durations (NTP, 1993; Bevan et al., 1996). Cytotoxic effects on differentiating spermatogonia were noted in (102/E1 × C3H/E1)F₁ mice 21 days after exposure to ³ 130 ppm (³ 288 mg/m³) for 5 days; a decrease in elongated spermatids was noted in mice exposed to 1300 ppm (2876 mg/m³) (Pacchierotti et al., 1998a).

No non-neoplastic effects were noted in the repro ductive organs of male or female Sprague-Dawley rats exposed to up to 8000 ppm (17 696 mg/m³) butadiene for 2 years (Hazleton Laboratories Europe Ltd., 1981a; Owen et al., 1987).

Both the mono- and diepoxide metabolites of butadiene induced ovarian toxicity (depletion of small and growing follicles) and alkylation with macromole cules in the ovary in B6C3F₁ mice repeatedly exposed via intraperitoneal injection. In contrast, effects in the ovary in Sprague-Dawley rats were observed only in rats exposed to the diepoxide at doses higher than those that were active in the mice (Doerr et al., 1996). Since the results of structure–activity studies with 4-vinylcyclo hexene and several of its analogues, butadiene mono epoxide and diepoxide, epoxybutane, and isoprene indicated that compounds that form only monoepoxides do not induce ovarian toxicity (Doerr et al., 1995), it appears that conversion to the bifunctional diepoxide may be required for the induction of these effects. A single intraperitoneal injection of DEB reduced various testicular cell populations and induced morphological changes in the epithelium of the seminiferous tubules in male B6C3F₁ mice (Spano et al., 1996).

8.6.2 Developmental toxicity

Few studies on the potential for butadiene to induce developmental effects have been identified. There was no evidence of teratogenicity following exposure of pregnant CD-1 mice to up to 1000 ppm (2212 mg/m³) butadiene on days 6 through 15 of gestation, although maternal toxicity (decreased body weight gain) and fetal toxicity (reduced fetal weight and skeletal abnormalities) occurred at 200 ppm (442 mg/m³) and above, and there was a slight reduction in male fetal body weight at 40 ppm (88 mg/m³) of questionable biological significance (Hackett et al., 1987b; Morrissey et al., 1990). In Sprague-Dawley rats exposed to 8000 ppm (17 696 mg/m³) butadiene on days 6 through 15 of gestation, there was an increased incidence of "major" abnormalities of the skull, spine, sternum, long bones, and ribs. Abnormalities believed to be associated with retarded embryonic growth were observed at 200 and 1000 ppm (442 and 2212 mg/m³). Maternal toxicity (decreased body weight gain or loss of body weight) was observed in all exposed groups (Hazleton Laboratories Europe Ltd., 1981b, 1982). However, there was no evidence of developmental toxicity in Sprague-Dawley rats exposed to up to 1000 ppm (2212 mg/m³) butadiene, also on days 6 through 15 of gestation, although maternal toxicity (decreased body weight gain) was noted at the highest concentration (Hackett et al., 1987a; Morrissey et al., 1990).

Although evidence of male-mediated teratogenicity was observed when male CD-1 mice exposed to 12.5 ppm (27.7 mg/m³) butadiene for 10 weeks were mated with unexposed females (Anderson et al., 1993), there was no increase in malformations when the study was repeated at 12.5 and 125 ppm (27.7 and 277 mg/m³) (Brinkworth et al., 1998). The authors suggested that the discrepant results may be a function of the statistical significance in the first study being due to the lack of abnormalities in controls (compared with 2.5% in exposed), whereas a low incidence was noted in exposed and control mice in the follow-up study. Similarly, there were no significant increases in fetal abnormalities in CD-1 mice following paternal exposure to up to 6250 ppm (13 825 mg/m³) butadiene for 6 h (Anderson et al., 1993) or concentrations up to 130 ppm (288 mg/ m³) for 4 weeks (BIBRA International, 1996a), although it was noted in the latter study that some females may have been sacrificed too early for detection of abnor malities. There was no evidence of male-mediated teratogenicity in offspring of male Sprague-Dawley rats exposed for 10 weeks to concentrations as high as 1250 ppm (2765 mg/m³) butadiene and then mated with unexposed females (BIBRA International, 1996b).

Although the haematopoietic system is a target of butadiene-induced toxicity, no effects on immune system function of biological significance were observed in the only relevant study identified in which B6C3F₁ mice were exposed to 1250 ppm (2765 mg/m³) butadiene for up to 24 weeks, although there were depressions in cellu larity and plaque-forming cells as well as histopathologi cal changes in the spleen (Thurmond et al., 1986).

Slight irritation of the eyes or respiratory tract was observed in volunteers exposed to very high concentra tions of butadiene (i.e., >2000 ppm [>4400 mg/m³]) in the few early clinical investigations identified (Larionov et al., 1934; Carpenter et al., 1944); however, these studies are inadequate for evaluation of the potential effects of butadiene in humans, as only subjective symp toms appear to have been evaluated.

9.2.1 Cancer

The carcinogenicity of butadiene has been investigated in several populations of workers occupationally exposed during its manufacture or use. Although most of these studies are limited by the paucity of historical monitoring data, there is evidence that occupational exposure to butadiene in the styrene- butadiene rubber industry is associated with excess mortality due to leukaemia and weaker evidence of an association with lymphosarcoma¹⁰ in butadiene monomer production workers. A summary of the measures of risk for lymphohaematopoietic cancers is presented in Table 5.

In the most recent update of the largest of the cohorts of male monomer workers (n = 2795) at the Port Neches butadiene production facility in Texas, USA (Divine & Hartman, 1996), mortality due to lympho haematopoietic cancer was significantly elevated (standardized mortality ratio [SMR]¹¹ = 147; 95% confidence interval [CI] = 106–198), due largely to a non-significant increase in the number of deaths due to lymphosarcoma and reticulosarcoma (SMR = 191, 95% CI = 87–364), based on nine cases. However, there was no association with duration of employment (SMRs of 261, 182, and 79, for <5, 5–19, and ³ 20 years of employment, respectively, based on six, two, and one cases). The greatest excess was observed in men first employed during the Second World War (observed cases [O] = 7; SMR = 241; 95% CI = 97–497), during which the greatest concentrations of butadiene likely occurred, although no data were presented (Divine et al., 1993). When the entire cohort was subdivided on the basis of a qualitative measure of exposure (based on job history information, discussions and reviews of classifications with long-term employees, and recent industrial hygiene surveys), the SMR for lymphosarcoma and reticulo sarcoma was greatest (SMR = 249; 95% CI = 100–513) in those whose jobs involved "varied" exposure to butadiene (i.e., the group with potentially the greatest exposure), although the number of cases in each sub group was small, and only one of the seven cases in this group had been exposed for more than 10 years. There was also a non-significant increase in mortality due to leukaemia in the varied exposure group (SMR = 154; 95% CI = 77–275; based on 11 cases, 10 of which had been employed less than 10 years), although there was no such increase in the total cohort (SMR = 113; 95% CI = 60–193). However, there was no association between estimates of cumulative exposure and any form of lymphohaematopoietic cancer.

Mortality due to lymphosarcoma and reticulo sarcoma was significantly increased in a smaller cohort study of 364 workers who had ever been employed in butadiene production at two Union Carbide plants in West Virginia, USA, based on four cases (SMR = 5.77; 95% CI = 1.57–14.8). The risk was greater in those employed for 2 years or more than in those employed for less than 2 years (SMRs of 8.27 and 3.03, respectively), although the number of cases in each category was very small. No significant increase in mortality due to leukaemia or aleukaemia was observed (observed/ expected = 2/1.62). However, no monitoring data were available to assess exposure of individuals in the cohort (E.M. Ward et al., 1995, 1996). In the only other study of monomer production workers, there were no deaths due to cancer of the lymphohaematopoietic system; however, the size of the cohort (n = 614) and duration of follow-up were insufficient to detect excess risks of lymphohaematopoietic cancer of less than fivefold (Cowles et al., 1994).

Several studies have been conducted in workers employed in the manufacture of synthetic rubber in North America who were exposed to butadiene as well as to styrene and other substances. The largest and most comprehensive study to date involved 15 649 workers employed at eight styrene-butadiene rubber manufactur ing facilities in North America (Delzell et al., 1995). The results of this study are emphasized here and considered to supplant those of earlier investigations, as there is considerable overlap in the cohort population with the earlier studies (i.e., 14 869 of these subjects had been employed at one of the two plants studied previously by Meinhardt et al. [1982] or at seven of the eight plants investigated by Matanoski et al. [1990, 1993] and Santos-Burgoa et al. [1992], although they had been employed for different time periods [Delzell et al. (1995) included several more years of follow-up] and selected using different inclusion criteria).¹² Estimates of cumula tive exposure and peak exposure frequency were derived for workers from six of the eight plants based on com plete work histories for 97% of these employees, infor mation on processes and plant conditions based on available records, walk-through surveys, and interviews with long-term employees, plant engineers, and man agers and were compared with monitoring data from surveys conducted from the late 1970s onward.

There was an increase in mortality due to leukae mia, which was of borderline statistical significance, in the overall cohort, based on 48 cases (SMR = 131; 95% CI = 97–174); this excess was concentrated in workers who had had 10 or more years of employment and 20 or more years since date of hire (O = 29; SMR = 201; 95% CI = 134–288). Similarly, there was a significant increase in mortality due to leukaemia in "ever hourly" workers (i.e., workers who had ever been paid on an hourly basis) whose jobs were most likely to have involved exposure to butadiene (O = 45; SMR = 143; 95% CI = 104–191), which was again concentrated in workers with longer duration of employment and time since hire and was greater in black workers than in white workers in this subgroup. The SMRs for leukaemia also increased with duration of employment for ever hourly workers. When examined by type of employment, the number of deaths due to leukaemia was significantly increased in production workers (O = 22; SMR = 159; 95% CI = 100–241), labourers (O = 16; SMR = 195; 95% CI = 112–317; concentrated among black workers), laboratory workers (O = 12; SMR = 462; 95% CI = 238–806), and black workers in other operations (O = 3; SMR = 680; 95% CI = 137–1986); no significant increases were observed in maintenance workers (O = 13; SMR = 107; 95% CI = 57–184). (Although analyses were not presented and no further information is given, Delzell et al. [1995], in the Discussion section of their report, reported that there was no increase in mortality due to leukaemia in 851 workers in butadiene production areas [which would not involve exposure to styrene], with very low numbers of observed and expected deaths [1 versus 2.1, respectively].) As well, the SMRs for lymphosarcoma were non-significantly increased for maintenance workers (O = 8; SMR = 192; 95% CI = 83–379) and labourers (O = 3; SMR = 123; 95% CI = 25–359), while there was no increase in mortality due to lymphosarcoma in production workers or laboratory workers. When individual plants were considered separately, there were non-statistically significant increases in mortality due to leukaemia at most (but not all) plants (SMRs ranged from 72 to 780, excluding groups in which zero cases were observed and less than one case was expected); numbers of observed cases of lymphosarcoma were too low to permit meaningful conclusions with respect to mortality at individual plants.

In regression analyses, mortality due to leukaemia was observed to increase with cumulative exposure to butadiene, as relative risk (RR) values for exposure categories of 0, >0–19, 20–99, 100–199, and ³ 200 ppm- years were 1.0, 1.4, 2.3, 2.6, and 4.2, respectively (for cases in which leukaemia was considered the underlying cause of death). There was only limited evidence of an association with cumulative exposure to peak levels of butadiene. The authors also investigated the potential influence of exposure to styrene or benzene on mortality and determined that the trend for increased risk with increased cumulative exposure to styrene was less pronounced, while exposure to benzene was considered to be too infrequent (few subjects were exposed) and too low to be a confounding factor. There was no association between cumulative exposure to butadiene and non- Hodgkin’s lymphoma in regression analyses.

Based on the results of this study, Delzell et al. (1995) concluded that there was a relationship between employment in the styrene-butadiene industry and leukaemia, with the increased risk of leukaemia being most strongly associated with exposure to butadiene or to butadiene and styrene in combination (although the association with butadiene remained after controlling for exposure to styrene). Data were insufficient to draw any firm conclusions with respect to an association with any specific form of leukaemia.

In a subsequent study in which Delzell et al. (1996) attempted to better define exposure of this cohort to peak levels of butadiene, the RR for leukaemia increased with increasing average annual number of peaks to which workers were exposed (RRs of 1.0, 2.3, and 3.1 for 0, >0–3288, and >3288 peaks), as well as again with cumulative exposure to butadiene (RRs of 1.0, 1.1, 2.0, 2.4, and 4.6 for 0, >0–19, 20–99, 100–199, and ³ 200 ppm-years). Although the analyses were not presented, adjusting for cumulative exposure to styrene apparently had little influence on the exposure–response relation ship. Risk of leukaemia also increased with duration of employment in areas in which there was "definite" expo sure to peaks (RRs of 1.0, 2.3, and 2.7 for 0, >0–4, and >5 years) and in areas for which elevated SMRs had been noted in the previous analyses (RRs of 1.0, 1.9, and 3.1 for 0, >0–4, and >5 "high SMR-years"). The authors noted that it was not possible to distinguish between the roles of estimated peak or cumulative exposure.

The estimates of exposure were further refined for workers at one of the plants included in the investigation by Delzell et al. (1995) through more extensive research of historical conditions (Macaluso et al., 1997). Although there was little change in classification of various workers as exposed or non-exposed, the revised estimates of cumulative exposure to butadiene for many job groups were generally greater (two- to threefold) than the original; the most substantial increase (by an order of magnitude) was determined for tasks among unskilled labourers during the 1950s and 1960s. It was not indicated in the report if the rank order of the cumu lative exposure estimates differed (although it is likely that it did not; Gerin & Siemiatycki, 1998). There was little change in estimated exposure to peak levels of butadiene or in cumulative exposure to styrene. These revised exposure estimates have not yet been incorpor ated into cancer mortality analyses.

Sathiakumar et al. (1998) re-examined the mortal ity of this cohort based on currently accepted terminol ogy for lymphohaematopoietic cancers (other than leukaemia). There were no significant increases in deaths in the overall cohort due to non-Hodgkin’s lymphoma, Hodgkin’s disease, multiple myeloma, or cancers of other lymphatic tissue, nor were there any associations between mortality due to these causes and duration of exposure and year of hire. Similarly, mortality due to these causes was not associated with any process group; however, the authors noted that an association for non- Hodgkin’s lymphoma may be obscured by the possibility that some cases of non-Hodgkin’s lymphoma had trans formed to leukaemia, with the latter form of cancer being recorded on the death certificate.

An association between exposure to butadiene and leukaemia, as well as Hodgkin’s disease, was also observed in a recent independently conducted nested case–control study of 58 cases of lymphohaematopoietic cancers from a cohort of styrene-butadiene rubber workers (from many of the same plants investigated by Delzell et al. [1995]), in which exposure was estimated based on analyses of monitoring data obtained in the last 15–20 years of operation (Matanoski et al., 1997).

Irons & Pyatt (1998) noted the concordance between the potential for exposure to dimethyldithiocar bamate (a "stopper" used in the polymerization process in the styrene-butadiene rubber industry) and mortality due to leukaemia in various processes. However, although the biological plausibility of a potential association is recognized (since dimethyldithiocarbamate is a potent inhibitor of clonogenic response in human CD34⁺ bone marrow cells), it is not possible to draw any conclusion concerning its potential role in the observed increases in leukaemia mortality at this time in view of the current lack of quantification of exposure levels of this substance in the plants examined and the absence of data on its leukaemogenic potential.

Other epidemiological studies in populations occupationally exposed to butadiene have been identified in the literature (e.g., McMichael et al., 1974, 1976; Andjelkovich et al., 1976, 1977; Linet et al., 1987; Siemiatycki, 1991; Bond et al., 1992; Downs et al., 1993). However, owing to limitations of these studies (including small numbers of observed and expected cases of lymphohaematopoietic cancer or lack of expo sure characterization), they contribute little to evaluation of the association between exposure to butadiene and these forms of cancer.

Significantly increased mortality due to cancers other than those of the lymphohaematopoietic system has not been consistently observed in these studies.

9.2.2 Non-neoplastic effects

Mortality due to all causes was similar to or significantly lower than that expected in all of the major cohorts of workers potentially exposed to butadiene. Although increased mortality due to arteriosclerotic or ischaemic heart disease or circulatory disease in general has been observed in some subgroups of workers in some of these cohorts (McMichael et al., 1974, 1976; Matanoski et al., 1990; Delzell et al., 1995), the potential association with exposure to butadiene has not been extensively investigated.

There were no differences in morbidity or various haematological parameters between 438 workers exposed to mean concentrations of butadiene ranging up to 10 ppm (22 mg/m³) (with a maximum time-weighted average concentration of 143 ppm [316 mg/m³]) and 2600 unexposed workers at a butadiene production facility in Texas, USA (Cowles et al., 1994). However, Checkoway & Williams (1982) observed changes in haematological parameters consistent with bone marrow depression in eight workers exposed to high concentra tions of butadiene (up to about 53 ppm [117 mg/m³]) when compared with values for 145 workers exposed to much lower levels (i.e., <1 ppm [<2.2 mg/m³]).

9.2.3 Genotoxicity

The potential genotoxicity of butadiene has recently been investigated in several studies of groups of workers exposed in the production of butadiene, styrene- butadiene rubber, or polybutadiene rubber. Although the data available to date are not completely consistent, they indicate that there is some evidence that exposure to butadiene induces genetic effects in occupationally exposed populations and that sensitivity to the induction of these effects is related to genetic polymorphism for enzymes involved in the metabolism of butadiene, most notably those within the glutathione-S-transferase class. The results of several in vitro studies in human lympho cytes have demonstrated that sensitivity to DEB-induced sister chromatid exchanges and micronuclei is associated with the presence or absence of homozygous deletion of the GSTT1 gene, which codes for GSTtheta (Kelsey et al., 1995; Norppa et al., 1995; Wiencke et al., 1995; Landi et al., 1996; Pelin et al., 1996; Vlachodimitropoulos et al., 1997), for which the prevalence of the null genotype is reported to be between 15 and 30% (Nelson et al., 1995; Abdel-Rahman et al., 1996; Bailey et al., 1998). Simi larly, sensitivity to sister chromatid exchanges induced by EB appears to be related to genotype for GSTM1, which codes for GSTµ (Wiencke & Kelsey, 1993; Uuskula et al., 1995), and possibly also GSTT1 genotype in GSTM1-null individuals (Bernardini et al., 1998). However, there were no differences in sensitivity to sister chromatid exchanges induced by EBdiol in individuals with and without deletions for GSTT1 or GSTM1 (Bernardini et al., 1996).

Although no increased frequencies of sister chro matid exchanges, chromosomal aberrations, or micro nuclei were observed in earlier studies in butadiene production workers in Portugal and the Czech Republic compared with controls (Sorsa et al., 1994, 1996b), positive results for chromosomal aberrations and sister chromatid exchanges were obtained in the most recent study of the Czech workers (Tates et al., 1996; Šrám et al., 1998). When genotype was considered, there was a significant increase in the frequency of chromosomal aberrations in both exposed and control subjects from both plants who were deficient for the GSTT1 gene (Sorsa et al., 1996a).

An increased frequency of hprt^– mutants in periph eral blood lymphocytes has been observed in two studies of exposed workers at a butadiene production facility in Texas, USA (Legator et al., 1993; Ward et al., 1994; Au et al., 1995) and in preliminary results of a study of styrene-butadiene rubber workers from the same region (J.B. Ward et al., 1996).¹³ Although analyses by genotype are not yet available, it was noted that the highest frequency of hprt^– variants occurred in an individual who was GSTT1 null. In contrast to the observations in the Texan plants, however, no increase in hprt^– mutant frequency was observed in workers exposed to similar levels of butadiene at the monomer plant in the Czech Republic (Tates et al., 1996) or in a population of polybutadiene rubber workers in China (Hayes et al., 1996) (no information on genotype was presented). These investigations involved different analytical methodologies (autoradiographic versus clonal assays), which may account for the discordance in the results; in addition, differences in occupational scenarios, exposure levels, age, smoking habits, or other lifestyle factors may have contributed to the discrepancy. Current ongoing research (including genotyping) may explain the differences in the results.

Decreased DNA repair ability was also observed in peripheral blood lymphocytes of exposed workers at the monomer production and styrene-butadiene rubber facilities in Texas in both a gamma-radiation challenge assay and a CAT-Host Cell Reactivation assay (Hallberg et al., 1997).¹⁴ However, the difference between exposed and "unexposed" monomer workers in the response to the challenge assay was no longer significant after ambient levels in the plant were reduced. Similarly, the effect on DNA repair ability in styrene-butadiene rubber workers was less when only non-smokers were considered. The detection of alkylated DNA (the same adduct as detected in the liver of mice and rats exposed to butadiene; Jelitto et al., 1989; Koivisto et al., 1997) in the urine of an exposed worker (Peltonen et al., 1993) also provides some evidence of the interaction of butadiene or its metabolites with genetic material in humans.

Although the metabolism of butadiene appears to be qualitatively similar across species, there are extensive data that indicate that the putatively active epoxide metabolites are formed to a greater degree in mice than in rats. Similarly, although in vivo data are limited, humans appear to metabolize butadiene to the mono- and diepoxide metabolites to a much lesser extent than mice. However, based on the observed variability in the formation of adducts of haemoglobin with butadiene metabolites in occupationally exposed human popula tions, there appears to be interindividual variation in humans, which is likely related to polymorphism for genes that code for enzymes involved in the metabolism of butadiene. The weight of evidence for the carcino genicity, genotoxicity, and non-neoplastic effects of butadiene needs to be considered, therefore, in the context of these interspecies and interindividual variations.

10.1.1 Carcinogenicity and genotoxicity

Data supporting the interspecies differences in production of active epoxide metabolites are in concordance with the observed difference in sensitivity between mice and rats (at least for the few strains investigated) to butadiene-induced carcinogenicity, in that the substance appears to be much more potent in mice than in rats. Although butadiene was a multi-site carcinogen in both mice and rats at all exposure levels tested (Hazleton Laboratories Europe Ltd., 1981a; NTP, 1984, 1993; Irons et al., 1989), the concentrations that induced tumours in the only study available in rats were much greater than those that were tumorigenic in mice (i.e., ³ 1000 ppm [³ 2212 mg/m³] versus ³ 6.25 ppm [³ 13.8 mg/m³]).

Species differences in sensitivity to genetic effects induced by butadiene have also been observed. Although butadiene was mutagenic in somatic cells of both mice and rats, its mutagenic potency was greater in mice. Other genotoxic end-points (chromosomal aberrations, sister chromatid exchanges, and micronuclei) were noted in somatic cells of mice but not in those of rats exposed to much higher concentrations. Butadiene was genotoxic in germ cells of male mice in multiple assays, while negative results were obtained in the single dominant lethal study in rats. Unlike the observations with the parent compound, however, there is little evidence that there are species differences in the sensitivity to geno toxic effects induced by the epoxide metabolites of butadiene (EB, DEB, and EBdiol), although there was some indication of interstrain variability. These data suggest that interspecies differences in sensitivity to butadiene-induced genotoxicity are related to quantitative differences in the formation of active metabolites.

There is also limited evidence of the genotoxicity of butadiene in exposed workers; although data are not completely consistent, increased frequencies of chromo somal aberrations, sister chromatid exchanges, and hprt^– mutations and decreased DNA repair capability have been reported in some studies of workers in the mono mer and/or styrene-butadiene rubber manufacturing industries (Legator et al., 1993; J.B. Ward et al., 1994, 1996; Au et al., 1995; Tates et al., 1996; Hallberg et al., 1997; Šrám et al., 1998).¹⁵ The discrepancy in the results may be due to the use of different methods for the detection of mutations or differences in exposure levels. In addition, since sensitivity to induction of genetic effects by butadiene and its metabolites has been linked to genotype for glutathione-S-transferase enzymes in several in vitro and a few in vivo studies, interpretation of the inconsistent observations in the available database is complicated by the lack of information on genotype for most of the small populations examined.

There have been several epidemiological investi gations of the carcinogenicity of butadiene that serve as a basis for assessment of the weight of evidence for causal ity based on traditional criteria. In the most recent cohort study (Delzell et al., 1995), which is also the largest and most comprehensive investigation conducted to date and that in which exposure was most extensively character ized, an association between exposure to butadiene in the styrene-butadiene rubber industry and leukaemia was observed (i.e., there was a quantifiable exposure– response relationship). SMRs for leukaemia were elevated for the overall cohort of workers from eight plants; the strength of this association was generally greater when specific subgroups with greater potential for exposure were considered. In addition, there was an increase in the RR for leukaemia with increased cumu lative exposure to butadiene in workers from the six plants for which exposure was best characterized. The association between leukaemia and exposure to buta diene remained when the potential role of two other substances present in the work environment (i.e., styrene and benzene) was considered. Although further refine ment of the estimates of exposure at one of these plants resulted in increases for several job categories (Macaluso et al., 1997), it is unlikely that these changes would affect the relative ranking of the categories and analyses in which exposed workers were compared with "non- exposed" workers (Gerin & Siemiatycki, 1998); therefore, these results are not inconsistent with the association observed by Delzell et al. (1995).

However, no increase in mortality due to leukae mia was observed in studies of workers involved in the production of butadiene monomer who were not con comitantly exposed to the other substances present in the styrene-butadiene rubber industry (E.M. Ward et al., 1995, 1996; Divine & Hartman, 1996). Although there was some evidence of increased mortality due to lymphosarcoma and reticulosarcoma in the subgroup of workers potentially exposed to the highest concentrations of butadiene in the largest of these investigations, there was no association with duration of employment or estimated cumulative exposure (based on qualitative ranking of potential for exposure). Although mortality due to lymphosarcoma was non-significantly elevated in some process groups in the styrene-butadiene rubber cohort (Delzell et al., 1995), there were no consistent patterns (other than for leukaemia), even when currently accepted terminology for lymphohaematopoietic cancers was used (Sathiakumar et al., 1998).

The traditional criterion of consistency for the observed association between exposure to butadiene and leukaemia is fulfilled, at least in part, in that similar excesses were observed among plants in the large cohort study of styrene-butadiene rubber workers (Delzell et al., 1995); i.e., there is internal consistency. A similar exposure–response was also noted in an independent nested case–control study of mostly the same population in which different exposure assessment methodology was employed (Matanoski et al., 1997). Observation of external consistency with results of other cohort studies of styrene-butadiene rubber workers is largely precluded, in view of the scope of the large epidemiological cohort study that included a large proportion of all of the styrene-butadiene rubber workers in North America. Indeed, it is difficult to envisage additional studies in this occupational group that would contribute meaningfully to weight of evidence for consistency of the observed association.

One criterion for causality of observed associations in epidemiological studies, namely coherence, may not have been adequately fulfilled, in view of the difference in the specific form of lymphohaematopoietic cancer in excess in available investigations for the two principal types of populations of workers studied. Indeed, increases in lymphosarcoma and reticulosarcoma have been observed in monomer production workers, whereas increases in leukaemia have been observed in styrene- butadiene rubber workers. Although it is plausible that this difference may be related to variation in the extent of information available for characterization of exposure or to the nature of exposures in the two industries, this has not been systematically investigated. There is also the possibility of misclassification of cause of death on death certificates (although Sathiakumar et al. [1998] did not observe an association with forms of lymphohaemato poietic cancer other than leukaemia in the large cohort of styrene-butadiene rubber workers when causes of death were examined using current terminology). The potential for transformation of one form of lymphohaematopoietic cancer to another (e.g., non-Hodgkin’s lymphoma to leukaemia) has also been noted (Sathiakumar et al., 1998). In addition, available data for the large study of styrene-butadiene rubber workers were insufficient to determine if butadiene was causally associated with a specific form of leukaemia. Moreover, it is noteworthy that these different tumours observed in styrene- butadiene rubber workers and monomer production workers are of the same organ system, and perhaps even share the same pluripotential stem cell.

An association between exposure to butadiene and the induction of leukaemia is also biologically plausible. The haematopoietic system is a target for butadiene- induced effects in rodents (i.e., lymphocytic lymphomas [NTP, 1993], cytogenetic effects in bone marrow [Cunningham et al., 1986; Irons et al., 1986a, 1987; Tice et al., 1987; NTP, 1993; Leavens et al., 1997], and suppression of stem cell differentiation [Irons et al., 1996]). Aneuploidy, which is believed to be associated with leukaemia in humans, has been induced in human lymphocytes exposed in vitro to the mono- and diepox ide metabolites of butadiene (Vlachodimitropoulos et al., 1997; Xi et al., 1997). Moreover, the presence of rele vant metabolizing enzymes in progenitor cells believed to be important targets for the induction of leukaemia in humans (i.e., CD34⁺ cells) has been demonstrated in studies of the metabolism of benzene (a documented human leukaemogen) (Schattenberg et al., 1994; Ross et al., 1996) (although exposure of human CD34⁺ cells to EB at "physiologically relevant concentrations" did not alter cytokine-induced clonogenic response, an early change frequently observed in the development of leukaemia; Irons et al., 1996).¹⁶ Therefore, available data also support the biological plausibility of an association between exposure to butadiene and leukaemia observed in humans, although the active metabolite has not been identified.

Therefore, although not completely convincing in their own right, the available epidemiological studies of the association between leukaemia and exposure to butadiene in occupationally exposed human populations fulfil several of the traditional criteria for causality, including strength of association (RR of 4.2 in the highest exposure group [based on five cases], which would be considered moderately strong), quantifiable exposure–response relationship, temporal relationship (the critical investigation [i.e., Delzell et al., 1995] is a historical cohort study), biological plausibility, and, to some degree, consistency, although the criterion for coherence is not fully satisfied.

Assessment of the weight of evidence for carcinogenicity in human populations should not, however, be considered in isolation from the extensive supporting data on carcinogenicity, genotoxicity, and inter- and intraspecies variations in metabolism and response. The association between exposure to butadiene and development of cancer is supported by limited evidence of genetic damage in exposed workers, as well as the wealth of evidence that butadiene is carcinogenic and/or genotoxic in all species of experimental animals tested (mice, rats, and hamsters), inducing a wide range of tumours and genetic damage at relatively low concentrations in mice (i.e., within the same order of magnitude as current occupational health limits). Moreover, while there are quantitative differences in the potency of the substance to induce tumours in various species, likely related to observed quantitative differences in metabolism, there are indications of considerable interindividual variations in the metabolism of butadiene in the human population, consistent with expectations for a complex metabolic pathway.

The observation of an association between exposure in the occupational environment and leukaemia that fulfils several of the traditional criteria for causality of associations observed in epidemiological studies, as well as supporting limited data on genotoxicity in human populations and the well documented carcinogenicity and genotoxicity at relatively low concentrations in some species of experimental animals, provides weight of evidence that butadiene is carcinogenic in humans.

Although relevant data in humans are limited, the results of in vivo studies in experimental animals indicate that butadiene induces mutations in somatic cells and male germ cells as well as male-mediated heritable clastogenic damage. While most of the studies have been conducted in mice, rats appear to be less sensitive to these effects, which is consistent with species differences in metabolism. However, in view of the likely consider able heterogeneity in the metabolism of butadiene in human populations, butadiene may be a human somatic and germ cell genotoxicant.

10.1.2 Non-neoplastic effects

The available data on effects of butadiene other than carcinogenicity or genotoxicity are limited. Based on the limited data available, species differences in the ability of butadiene to induce other non-neoplastic effects again appear to be consistent with variations in metabolism of butadiene to active metabolites. However, butadiene is of low acute toxicity in both rats and mice, in contrast to its ability to induce cancer and genetic damage at relatively low concentrations in mice.

Haematological effects suggestive of macrocytic anaemia have been consistently observed in mice (two strains) following short-term, subchronic, or chronic exposure to butadiene at concentrations similar to or lower than those that induced general toxicity (as indicated by decreased body weight gain and increased organ weights) (Irons et al., 1986a, 1986b; NTP, 1993; Bevan et al., 1996). For example, changes in haemato logical parameters were noted in mice exposed to ³ 62.5 ppm (³ 138 mg/m³) butadiene for 9 months or longer in the NTP bioassay. Butadiene also induced effects on bone marrow (including atrophy, decreased cellularity, regeneration, and alterations in stem cell development) in mice (Irons et al., 1986a, 1986b; Leiderman et al., 1986; NTP, 1993), although available data are inadequate to assess the potential effects on immune system function. While effects on the blood and bone marrow have not been reported in rats in recent investigations (including the only identified chronic bioassay; Hazleton Laboratories Europe Ltd., 1981a), the database is considerably more limited. In addition, the lack of observation of haematotoxicity in rats may again reflect the species differences in metabolism. Although the available epidemiological studies are too limited to assess the haematotoxicity in humans, available data support the haematopoietic system being a critical target for butadiene-induced toxicity, since the lymphohaematopoietic system is a target for butadiene- induced leukaemia in humans. However, it has not been established if the non-neoplastic effects observed in animals may be preliminary to, or associated with, the development of lymphohaematopoietic cancers.

The reproductive organs are also critical targets of butadiene-induced non-neoplastic effects in mice. Ovarian atrophy, the severity and incidence of which increased with concentration or duration of exposure, was observed at all concentrations (i.e., ³ 6.25 ppm [³ 13.8 mg/m³]) in the chronic bioassay conducted by the NTP (1993); in all exposure groups, the level of degen eration at 2 years, characterized by lack of oocytes, follicles, or corpora lutea, was incompatible with reproductive capacity. Although recent re-examination of some of the tissue samples indicated that the atrophy observed in the ovaries may be related to senile changes,¹⁷ it may be that butadiene is exacerbating these changes. It should be noted, though, that the incidence of these lesions was increased as early as 9 months (although the slides from these interim sacrifices have not been re-examined). That butadiene is causally associated with these lesions is also difficult to dismiss on the basis of currently available data, in view of the consistency with the results of other studies, including the earlier NTP (1984) bioassay and a subchronic study at higher concentrations (Bevan et al., 1996) in which such lesions were also observed, the presence of a clear dose–response relationship, and biological plausibility. Based on the observation of depletion of ovarian follicles and alkylation with ovarian macromolecules in mice following intraperitoneal administration of the monoepoxide or diepoxide metabolite and in rats administered the diepoxide (Doerr et al., 1995), it is possible that the ovarian toxicity is mediated through generation of the active epoxide metabolites.

Testicular atrophy was noted only in male mice exposed to concentrations greater than those that induced effects in females (NTP, 1993). Consistent with meta bolic differences, butadiene did not induce ovarian or testicular toxicity in the limited number of available studies in rats, although, as noted above, the diepoxide metabolite was ovotoxic in both species (Doerr et al., 1995, 1996).

Although available data are limited, there is no conclusive evidence that butadiene is teratogenic in mice or rats following maternal or paternal exposure or that it induces significant fetal toxicity at concentrations below those that are maternally toxic.

Available epidemiological data are inadequate for evaluation of potential reproductive or developmental toxicity; in fact, none of the identified analytical studies was conducted in women. However, in view of the qualitative similarities in the metabolism of butadiene in mice, rats, and humans and the likely variation across the gen eral population associated with genetic polymorphism for the relevant enzymes, and on the basis of the observed ovarian toxicity in butadiene-exposed mice, butadiene may be a reproductive toxicant in humans, although additional work to clarify the relevance of these observed effects is clearly desirable.

Available data on other systemic or organ-specific effects are inadequate to determine if such effects might be considered critical.

As per the approach adopted for several genotoxic carcinogens, measures of the potency of a substance to induce effects for which there is believed to be no threshold may be used to establish guideline values for environmental media. Since air is the principal route of exposure to butadiene in the general environment (available data indicate that other routes contribute negligibly), quantitation of exposure–response for cancer as well as non-cancer effects is limited to exposure by inhalation.

In order to eliminate the uncertainty associated with extrapolation from animal species, quantitative measures of carcinogenic potency (i.e., tumorigenic concentrations, or TCs)¹⁸ have been developed on the basis of available epidemiological data. This is based on the conclusion that the weight of evidence for an association between butadiene and leukaemia satisfies several of the traditional criteria for causality in epidemiological studies. However, uncertainties in the exposure estimates for the critical cohort of workers as well as confounding or effect-modifying aspects that could impact on quantitative estimates of risk are recognized. In view of these factors and to serve as a basis for comparison, quantitative measures of cancer potency have also been developed on the basis of results of long-term bioassays in rats and mice, with those in mice being considered justifiably conservative, considering the likely heterogeneity in metabolic transformation of butadiene in humans. (See discussion of relevance of specific tumour types in animals to humans in section 10.2.1.2.)

In addition to inducing tumours at multiple sites in experimental animals, butadiene is also genotoxic in somatic and germ cells and induces reproductive and haematological effects in animals. As a measure of exposure–response for non-cancer effects, where con sidered appropriate, benchmark concentrations¹⁹ have been calculated on the basis of data from long-term studies in mice.

Several physiologically based pharmacokinetic (PBPK) models have been developed as a basis for reducing uncertainty in interspecies extrapolations for butadiene by various groups of investigators. However, none of the models currently available has adequately accounted for the distribution of metabolites in the compartments included; the principal researchers in this field have concluded that there are likely more factors involved in butadiene metabolism than have been included in the models developed to date (Csanády et al., 1996; Sweeney et al., 1997). In addition, none of the models has included the formation of EBdiol, a puta tively active metabolite that is believed to be important in humans, since it has been observed to bind to haemo globin to a greater degree than EB in workers exposed to butadiene. Nor has bone marrow been incorporated as a compartment, although it appears to be a target site of butadiene-induced toxic effects. Moreover, none of the PBPK models has been validated in humans. For these reasons, therefore, such models have not been used to quantitatively account for interspecies variations in metabolism in the quantitation of exposure–response for critical end-points based on studies in experimental animals presented here.

In addition, owing to its relatively slow metabo lism, butadiene achieves a steady state during prolonged inhalation exposure. Exposures of the same concentra tion and duration would be expected to result in equiva lent toxicity across species, and interspecies scaling to account for variations in inhalation rate to body weight ratios or body surface areas between humans and animals is not considered necessary.

10.2.1 Carcinogenicity

In only one epidemiological investigation of the association between butadiene and leukaemia have data on exposure of the study population been sufficiently characterized to permit quantitation of exposure–response (Delzell et al., 1995). The Delzell et al. (1995) study also presents results for the largest cohort studied to date (including subjects from eight plants, six of which were included in the exposure–response analyses); it is also considered to subsume the observations of mortality in workers at these plants reported previously by other researchers (i.e., Meinhardt et al., 1982; Matanoski et al., 1990, 1993; Santos-Burgoa et al., 1992), because of the considerable overlap in the cohort definition. The exposure assessment of study subjects was of extremely high quality, being very thorough and based on research of plant records concerning work histories, processes and local emissions, and consultation with staff from each plant, and is, therefore, considered appropriate for quantification of exposure–response (limited industrial hygiene monitoring data were also available, although used primarily for comparison with estimated concentrations). For comparison with esti mates based on the data from the cohort study, carcino genic potency was also calculated on the basis of the results of the case–control study nested within essentially the same population of workers (Matanoski et al., 1997), although data available in the published report were too limited to permit detailed analysis here.

A detailed description of the methods employed and the assumptions made in the derivation of tumori genic potency estimates (i.e., TC₀₁, or the concentration associated with a 1% increase in mortality due to leukaemia based on the observations in Delzell et al. [1995]) is presented in Appendix 4. Although several mathematical models were applied, the choice of model had little impact on the resulting TC₀₁, as there was only a threefold variation in the range of values. However, the estimated TC₀₁ in which confidence is highest (i.e., that for which the model provided the best fit) is 1.7 mg/m³. The TC₀₁ calculated on the basis of the data of Matan oski et al. (1997) was only slightly lower than this value.

As described in section 10.1.1, butadiene induced an increase in the incidence of tumours at multiple sites in both B6C3F₁ mice (liver, lung, Harderian gland, mam mary gland, ovaries, forestomach, Zymbal gland, and kidney, along with malignant lymphomas, histiocytic sarcomas, and cardiac haemangiosarcomas) and Sprague-Dawley rats (mammary gland, thyroid gland, uterus, Zymbal gland, pancreas, and testes). As discussed above, consistent with the species differences in metabo lism, mice were much more sensitive to butadiene-induced cancer than were rats for the strains investigated. Based on data available (i.e., evidence from genotoxicity studies that butadiene and its metabolites are active in both species), this difference in sensitivity is quantitative rather than qualitative and is related to the greater amounts of putatively active metabolites formed in mice compared with rats. In addition, the different profiles of tumours observed in the two species may be related to differential roles of the epoxide metabolites in the induc tion of the various tumours; i.e., the diepoxide may be more critical to tumour induction in mice than is EB (since it was reported recently that formation of DEB increased with level of exposure to butadiene in mice but not in rats; Thornton-Manning et al., 1998), while the monoepoxide or monoepoxide diol may be more important in rats.

The relevance for extrapolation to humans of exposure–response for some of the types of tumours observed in rodents has been questioned. For example, Irons et al. (1989) hypothesized that the thymic lympho ma/leukaemia induced in B6C3F₁ mice may be related to the presence of an endogenous ecotropic retrovirus, as a much lower incidence was observed in Swiss mice that do not possess this retrovirus (although the incidence was significantly elevated compared with controls). Therefore, although the haematopoietic system is a target for the induction of cancer by butadiene in humans, the observed exposure–response relationship for this end- point is not considered appropriate for quantitative extrapolation to humans — on the basis that this retro virus is not present in humans and its presence in B6C3F₁ mice renders this strain quite susceptible to induction of lymphoma — although the relevant infor mation is included for comparative purposes.

It has also been suggested that the tumours observed in the study in rats (i.e., mammary gland, thyroid gland, pancreas, uterus, and testes) and some of the tumours induced in mice (i.e., ovaries and mammary gland) may be mediated through effects on the endocrine system. Indeed, tumours at these sites are often associ ated with disruption of hormonally mediated functions. In addition, non-neoplastic or pre-neoplastic effects, including atrophy, degeneration, and hyperplasia, have also been observed in mice exposed subchronically to butadiene. However, the mechanism by which butadiene induces tumours at these sites has not yet been ade quately investigated; i.e., it has not been established whether these tumours are induced via a mechanism for which there may be a threshold of exposure (e.g., through induction of hormonally mediated effects), although the possibility is recognized. In addition, the results of in vivo genotoxicity assays indicate that buta diene or its metabolites induce genetic effects in the reproductive organs of multiple strains of mice.

Based on these considerations, estimates of carcinogenic potency were calculated on the basis of the malignant lymphomas, histiocytic sarcomas, cardiac haemangiosarcomas, alveolar/bronchiolar adenomas or carcinomas, hepatocellular adenomas or carcinomas, squamous cell papillomas or carcinomas of the fore stomach, adenomas or carcinomas of the Harderian gland, granulosa cell tumours of the ovaries, and adenoacanthomas, carcinomas, or malignant mixed tumours of the mammary gland observed in B6C3F₁ mice in the chronic bioassay conducted by the NTP (1993) and the mammary gland tumours, pancreatic exocrine adenomas, Leydig cell tumours, Zymbal gland carcinomas, thyroid follicular cell adenomas or carcinomas, and uterine sarcomas in Sprague-Dawley rats reported by Hazleton Laboratories Europe Ltd. (1981a). It is noted that the characterization of exposure–response is much better in the study in mice (which involved five closely spaced exposure levels) than in the bioassay in rats (in which only two more widely spaced exposure levels were used, the higher of which was likely above the level of metabolic saturation). (Although there were also increased incidences of tumours at several sites in B6C3F₁ mice in the "stop-exposure" study conducted by the NTP [1993], only TC₀₅s determined on the basis of the 2-year study were included, as the latter study provides better information for characterization of exposure–response in mice following long-term exposure [i.e., more exposure levels for up to 2 years].)

The methods employed for development of estimates of tumorigenic potency (i.e., the concentrations associated with a 5% increased incidence of tumours, or TC₀₅s) based on these data are described in Appendix 4. TC₀₅s based on observations in mice ranged from 2.3 mg/m³ (95% lower confidence limit [LCL] = 1.7 mg/m³) for Harderian gland tumours in males to 99 mg/m³ (95% LCL = 23 mg/m³) for malignant lymphomas in males. For rats, calculated TC₀₅s ranged from 6.7 mg/m³ (95% LCL = 4.7 mg/m³) to 4872 mg/m³ (95% LCL = 766 mg/m³) for tumours of the mammary gland and Zymbal gland in females, respectively.

Although not presented here, modelling of the incidence of micronucleated polychromatic erythrocytes in B6C3F₁ mice exposed to butadiene for up to 15 months in the NTP bioassay resulted in a benchmark concentration (BMC₀₅) that was very similar to the lower end of the range of estimates of tumorigenic potency.

There have been recent attempts to quantitatively estimate risk of heritable genetic damage in humans based on a parallelogram approach and data on male-mediated heritable translocations and bone marrow micronuclei in mice and chromosomal aberrations in lymphocytes of exposed workers (Pacchierotti et al., 1998b). In view, however, of the reported ovarian atrophy due to reduction of primordial follicles (to a degree that would preclude reproduction) following chronic exposure of mice to concentrations of butadiene considerably lower than those associated with adverse effects on the testes, investigation of the response of female germ cells in mice to butadiene is desirable, since this may well be the most sensitive end-point for devel opment of quantitative estimates of heritable damage. (Determination of putatively toxic metabolites in the ovaries of butadiene-exposed female mice would also be informative.) For this reason, quantitation of exposure– response for heritable genetic damage is not presented here. However, in view of the apparent greater sensitivity of the reproductive organs in female mice, a benchmark concentration was derived for non-neoplastic effects in the ovary, which is considerably more protective than that for male-mediated heritable damage developed by Pacchierotti et al. (1998b). (Although the relative role of butadiene in the induction of the observed atrophy in mice in the NTP study is unclear, as discussed in section 10.1.2, information currently available is not considered a sufficient basis upon which to dismiss this end-point as being inappropriate for quantification of exposure– response. However, this uncertainty should be kept in mind in the interpretation or application of the BMC₀₅s derived below.)

Ovarian atrophy was observed in both long-term NTP (1984, 1993) bioassays in mice and a subchronic study (Bevan et al., 1996). Although limited, available data indicate that rats are less sensitive to induction of this effect, which may, again, be a consequence of inter species variations in metabolism. Therefore, although additional research into the etiology of the observed ovarian atrophy in mice would be desirable, the data from the later NTP study are considered most appropri ate for characterization of exposure–response (i.e., development of a BMC₀₅). In this investigation, the incidence of atrophy of the ovaries was significantly increased in an exposure-related manner at all concen trations tested (i.e., ³ 6.25 ppm [³ 13.8 mg/m³]). The severity of this effect also increased with exposure.

The derivation of the BMC₀₅ for ovarian atrophy is presented in detail in Appendix 4. Because the expo sure–response curve plateaus at the higher exposure levels, the two highest exposure groups were omitted from the calculations. The resulting BMC₀₅ for ovarian atrophy in mice was determined to be 0.57 mg/m³ (95% LCL = 0.44 mg/m³), when all degrees of severity were considered. If only lesions of moderate or marked severity were considered, the resulting BMC₀₅ would be about fivefold higher.

Haematotoxicity is considered to be a critical effect associated with exposure to butadiene. Although the haematopoietic system appears to be a target for butadiene-induced cancer in humans, available data on the potential non-neoplastic effects on this system are inadequate for quantitation of exposure–response. However, since statistically significant changes were observed in mice only at concentrations greater than those that induced other toxic effects, and since benchmark concentrations derived for effects on the blood are greater than those for these other effects, quantitation of the exposure–response for haematological effects has not been presented here.

10.3.1 Sample exposure characterization

The principal source of environmental exposure to butadiene is air. Although few data were identified regarding levels in drinking-water and food, due to its physical/chemical properties (e.g., vapour pressure and partition coefficients) and environmental release patterns (i.e., principally atmospheric emissions), intake of butadiene in these media is expected to be negligible in comparison with that in air.

As an example of population exposure characterization, estimates are presented on the basis of data available for Canada. Based on concentrations measured in outdoor air in several rural, suburban, and urban locations across Canada²⁰ (see section 6.1.1), 95% of the general population can be expected to be exposed to average concentrations of up to 1.0 µg/m³. However, since levels are generally greater in highly urbanized areas, estimated "reasonable worst-case exposure" is expected to be up to 1.3 µg/m³ (95th percentile). In areas influenced by industrial point sources, exposure could be as high as 6.4 µg/m³, based on the 95th percentile of concentrations measured near a source in Ontario (MOEE, 1995).

Individuals may also be exposed to butadiene for short durations while at self-service gasoline filling stations or in parking garages; however, these intakes are still much less than average daily intakes for the general population from inhalation of background concentrations in outdoor and indoor air.

Although available Canadian data indicate that butadiene is detected with greater frequency in indoor air than in outdoor air, there are insufficient data to characterize the distributions of concentrations of butadiene in various indoor environments. In general, butadiene is detected more frequently and at higher concentrations in indoor environments contaminated by ETS than in areas where smoking does not occur. Non- smokers who spend a considerable proportion of their time in indoor environments where ETS is present can be exposed to concentrations of butadiene that are an order of magnitude higher than the average levels in the outdoor air. Tobacco use (e.g., 20 cigarettes per day) can increase the daily intake of butadiene by smokers by five times over the daily intake by non-smokers in ETS- contaminated indoor locations. The daily intake of butadiene by smokers can be 100 times greater than the daily intake of non-smokers who are not exposed to ETS.

10.3.2 Sample risk characterization

Butadiene is released to air from both industrial point sources and more dispersive, non-point sources, the latter due to its production primarily during incomplete combustion. Based on estimates derived using monitoring data from Canada, intake for the general population is primarily from air, with intake from other media likely being negligible in comparison. The focus of the human health risk characterization is, therefore, the general population exposed in outdoor and indoor air in the general environment and those exposed through air in the vicinity of industrial point sources.

For compounds such as butadiene, where data are sufficient to support a plausible mode of action for induction of tumours by direct interaction with genetic material, estimates of exposure are compared with quantitative estimates of cancer potency to characterize risk.

Tumorigenic concentrations were calculated on the basis of data from both epidemiological studies and investigations in experimental animals. For the critical epidemiological investigation (Delzell et al., 1995), a TC₀₁ (i.e., the concentration associated with a 1% increase in mortality due to leukaemia) was considered the appropriate measure of carcinogenic potency, since the majority of the observable data fell within this range. Although four different mathematical models were considered, the TC₀₁ generated by the model with the best fit was 1.7 mg/m³.

Quantitative estimates of carcinogenic potency derived on the basis of data in experimental animals were calculated as TC₀₅s (i.e., the concentration associated with a 5% increase in tumour incidence). Based on the 2-year bioassay in mice (NTP, 1993), TC₀₅s ranged from 2.3 mg/m³ (95% LCL = 1.7 mg/m³) to 99 mg/m³ (95% LCL = 23 mg/m³). The TC₀₅s derived on the basis of the more limited study in rats (Hazleton Laboratories Europe Ltd., 1981a) ranged from 6.7 mg/m³ (95% LCL = 4.7 mg/m³) to 4872 mg/m³ (95% LCL = 766 mg/m³).

The values derived on the basis of studies in humans are preferred as the basis for comparison with estimates of exposure to characterize risk. While there are a number of uncertainties in the use of the epidemi ological data for both hazard evaluation and exposure– response analyses (section 10.4), these are likely far less than uncertainties associated with interspecies extrapo lation. Moreover, estimated potency for humans is similar to that developed on the basis of the cancer bioassays in experimental animals. (Indeed, although in an area of the exposure–response curve where data were more sparse, it is noteworthy that TC₀₅s calculated on the basis of epidemiological data [as opposed to the TC₀₁s presented above] are within the range of values derived from the studies in rodents.)

Based on the sample exposure scenarios presented above (section 10.3.1), 95% of the population in Canada is exposed to concentrations of butadiene in outdoor air of 1.0 µg/m³ or less. For the proportion of the general population that is regularly exposed to higher concentra tions of butadiene in urban areas (i.e., the "reasonable worst-case scenario"), the 95th percentile of the distrib ution of concentrations is 1.3 µg/m³. In the only area of Canada identified as having an industrial point source, the 95th percentile of the distribution of concentrations is 6.4 µg/m³.

The margins between carcinogenic potency and estimated exposure for the general population (including ambient and reasonable worst case) and those in the vicinity of a point source are presented in Table 6. Equivalent low-dose risk estimates are also presented in this table.

In view of the relative potency of butadiene to induce some non-cancer effects, these end-points are also important in risk characterization. As presented above, a benchmark concentration (BMC₀₅) of 0.57 mg/m³ (95% LCL = 0.44 mg/m³) was derived on the basis of data for the incidence of ovarian atrophy of all severities (i.e., female reproductive toxicity) in mice exposed to butadiene for up to 2 years (NTP, 1993). And while there is uncertainty about the relevance of the ovarian atrophy observed in mice for humans (section 10.4), the BMC₀₅ is slightly less than the lower end of the range of estimates of cancer potency based on the incidence of tumours in the same study in mice, as well as the TC₀₅ for cancer based on the epidemiological data. The mode of induction of ovarian atrophy is unknown. However, if it is (reasonably) assumed that the mode of action is related to that by which tumours are induced (i.e., direct interaction with genetic material), risk to human health for reproductive effects may be characterized in the same manner as presented for cancer. Therefore, estimates of the margin between the BMC₀₅ for ovarian toxicity and a sample exposure characterization are presented in Table 7. It should be noted, though, that even if the mode of induction of ovarian atrophy does not involve direct interaction with genetic material, the margin between exposure and effect level (i.e., for which a tolerable concentration is normally developed) is still small — i.e., exposure levels in Canada are 90–570 times lower than the benchmark concentration, as presented in Table 7.

There is a high degree of certainty that butadiene is being released to ambient air in Canada in significant amounts in vehicular exhaust. There is a moderate degree of certainty that exhaust emissions of butadiene are lower in well maintained vehicles equipped with catalytic converters than in older non-equipped vehicles, and that evaporative emissions during refuelling and vehicle operation contribute less to concentrations of butadiene in ambient air than do emissions in vehicular exhaust.

There is a moderate degree of certainty that butadiene is not being released to the Canadian environment in significant amounts from industrial activities in Canada, as only a single major point source (i.e., in Sarnia, Ontario) of discharge to the atmosphere has been identified. Although there is some uncertainty that the available measurements of butadiene in samples taken over a few days in the vicinity of this source are repre sentative of population exposure over the long term, since the samples were taken at distances of up to a few kilometres from the source, there is a moderate degree of certainty that a segment of the population would be exposed to the measured concentrations. There is a high degree of certainty that populations in rural areas are exposed to lower concentrations of butadiene in ambient air than are communities in more densely populated areas.

Available data on concentrations of butadiene in ambient air in Canada are quite extensive. A large pro portion of the numerous samples from several sampling sites across the country contained concentrations of butadiene above the level of detection. Therefore, there is a high degree of certainty in the estimations of expo sure to butadiene via ambient air.

The most limiting aspect of the exposure assess ment is the lack of sufficient data on the concentrations of butadiene in indoor air. This is an important short coming, since humans spend significantly more time in indoor environments than outdoors. In the absence of indoor sources, it is reasonably certain that concentra tions of butadiene in indoor environments are similar to the concentrations in the local ambient air.

Higher concentrations of butadiene have been measured in indoor air where ETS was known to be present. However, the data on concentrations of buta diene in ETS-contaminated indoor air are highly variable and are not sufficient to reasonably define the range of mean concentrations. Nevertheless, there is a high degree of certainty that non-smokers spending a considerable proportion of their time in indoor environments where ETS is present are exposed to higher concentrations of butadiene than are non-smokers who are not exposed to ETS. There is a high degree of certainty that smokers are exposed to higher concentrations of butadiene and have significantly higher daily intakes than do non-smokers.

There is somewhat less certainty that butadiene monomer is not released in detectable amounts from consumer products (e.g., synthetic materials) incorpor ating this compound in their production. Although there may be contributions to indoor concentrations of buta diene from certain cooking activities, the data are not sufficient to identify specific sources or activities or to identify a range of emissions of butadiene during cooking.

Although data on levels of butadiene in foodstuffs are scarce, based on the physical and chemical properties of the substance and the fact that it is released primarily to ambient air (where it is likely to remain without partitioning to other media), there is a reasonable degree of certainty that food does not represent a major source of exposure. Similarly, although the database for con centrations of butadiene in drinking-water is limited, there is a reasonable degree of certainty that drinking- water is not an important source of exposure for the general public in Canada, based on the volatility and release patterns of the compound.

There is some degree of uncertainty that the weight of epidemiological evidence for the association between butadiene and leukaemia satisfies criteria for causality. In particular, the need for coherence is seemingly not addressed, since the observed increase in mortality due to leukaemia in styrene-butadiene rubber workers was not observed in the cohorts of monomer workers (although there was some evidence of an association with other forms of lymphohaematopoietic cancer, particularly in short-term workers). This may be related to the nature of exposure to both butadiene and other substances in these two industries. However, in view of the overwhelming evidence of carcinogenicity and genotoxicity in experimental animals, available infor mation on species differences in sensitivity likely being related to differences in metabolism, and the potential for considerable interindividual variability in metabolism to putatively toxic metabolites in the human population, along with the limited evidence of genotoxicity in occu pationally exposed populations, there is a high degree of confidence that butadiene is likely to be carcinogenic in humans. Based on the extensive database on the geno toxicity of butadiene and its principal metabolites both in vitro and in vivo in both somatic and germ cells, confi dence that butadiene induces tumours (and possibly other effects) through direct interaction with genetic material is high.

Although the assessment of the exposure of the critical cohort of workers is likely one of the most comprehensive published to date, there is also uncer tainty in the estimates of carcinogenic potency derived on the basis of this study, due primarily to the fact that the estimates of exposure are based on few actual histori cal monitoring data.²¹ For example, when the exposure of workers at one plant was re-examined, there were 2- to 3-fold changes in the estimates for several job groups (with a 10-fold increase for one job group). In addition, with the exception of incorporating exposure to styrene as a stratification variable in the analyses, potential interactions between various occupational exposures could not be taken into account in the derivation of the carcinogenic potency based on the observations in this cohort. It has also been demonstrated that genetic poly morphism for several of the enzymes involved in metab olism of butadiene affects sensitivity to toxic effects induced by the substance. Also, since information on genotype for the relevant enzymes was not available for this large cohort and only a small amount of information on the distribution in the general population has been identified, it is not possible to determine how represen tative the study cohort is of the genetic susceptibility to butadiene of the general public.

With respect to the quantitation of exposure– response and derivation of potency estimates based on the epidemiological data, the inability of any of the models to consistently predict leukaemia rates in the validation study contributes to additional uncertainty. In addition, the small number of leukaemia cases being modelled contributes to model instability. However, the fact that the range of potency estimates for the four models is narrow (i.e., 1.4–4.3 mg/m³) increases the confidence in the calculated potencies. There is also some uncertainty associated with the fact that the potency estimates were based only on cases in which leukaemia was considered the underlying cause of death, rather than all cases, which could result in an underesti mation of leukaemogenic potency.

In view of the likely variability in metabolism of butadiene across the human population related to genetic polymorphism for relevant enzymes, estimates of car cinogenic potency as well as benchmark concentrations for non-cancer effects based on studies in mice are con sidered justifiably conservative. However, because of the high mortality in the study in mice in which exposure– response could best be characterized and the limitations in the study in rats (high mortality at the higher of only two widely spaced exposure levels), there is a moderate degree of uncertainty in estimates of carcinogenic potency derived on the basis of investigations in exper imental animals. In addition, since the available PBPK models were considered inadequate, the potency esti mates developed here are based on an inhaled concentra tions exposure metric; i.e., none of the available infor mation on species differences in metabolism is taken into account. It is noteworthy that if the calculated margins between exposure and carcinogenic potency presented above that serve to characterize risk were derived on the basis of the 95% LCLs of the TC₀₅s for tumours in mice, the values would differ by only 1.4- to 3.3-fold (i.e., within the same order of magnitude) from those calcu lated on the basis of the point estimates; similarly, use of the 95% LCLs of the TC₀₅s for tumours in rats would result in a 1.1- to 6.4-fold difference in the margins between exposure and potency. In addition, although confidence in the use of the potency estimate for lymphomas in mice is low, due to the inherent sensitivity associated with the presence of an endogenous retro virus, it is noteworthy that the TC₀₅ for this tumour would not be limiting, as it falls within the range of values determined for cancers at other sites. Also, it should be noted that, although these margins and measures of risk presented above were based on com parison of the 95th percentile of the exposure data for each scenario, use of the median concentration (i.e., the 50th percentile) and either the point estimates of carcinogenicity or the associated 95% LCLs would result in a 5-fold difference in the resulting values for the general population and a 10-fold difference in values for those in an area influenced by a point source.

There is uncertainty about the relevance of the ovarian atrophy observed in mice to humans, based on lack of data on the relative role of butadiene in the etiology of these lesions. Although the observed effects in the 2-year bioassay may have been related to senile changes, possibly exacerbated by butadiene, atrophy of the ovaries was detected in these mice as early as 9 months, and there is consistent evidence in other chronic and subchronic studies that the ovaries are tar gets of toxic effects induced by butadiene or its epoxide metabolites. As a result of this uncertainty, quantitative measures of dose–response developed on the basis of ovarian atrophy must necessarily be interpreted with caution. In addition, the BMC₀₅ presented above was based on inclusion of atrophy of all severities, including "minimal" severity, the biological significance of which is unclear. If only lesions of moderate or marked severity are considered, the resulting BMC₀₅ and hence the cal culated margin between exposure and effect level and risk estimates would differ by about fivefold. (Use of the 95% LCLs of the BMC₀₅s for atrophy of all severities or of only moderate or marked severity would result in only a 1.5- or 3-fold difference in the measure of risk.) How ever, in view of the weight of evidence of causality for the association between butadiene and these effects in mice and the relatively low value for the measure of dose–response compared with that for other types of effects, additional investigation in this area is deemed to be of high priority.

An IARC Working Group that convened in 1998 has classified 1,3-butadiene as probably carcinogenic to humans (Group 2A) based on limited evidence of carcinogenicity in humans and sufficient evidence of carcinogenicity for 1,3-butadiene and 1,2:3,4-diepoxy butane in experimental animals (IARC, 1999).

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Copies of the Canadian Environmental Protection Act Priority Substances List Assessment Report on 1,3-butadiene (Government of Canada, 2000) are available upon request from:

Inquiry Centre

Environment Canada

Main Floor, Place Vincent Massey

351 St. Joseph Blvd.

Hull, Quebec

Canada K1A 0H3

or on the Internet at:

www.ec.gc.ca/cceb1/eng/public/index_e.html

Unpublished supporting documentation for the health effects assessment, which presents additional information, is available upon request from:

Environmental Health Centre

Room 104

Health Canada

Tunney’s Pasture

Ottawa, Ontario

Canada K1A 0L2

Sections of the Assessment Report and supporting documentation on genotoxicity and reproductive and developmental toxicity were reviewed by D. Blakey and W. Foster, respectively, of the Environmental and Occupational Toxicology Division of Health Canada. A review of the exposure assessment included in the critical epidemiological studies was prepared under contract by M. Gerin and J. Siemiatycki of the Institut Armand-Frappier, University of Quebec.

In the first stage of external review, sections of the supporting documentation pertaining to human health were considered by the following individuals, primarily to address adequacy of coverage: J. Aquavella, Monsanto Company; M. Bird, Exxon Biomedical Sciences, Inc.; J.A. Bond, Chemical Industry Institute of Toxicology; I. Brooke, United Kingdom Health and Safety Executive; G. Granville, Shell Canada Ltd.; R. Keefe, Imperial Oil Ltd.; A. Koppikar, US Environmental Protection Agency; R.J. Lewis, Exxon Biomedical Sciences, Inc.; K. Peltonen, Finnish Institute of Occupational Health; and F. Ratpan, Nova Chemicals

In the second stage of external review, accuracy of reporting, adequacy of coverage, and defensibility of conclusions with respect to hazard characterization and exposure–response analyses were considered in written review by BIBRA International and the following individuals: R.J. Albertini, University of Vermont; J.A. Bond, Chemical Industry Institute of Toxicology; I. Brooke, United Kingdom Health and Safety Executive; J. Bucher, US National Toxicology Program; B. Davis, US National Toxicology Program; E. Delzell, University of Alabama at Birmingham; B.J. Divine, Texaco; A.A. Elfarra, University of Wisconsin-Madison; E. Frome, Oak Ridge National Laboratory; B.D. Goldstein, Environmental and Occupational Health Sciences Institute; R.F. Henderson, Lovelace Respiratory Research Institute; R.D. Irons, University of Colorado Health Sciences Center; A. Koppikar, US Environmental Protection Agency; J. Lubin, US National Cancer Institute; J. Lynch, Exxon Biomedical Sciences, Inc. (retired); R.L. Melnick, US National Toxicology Program; K. Peltonen, Finnish Institute of Occupational Health; A.G. Renwick, University of Southampton; J. Siemiatycki, Institut Armand-Frappier; L.T. Stayner, US National Institute for Occupational Safety and Health; J.A. Swenberg, University of North Carolina; R. Tice, Integrated Laboratory Systems, Inc.; and J.B. Ward, Jr., University of Texas Medical Branch.

In the third and final stage of external expert review, adequacy of incorporation of the comments received during the second stage was considered at a final meeting of a panel of the following members convened by Toxicology Excellence in Risk Assessment (TERA) in November 1998: H. Clewell, K.S. Crump Division of ICF Kaiser; M.L. Dourson, TERA; and L. Erdreich, Bailey Research Associates, Inc.

The health-related sections of the Assessment Report were reviewed and approved by the Health Protection Branch Risk Management meeting. The entire Assessment Report was reviewed and approved by the Environment Canada/Health Canada CEPA Management Committee.

Concurrent with review of the draft CICAD, there was also a public comment period for the source national assessment, in which the Priority Substances List Assessment Report was made available for 60 days (2 October to 1 December 1999). A summary of the comments and responses is available on the Internet at www.ec.gc.cceb1/eng/public/index_e.html.

The draft CICAD on 1,3-butadiene was sent for review to institutions and organizations identified by IPCS after contact with IPCS national Contact Points and Participating Institutions, as well as to identified experts. Comments were received from:

M. Baril, International Programme on Chemical Safety/ Institut de Recherche en Santé et en Sécurité du Travail du Québec, Canada

R. Benson, Drinking Water Program, US Environmental Protection Agency, USA

T. Berzins, National Chemicals Inspectorate (KEMI), Sweden

R. Cary, Health and Safety Executive, United Kingdom

R. Chhabra, National Institute of Environmental Health Sciences, National Institutes of Health, USA

P. Edwards, Department of Health, United Kingdom

H. Gibb, National Center for Environmental Assessment, US Environmental Protection Agency, USA

R. Hertel, Federal Institute for Health Protection of Consumers and Veterinary Medicine, Germany

J. Heuer, Federal Institute for Health Protection of Consumers and Veterinary Medicine, Germany

J. Jinot, US Environmental Protection Agency, USA

C. Kimmel, US Environmental Protection Agency, USA

A.M. Koppikar, US Environmental Protection Agency, USA

S. Kristensen, National Industrial Chemicals Notification and Assessment Scheme (NICNAS), Australia

N. Moore, BP Amoco Chemicals (commented through Department of Health, United Kingdom)

H. Nagy, National Institute of Occupational Safety and Health, USA

S. Tarkowski, Nofer Institute of Occupational Medicine, Poland

L. Vodickova, National Institute of Public Health, Centre of Industrial Hygiene and Occupational Diseases, Czech Republic

P. Yao, Institute of Occupational Medicine, Chinese Academy of Preventive Medicine, People’s Republic of China

K. Ziegler-Skylakakis, GSF-Forschungszentrum für Umvelt und Gesundheit, Germany (transmitted comments from BUA and industry representatives)

Mr H. Ahlers, Education and Information Division, National Institute for Occupational Safety and Health, Cincinnati, OH, USA

Dr T. Berzins, National Chemicals Inspectorate (KEMI), Solna, Sweden

Dr R.M. Bruce, Office of Research and Development, National Center for Environmental Assessment, US Environmental Protection Agency, Cincinnati, OH, USA

Mr R. Cary, Health and Safety Executive, Liverpool, United Kingdom (Rapporteur)

Dr R.S. Chhabra, General Toxicology Group, National Institute of Environmental Health Sciences, Research Triangle Park, NC, USA

Dr H. Choudhury, National Center for Environmental Assessment, US Environmental Protection Agency, Cincinnati, OH, USA

Dr S. Dobson, Centre for Ecology and Hydrology, Monks Wood, Abbots Ripton, United Kingdom (Chairman)

Dr H. Gibb, National Center for Environmental Assessment, US Environmental Protection Agency, Washington, DC, USA

Dr R.F. Hertel, Federal Institute for Health Protection of Consumers and Veterinary Medicine, Berlin, Germany

Ms K. Hughes, Priority Substances Section, Environmental Health Directorate, Health Canada, Ottawa, Ontario, Canada

Dr G. Koennecker, Chemical Risk Assessment, Fraunhofer Institute for Toxicology and Aerosol Research, Hanover, Germany

Ms M. Meek, Existing Substances Division, Environmental Health Directorate, Health Canada, Ottawa, Ontario, Canada

Dr A. Nishikawa, Division of Pathology, Biological Safety Research Centre, National Institute of Health Sciences, Tokyo, Japan

Dr V. Riihimäki, Finnish Institute of Occupational Health, Helsinki, Finland

Dr J. Risher, Agency for Toxic Substances and Disease Registry, Division of Toxicology, US Department of Health and Human Services, Atlanta, GA, USA

Professor K. Savolainen, Finnish Institute of Occupational Health, Helsinki, Finland (Vice-Chairman)

Dr J. Sekizawa, Division of Chem-Bio Informatics, National Institute of Health Sciences, Tokyo, Japan

Dr S. Soliman, Department of Pesticide Chemistry, Faculty of Agriculture, Alexandria University, Alexandria, Egypt

Ms D. Willcocks, National Industrial Chemicals Notification and Assessment Scheme, Sydney, NSW, Australia

Dr R.J. Lewis (representative of European Centre for Ecotoxicology and Toxicology of Chemicals), Epidemiology and Health Surveillance, ExxonMobil Biomedical Sciences, Inc., Annandale, NJ, USA

Dr A. Aitio, International Programme on Chemical Safety, World Health Organization, Geneva, Switzerland (Secretary)

Dr P.G. Jenkins, International Programme on Chemical Safety, World Health Organization, Geneva, Switzerland

Dr M. Younes, International Programme on Chemical Safety, World Health Organization, Geneva, Switzerland

The raw study data²² for the six plants investigated by Delzell et al. (1995) were used to calculate the potency estimates. The data consisted of the cumulative occupational exposures to butadiene and styrene at each year of each subject’s life (person-year), beginning with his entry into the cohort and terminating with death or other exit from the cohort. The data also contained information on race, age, calendar year, and years since hire of each subject.

The response of interest was cases of death due to all forms of leukaemia, as information on the specific type of leukaemia was insufficient; only cases in which leukaemia was considered the underlying cause of death were considered in these analyses. Exposure estimates were cumulative occupational exposures in ppm-years assumed to be incurred for 8 h/day, 240 days/year over a 45-year working life.

The objective of this exposure–response analysis was to compute the carcinogenic potency, expressed as the TC₀₁, or the concentration of butadiene associated with a 1% excess probability of dying from leukaemia. This analysis involved two stages. First, the relationship between exposure and the death rate due to leukaemia within the cohort was modelled. This was accomplished by collapsing (or stratifying) the data into discrete exposure cate gories and then modelling the mean exposure in each category versus the death rates due to leukaemia. In the second stage of analysis, the TC₀₁ was calculated based on this exposure–response relationship and the background mortality rates in the Canadian population.

In addition to stratifying by exposure, the data were stratified by race, age, calendar year, years since hire, and styrene exposure in order to incorporate this information into the exposure–response relationship. Each of these variables was collapsed into a small number of discrete categories in order to reduce the number of strata, thereby improving model stability. These variables and their categories are presented in Table A-1. Exposure, defined as the mean cumulative exposure per person-year, was calculated for person-years falling into each possible combination of the stratification variables.

Table A-1: Stratification variables for exposure–response modelling of epidemiological data from Delzell et al. (1995).

The data were imported to Epicure (1993)²³ for exposure– response modelling. All fitted models were of the form:

RR = O/E = g(D(t))

where RR is the rate ratio, O and E are the observed and expected numbers of leukaemia deaths, D(t) is cumulative butadiene expo sure up to time t, and g is the exposure–response model, which is constrained to pass through one at zero exposure. Four different models, discussed in more detail below, were fitted to the data. At the model-fitting stage, the expected number of deaths is calculated on the basis of the non-exposed person-years in the cohort, and not from population background rates.

Once the fitted exposure–response model was obtained, the lifetime probability of death due to leukaemia was computed using lifetable methods taking into account the death rates in the Cana dian population. The derivation of the formula used for the lifetime probability of death due to leukaemia proceeds as follows.

Let d(t) represent the exposure concentration of butadiene in ppm at age t years, and let D(t) denote the cumulative exposure in ppm-years, with:

This formulation of cumulative exposure allows for the possibility of non-constant exposure scenarios.

At a cumulative exposure of D(t) ppm-years, the probability of dying from leukaemia by age t is given by:

where h_R(D(t);t) is the mortality rate from leukaemia at age t given a cumulative exposure to butadiene of D(t), and S(t) is the probability of survival up to age t. Equation (1) follows from the argument that the probability of death by age t is equal to the probability of death at age t multiplied by the probability of surviving up until age t. In lifetable analysis, the mortality and survival rates are constant for each year, so the integral in (1) can be replaced by a summation over year.

Exposure to butadiene is assumed to augment the back ground rate of leukaemia in a multiplicative fashion. In other words, the mortality rate, given exposure to butadiene, is equal to the background exposure rate multiplied by the excess risk due to exposure to butadiene. This is known as the "proportional hazard" model and is expressed as:

where h(t) is the background mortality rate from leukaemia in the population, calculated from Canadian age-specific death rates²⁴ due to leukaemia, and g(D(t)) is the fitted exposure–response model, or excess risk at age t.

The survival rate, S(t), appearing in equation (1) is computed from Canadian age-specific death rates due to all causes, where the reported Canadian leukaemia mortality rate is replaced by the modelled rate in order to incorporate exposure to butadiene. The formula describing the probability of survival up to age i is given by:

where h_j^* and h_j are the Canadian mortality rates due to all causes and due to leukaemia at age j, respectively, and g_j = g(D(j)) is the excess risk at age j.

Substituting equation (2) into (1), the lifetime probability of death due to leukaemia is given by:

where 1–70 years is the standard lifetime for a human.

The TC₀₁ is computed by determining the exposure D(t) at which the excess risk is equal to 0.01. That is,

If a constant exposure d is assumed for an individual from birth to age 70 years, then d(t) = d ppm and the cumulative exposure D(t) = d × t ppm-years. The TC₀₁ is then the ambient exposure level d (in ppm) at which the excess risk equals 0.01 at t = 70 years.

In separate analyses, exposures were lagged by n = 2, 5, 10, 15, 20, and 25 years to determine if the models would provide better fits if the most recent n years of exposure were ignored. An n-year lag was achieved by resetting an individual’s cumulative exposure at each year to be equal to the exposure he had accumulated n years prior. In so doing, the last n years of exposure do not affect the probability of developing leukaemia. The data were first stratified on unlagged cumulative exposure, and then the individual exposures were lagged. Thus, the number of strata remains constant when using different lag periods, and models with different lags may be directly compared (Preston et al., 1987).

To assess the predictive power of the exposure–response models, a validation study was performed in which individuals in the cohort were divided randomly into two groups. The models were fit separately to both groups, and then a likelihood ratio test was performed to determine if the estimated parameters were equal. The process of dividing and fitting was repeated 1000 times to characterize the variability due to the random splitting process. If the models provided consistent fits, then the likelihood ratio test would be expected to reject at a rate equal to the desired signif icance level of the test (i.e., at a significance level of 0.05, the fitted parameters should be significantly different 1 in 20 times). If the tests are significant more often than this, the confidence in the predictive power of the models is reduced.

Four different exposure–response models were examined and are presented in Table A-2. These models are identical to those fitted in the Delzell et al. (1995) report except that model 2 is more general and flexible than the square root model used by those authors. Preliminary analysis indicated that all stratification vari ables except race significantly affected the model fit. Since race was only marginally insignificant, all variables were used to stratify the data prior to model fitting.

The four models were fitted while stratifying on race, age, calendar year, years since hire, and styrene exposure. The results of the model fitting are displayed in Table A-2. (Note: A smaller deviance roughly indicates a better fit.) A graphic representation of the data and the fitted models is shown in Figure A-1. Judging from the model deviances and the shape of the curves relative to the data, especially in the low-dose region, model 1 provides the best fit to the data.

For purposes of comparison, the same models were fitted using the median exposure as per the Delzell et al. (1995) report. These analyses indicated that there is little difference between using mean or median exposures. Models including age as a multiplying factor of e^gamma•age instead of as a stratification variable were also fitted, but these models did not fit as well. Since cumulative exposure and years since hire may be confounded, their interaction was examined. The interaction was not significant for any of the models. The same models were refitted excluding the largest exposure group (200+ ppm-years), but this did not significantly affect any of the parameter estimates. The four models were also refitted allowing for different background rates for control and exposed populations. Different background rates might be neces sary in occupational studies where lifetime non-exposed workers may differ fundamentally from exposed workers as a result of differences in jobs and work areas. Results of this analysis indicated that different background rates are not necessary for these data.

The parameter estimates obtained in the present analysis are also not significantly different from those presented in the Delzell et al. (1995) report. The differences in parameter estimates are likely due to the different levels used in the stratification vari ables. Table A-2 compares the parameter estimates obtained in this analysis with those of the Delzell et al. (1995) report.

The TC₀₁s were calculated for each model using the lifetable methods described above, and the resulting ambient occupational exposures per person-year were converted to environmental exposures by assuming that the exposures occurred for 8 h/day,

Table A-2: Parameter estimates and model deviances for each of four models fit to mean cumulative exposure per person-year for Delzell et al. (1995) study and comparison with parameter estimates from Delzell et al. (1995) analyses.

a p-value of likelihood ratio test of equality of parameters.

b For the Delzell et al. (1995) analysis, alpha was fixed at 0.5, and only beta was estimated.

240 days/year. This amounts to multiplying the TC₀₁ by:

To convert the ambient exposures from ppm to mg/m³, the TC₀₁s are further multiplied by 2.21, the conversion factor for butadiene. The occupational and equivalent environmental TC₀₁s are presented in Table A-3. Environmental TC₀₁s for each of the four models ranged from 1.4 to 4.3 mg/m³. TC₀₁s calculated excluding the largest exposure group were slightly smaller, ranging from 0.6 to 1.6 mg/m³, while those calculated on the basis of median exposures were similar, ranging from 0.4 to 5.0 mg/m³.

TC₀₁s were also calculated using the parameter estimates from the Delzell et al. (1995) report and are compared with the TC₀₁s developed here in Table A-3. They ranged from 3.1 to 14.3 mg/m³.

Lagged exposure analysis

The same four models were fitted when exposures were lagged by 2, 5, 10, 15, 20, and 25 years. The resulting model fits are displayed in Table A-4. Since the deviances are similar for each lag period, this analysis indicates that lagging exposures does not dramatically improve the fit of any of the four models. In fact, TC₀₁s for all four models and all lag periods ranged from 0.8 to 4.3 mg/m³.

Validation study

With respect to model validation, the p-values for the tests of equality of the parameters are displayed in Table A-5. If the models were providing consistent fits between the two halves, the propor tion of p-values less than the significance level of alpha would be alpha. The results of the simulation study indicate that the test is rejecting more often than would be expected if the models were providing the same fits to both halves of the data. For model 1, the test was rejected at a significance level of 1% in 7.4% of the runs, whereas a rejection rate of 1% of the runs would be expected if the model was fitting consistently. The results of this analysis reduce the confidence in the power of the models to predict leukaemia mortality rates.

Table A-3: Carcinogenic potency estimates (TC₀₁s) for models fit to mean cumulative exposure per person-year based on Delzell et al. (1995) study and comparison with estimates from Delzell et al. (1995) analyses.

Model	Current analysis		Delzell et al. (1995) analysis
	Occupational TC01 (mg/m3)	Environmental TC01 (mg/m3)	Environmental TC01 (mg/m3)
1) RR = (1 + dose)	7.8	1.7	14.3
2) RR = 1 + beta•dose	6.5	1.4	6.4
3) RR = ebeta•dose	19.8	4.3	3.1
4) RR = 1 + beta•dose	13.8	3.0	4.5

Summary

It is noteworthy that the choice of the exposure–response model does not have a large impact on the resulting TC₀₁; as indicated in Table A-3, the values are similar, ranging from 1.4 to 4.3 mg/m³. However, if a best model must be chosen, it would be

model 1, owing to the smaller deviance (Table A-2), the shape of the curve relative to the data in the low-dose region (Figure A-1), and the fact that it has one fewer parameter than model 2, which provides a similar fit. The TC₀₁ for model 1 is 1.7 mg/m³.

It is difficult, though, to assess how well any of these models truly describes the data. It is noted that the plot in Figure A-1 provides only a rough indication of the shape of the data, since each point on the plot is an average of data in many strata. The results of the validation study reduce confidence in the ability of the models to predict leukaemia mortality.

The choice of exposure lag does not greatly improve the fit of any of the four models, and it does not affect the resulting TC₀₁. Including all lagged models, the range of TC₀₁s is still from 0.8 to 4.3 mg/m³.

For comparison with these values, potency estimates were also calculated on the basis of the recent case–control study of styrene-butadiene rubber workers by Matanoski et al. (1997). Although workers were from plants subsumed in the Delzell et al. (1995) study, exposure was independently characterized. Treating the odds ratio presented by these authors as a rate ratio (since leukaemia is a rare disease) and using their model and parameter estimates as well as the same lifetable methods described above, the TC₀₁ for environmental exposure was calculated to be 0.4 mg/m³. It is reassuring, therefore, that this value is only slightly lower than the estimates derived on the basis of the Delzell et al. (1995) cohort study data.

Tumorigenic concentrations based on data from studies in experimental animals

Estimates of carcinogenic potency were calculated on the basis of the incidences of malignant lymphomas, histiocytic sarcomas, cardiac haemangiosarcomas, alveolar/bronchiolar adenomas or carcinomas, hepatocellular adenomas or carcinomas, squamous cell papillomas or carcinomas of the forestomach, adenomas or carcinomas of the Harderian gland, granulosa cell tumours of the ovaries, and adenoacanthomas, carcinomas, or malignant mixed tumours of the mammary gland observed in B6C3F₁ mice in the chronic bioassay conducted by the NTP (1993) and the mammary gland tumours, pancreatic exocrine adenomas, Leydig cell tumours, Zymbal gland carcinomas, thyroid follicular cell adenomas or carcinomas, and uterine sarcomas in Sprague-Dawley rats reported by Hazleton Laboratories Europe Ltd. (1981a). (The tumour incidence data for each of the sites considered are presented in Tables 2 and 3.)

In the NTP study, mice were exposed to 0, 6.25, 20, 62.5, 200, or 625 ppm (0, 13.8, 44.2, 138, 442, or 1383 mg/m³) butadiene for 6 h/day, 5 days/week, for 103 weeks. Survival of mice decreased with increasing exposure concentration; therefore, to minimize the effect of the high mortality rate, the poly-3 adjusted data (Bailer & Portier, 1988; Portier & Bailer, 1989) presented in the NTP (1993) report were used in these calculations. For some tumour types, the adjusted data still demonstrated downward curvature at the highest concentration. In these cases, the high- exposure group was excluded in the determination of the TC₀₅. The TC₀₅s were calculated for these end-points by first fitting a multi stage model to the data. The multistage model is given by:

where d is dose, k is the number of dose groups in the study minus one, P(d) is the probability of the animal developing a tumour at dose d, and q_i > 0, i = 1,..., k are parameters to be estimated.

The models were fitted using GLOBAL82 (Howe & Crump, 1982), and a chi-square lack of fit test was performed for each model fit. The degrees of freedom for this test are equal to k minus the number of q_i’s whose estimates are non-zero. A p-value less than 0.05 indicates a significant lack of fit. The lower confidence limits presented are approximate, based on output from GLOBAL82. Results from the model fitting are displayed in Table A- 6. Plots of the data and the fitted models are shown in Figure A-2.

TC₀₅s were determined as the doses D (in mg/m³) that satisfy

and then adjusted by multiplying by:

where, in the first term, which amortizes the dose to be constant over the lifetime of a mouse, w is the duration of the experiment (103 weeks). The second factor was suggested by Peto et al. (1984) and corrects for an experiment length that is unequal to the standard lifetime. Since tumours develop much more rapidly later in life, a greater than linear increase in the tumour rate is expected when animals are observed for tumours longer than their standard lifetime (or the reverse, when animals are observed for a period shorter than their standard lifetime). (Application of this factor does not impact greatly on the final values, since it is very close to one.) The selected TC₀₅ values for this study and their 95% lower confi dence limits (LCLs) are presented in Table A-6 and range from 2.3 mg/m³ (95% LCL = 1.7 mg/m³) or 1.1 ppm (95% LCL = 0.79 ppm) for Harderian gland tumours in males to 99 mg/m³ (95% LCL = 23 mg/m³) or 45 ppm (95% LCL = 10 ppm) for malignant lymphomas in males.

Table A-4: Parameter estimates and model deviances for each of four lagged-exposure models fitted to median cumulative exposure per person-year.

Model	Lag	Parameter estimates	Standard error	Deviance
1) RR = (1 + dose)alpha	None	alpha = 0.2850	SE(alpha) = 0.0976	171.5
	2 years	alpha = 0.2852	SE(alpha) = 0.0982	171.6
	5 years	alpha = 0.2883	SE(alpha) = 0.0995	171.6
	10 years	alpha = 0.3064	SE(alpha) = 0.1034	171.1
	15 years	alpha = 0.2955	SE(alpha) = 0.1079	172.4
	20 years	alpha = 0.2891	SE(alpha) = 0.1141	173.6
	25 years	alpha = 0.2898	SE(alpha) = 0.1334	175.4
2) RR = 1 + beta•dosealpha	None	alpha = 0.3999	SE(alpha) = 0.2733	172.0
		beta = 0.4557	SE(beta) = 0.8219
	2 years	alpha = 0.3992	SE(alpha) = 0.2739	172.0
		beta = 0.4602	SE(beta) = 0.8279
	5 years	alpha = 0.4024	SE(alpha) = 0.2737	172.0
		beta = 0.4647	SE(beta) = 0.8288
	10 years	alpha = 0.4245	SE(alpha) = 0.2755	171.4
		beta = 0.4693	SE(beta) = 0.8345
	15 years	alpha = 0.4835	SE(alpha) = 0.3397	172.6
		beta = 0.2878	SE(beta) = 0.5846
	20 years	alpha = 0.4720	SE(alpha) = 0.3558	173.9
		beta = 0.3243	SE(beta) = 0.6572
	25 years	alpha = 0.2960	SE(alpha) = 0.2833	175.3
		beta = 0.9293	SE(beta) = 1.5710
3) RR = ebeta•dose	None	beta = 0.0029	SE(beta) = 0.0014	176.7
	2 years	beta = 0.0029	SE(beta) = 0.0015	176.8
	5 years	beta = 0.0031	SE(beta) = 0.0015	176.7
	10 years	beta = 0.0034	SE(beta) = 0.0016	176.4
	15 years	beta = 0.0035	SE(beta) = 0.0018	177.0
	20 years	beta = 0.0033	SE(beta) = 0.0022	178.2
	25 years	beta = 0.0033	SE(beta) = 0.0022	178.2
4) RR = 1 + beta•dose	None	beta = 0.0099	SE(beta) = 0.0065	174.7
	2 years	beta = 0.0102	SE(beta) = 0.0067	174.7
	5 years	beta = 0.0109	SE(beta) = 0.0072	174.6
	10 years	beta = 0.0137	SE(beta) = 0.0089	173.8
	15 years	beta = 0.0158	SE(beta) = 0.0106	174.1
	20 years	beta = 0.0179	SE(beta) = 0.0129	175.7
	25 years	beta = 0.0179	SE(beta) = 0.0129	175.7

Table A-6: Carcinogenic potency estimates (TC₀₅s)^a of butadiene based on results of bioassays in experimental animals.

Tumour type	Males
	TC05 (mg/m3)	95% LCL (mg/m3)	02 b	dfc	pd
Mice (from NTP, 1993)
Alveolar/bronchiolar adenomas or carcinomas	2.4	1.4	1.0	3	0.79
Histiocytic sarcomas	12	8.4	7.6	5	0.18
Cardiac haemangiosarcomas	14	6.4	0.34	3	0.95
Forestomach squamous cell papillomas or carcinomas	29	13	6.1	3	0.11
Ovarian granulosa cell tumours	–	–	–	–	–
Mammary gland adenoacanthomas, carcinomas, or malignant mixed tumours	–	–	–	–	–
Hepatocellular adenomas or carcinomas	3.2	1.9	2.8	2	0.24
Harderian gland adenomas or carcinomas	2.3	1.7	0.5	2	0.77
Malignant lymphomase	99	23	3.3	3	0.35
Rats (from Hazleton Laboratories Europe Ltd., 1981a)
Mammary gland adenomas or carcinomas	–	–	–	–	–
Pancreatic exocrine adenomas	597	316	1.1	1	0.29
Leydig cell tumours	161	96	0	1	–
Thyroid follicular cell adenomas or carcinomas	–	–	–	–	–
Uterine sarcomas	–	–	–	–	–
Zymbal gland carcinomas	1023	905	1	1	0.32

Table A-6: (continued)

Tumour type	Females
	TC05 (mg/m3)	95% LCL (mg/m3)	02 b	dfc	pd
Mice (from NTP, 1993)
Alveolar/bronchiolar adenomas or carcinomas	5.2	3.2	9.1	4	0.06
Histiocytic sarcomas	21	12	5.4	4	0.25
Cardiac haemangiosarcomas	7.6	5.2	18	4	0.00
Forestomach squamous cell papillomas or carcinomas	14	8.1	4.3	4	0.36
Ovarian granulosa cell tumours	6.7	4.4	5.0	3	0.17
Mammary gland adenoacanthomas, carcinomas, or malignant mixed tumours	6.7	4.9	13	4	0.01
Hepatocellular adenomas or carcinomas	5.4	3.2	0.8	3	0.85
Harderian gland adenomas or carcinomas	4.7	2.7	1.5	2	0.47
Malignant lymphomase	23	6.9	3.9	3	0.27
Rats (from Hazleton Laboratories Europe Ltd., 1981a)
Mammary gland adenomas or carcinomas	6.7	4.7	0	0	–
Pancreatic exocrine adenomas	–	–	–	–	–
Leydig cell tumours	–	–	–	–	–
Thyroid follicular cell adenomas or carcinomas	142	113	0	1	–
Uterine sarcomas	189	113	0	0	–
Zymbal gland carcinomas	4872	766	0.06	2	0.97

a Values have been adjusted for lifetime exposure.

b Chi-squared goodness of fit statistic.

c Degrees of freedom.

d p-value of goodness of fit test (p-value < 0.05 indicates significant lack of fit).

e Values for malignant lymphomas presented here only for comparison; potency estimates for these tumours not considered relevant to humans due to the greater sensitivity of these mice to induction of this effect associated with the presence of an endogenous retrovirus.

Estimates of carcinogenic potency were also calculated based on the results of the bioassay in Sprague-Dawley rats (Hazleton Laboratories Europe Ltd., 1981a). In this study, rats were exposed to 0, 1000, or 8000 ppm (0, 2212, or 17 696 mg/m³) for 6 h/day, 5 days/week, for 105 (males) or 111 (females) weeks. A high mortality rate was observed at the higher concentration; therefore, this exposure group was excluded from the analysis, except for the potency estimates for pancreatic exocrine adenomas in males (for this end-point, exclusion of the high-exposure group would have resulted in the exposure–response relationship curving downwards). As for mice, a multistage model was fit to the data for rats using GLOBAL82 and adjusted to account for study duration (w) by multiplying by:

where the duration of the experiment was 105 weeks for males and 111 weeks for females. The exposure–response curves and estimated adjusted TC₀₅ values based on this study in rats are presented in Figure A-3 and Table A-6, respectively. The concentrations of butadiene estimated to be associated with a 5% increased incidence of tumours ranged from 6.7 mg/m³ (95% LCL = 4.7 mg/m³) or 3.0 ppm (95% LCL = 2.1 ppm) to 4872 mg/m³ (95% LCL = 766 mg/m³) or 2203 ppm (95% LCL = 346 ppm) for tumours of the mammary gland and Zymbal gland in female rats, respectively. Although the available data for analysis of exposure– response were more limited for rats than for mice, it is interesting to note the similarity in estimates of potency for mammary gland tumours (i.e., 6.7 mg/m³ in both species).

Table A-5: Model validation p-values for Delzell et al. (1995) study.

	Proportion of p-valuesa that are
Model	less than 0.01	less than 0.05	less than 0.1
1) RR = (1 + dose)alpha	0.074	0.167	0.252
2) RR = 1 + beta•dosealpha	0.084	0.19	0.286
3) RR = ebeta•dose	0.08	0.188	0.264
4) RR = 1 + beta•dose	0.103	0.214	0.303

a p-value of likelihood ratio test of equality of parameters fitted to each half of the data.

Based on modelling (using THC; Howe, 1995a) of the inci dence of micronucleated polychromatic erythrocytes in B6C3F₁ mice exposed to butadiene for up to 15 months in the NTP bio assay, BMC₀₅s for somatic cell mutations were very similar to the lower end of the range of the TC₀₅s for tumour induction.

Benchmark concentrations for ovarian atrophy in mice

Benchmark concentrations (BMC₀₅s) for ovarian atrophy were derived on the basis of the chronic bioassay conducted by the NTP (1993) in which B6C3F₁ mice were exposed to concentrations of 0, 6.25, 20, 62.5, 200, or 625 ppm (0, 13.8, 44.2, 138, 442, and 1383 mg/m³) butadiene for up to 2 years. There was a concentra tion-related increase in the incidence as well as the severity of ovarian atrophy, as summarized in Table A-7.

The exposure–response relationship for ovarian atrophy from this study was quantified by fitting the following model to the dose–response data (Howe, 1995b):

where d is dose, k is the number of dose groups in the study minus one, P(d) is the probability of the animal developing the effect at dose d, and q_i > 0, i =1,..., k and d₀ are parameters to be estimated. The models were fit using THRESH (Howe, 1995b), and the BMC₀₅s were calculated as the dose D that satisfies:

A chi-square lack of fit test was performed for each of the model fits. The degrees of freedom for this test are equal to k minus the number of q_i’s whose estimates are non-zero. A p-value less than 0.05 indicates a significant lack of fit.

The BMC₀₅ was then amortized to be constant over the standard life of a mouse by multiplying by:

Resulting BMC₀₅s and lack of fit information for all models fit are displayed in Table A-8.

The model fitted to all six exposure groups exhibited a significant lack of fit, likely due to the fact that the curve rises sharply and then plateaus at the three highest exposure groups. Plots of the data and fitted model are displayed in Figure A-4. Since a good fit in the range of the BMC₀₅ (in the vicinity of 6.25 ppm [13.8 mg/m³]) is desired, the model was refitted omitting the two highest exposure groups. This model again indicates a marginal lack of fit. The graph of this model (Figure A-5) indicates that this model provides a reasonable visual fit to the data, but the resulting BMC₀₅ is uncertain due to lack of fit of the model.

The BMC₀₅ for the model excluding the two highest dose groups was calculated to be 0.57 mg/m³, with a 95% LCL of 0.44 mg/m³.

If only those animals that had moderate or marked ovarian atrophy from all exposure groups were included, the resulting BMC₀₅ would be 9.6 mg/m³ (95% LCL = 7.6 mg/m³), although there is again a significant lack of fit (Figure A-6). If the highest exposure group is excluded, the BMC₀₅ for moderate or marked ovarian atrophy becomes 3.1 mg/m³, with a 95% LCL of 2.5 mg/m³ (Figure A-7).

Table A-7: Incidence and severity of ovarian atrophy observed in 2-year bioassay in mice (NTP, 1993).

Exposure level (ppm)	Number of animals examined	All severities	Minimal severity	Mild severity	Moderate severity	Marked severity
0	49	4	1	2	1	0
6.25	49	19	0	15	4	0
20	48	32	1	23	8	0
62.5	50	42	3	18	21	0
200	50	43	0	9	34	0
625	79	69	0	19	47	3

Table A-8: Benchmark concentrations for ovarian atrophy.

Ovarian atrophy	BMC05 (ppm)	95% LCL on BMC05 (ppm)	BMC05 (mg/m3)	95% LCL on BMC05 (mg/m3)	Chi-square	df	p-value
All severities	2.5	1.9	5.6	4.1	61	4	0.00
All severities, excluding top two dose groups	0.25	0.20	0.57	0.44	7.0	2	0.03
Moderate/marked severity	4.3	3.4	9.6	7.6	37.1	4	0.00
Moderate/marked severity, excluding top dose group	1.4	1.1	3.1	2.5	2.2	3	0.55

INTERNATIONAL CHEMICAL SAFETY CARD

1,3-BUTADIENE ICSC:0017

RÉSUMÉ D’ORIENTATION

Ce CICAD sur le 1,3-butadiène a été préparé par la Direction de l’Hygiène de l’Environnement de Santé Canada sur la base d’une documentation préparée simultanément dans le cadre du Programme d’évaluation des substances prioritaires, en application de la Loi canadienne sur la protection de l’environnement (LCPE). Les évaluations sanitaires des substances prioritaires effectuées en application de cette loi portent sur les effets potentiels que ces produits pourraient avoir sur la santé humaine en cas d’exposition indirecte dans l’environnement général. Cette mise au point prend en compte des données allant jusqu’à fin avril 1998. L’appendice 1 donne des informations sur la nature de l’examen par des pairs ainsi que sur les sources documentaires utilisées. Des renseignements sur l’examen de ce CICAD par des pairs sont donnés à l’appendice 2. Ce CICAD a été approuvé en tant qu’évaluation internationale lors de la réunion du Comité d’évaluation finale qui s’est tenue à Helsinki (Finlande) du 26 au 29 juin 2000. La liste des participants à cette réunion figure à l’appendice 3. La fiche internationale sur la sécurité chimique (ICSC 0017) du 1,3-butadiène, établie par le Programme international sur la sécurité chimique (IPCS, 1993) est également reproduite dans le présent document.

Le 1,3-butadiène (No CAS 106-99-0) résulte de la combustion incomplète de certains produits lors de processus naturels ou d’activités humaines. C’est également un produit chimique industriel principalement utilisé pour la préparation de polymères, notamment du polybutadiène et des caoutchoucs et latex styrène- butadiène ou des élastomères acrylonitrile-butadiène. Le 1,3-butadiène pénètre dans l’environnement avec les gaz d’échappement des véhicules à essence ou à gazole, lors de l’utilisation de combustibles à d’autres fins que le transport, de la combustion de la biomasse ou encore lors de son utilisation sur des sites industriels.

Le 1,3-butadiène ne persiste pas dans l’environne ment mais il est néanmoins répandu dans tout le milieu urbain du fait de la présence généralisée des sources de combustion où il prend naissance. C’est dans les villes et à proximité immédiate des installations industrielles que sont mesurées les concentrations atmosphériques les plus fortes.

L’exposition de la population générale au 1,3- butadiène est due principalement à sa présence dans l’air extérieur ou intérieur. Comparativement, la contribution des autres véhicules, comme les aliments ou l’eau de boisson, reste négligeable. En revanche, celle de la fumée de tabac peut être importante.

Le métabolisme du 1,3-butadiène se révèle être de même nature d’une espèce à l’autre, mais il y a néan moins des différences quantitatives notamment en ce qui concerne la proportion de métabolites présumés toxiques qui se forment; ainsi chez la souris le métabolisme oxydatif en mono-, puis en diépoxyde est plus important que chez le rat ou l’Homme. Par ailleurs, il peut égale ment y avoir des variations interindividuelles dans la capacicité de métabolisation chez l’Homme, variations qui s’expliquent par le polymorphisme génétique des enzymes en cause.

Le 1,3-butadiène présente une faible toxicité aiguë chez les animaux de laboratoire. On a toutefois observé une atrophie ovarienne à toutes les concentrations chez des souris exposées à ce composé pendant une période prolongée. D’autres effets sur les ovaires ont également été relevés lors d’études à court terme. Chez les souris mâles, on a constaté que le 1,3-butadiène provoquait une atrophie testiculaire, mais à des concentrations supéri eures à celles qui étaient toxiques pour les femelles. Les données limitées dont on dispose ne permettent pas de conclure que ce composé est tératogène pour la progéni ture des animaux de laboratoire mâles ou femelles qui y ont été exposés, ni qu’il est véritablement toxique pour le foetus à des concentrations inférieures à celles qui sont toxiques pour la mère.

Le 1,3-butadiène exerce également divers effets sur le sang et la moelle osseuse de la souris; en ce qui con cerne le rat, les données, bien que limitées, ne mettent pas en évidence d’effets de ce genre.

L’ensemble des études montre qu’une fois inhalé, le 1,3-butadiène se révèle fortement cancérogène chez la souris, provoquant des tumeurs de localisations multiples à toutes concentrations. Il se révèle également cancéro gène chez le rat à toutes les concentrations, selon la seule étude adéquate dont on dispose. Chez cette espèce, les seules concentrations étudiées étaient beaucoup plus fortes que pour la souris, mais il apparaît néanmoins que le rat est l’espèce la moins sensible, si l’on prend comme élément de comparaison l’incidence des tumeurs. Il est probable que la plus grande sensibilité observée chez les souris est liée aux différences interspécifiques mention nées plus haut concernant le métabolisme, et notamment le métabolisme qui conduit à la formation d’époxydes actifs.

Le 1,3-butadiène a une action mutagène sur les cellules somatiques de la souris et du rat, avec une activité plus forte chez la souris. On a également constaté d’autres lésions génétiques dans les cellules somatiques de cette dernière espèce, à l’exclusion de celles du rat. Le composé s’est révélé systématiquement génotoxique pour les cellules germinales de la souris, mais apparemment pas pour celles du rat, selon la seule étude retrouvée.

Toutefois il n’y a apparemment pas de différence interspécifique pour ce qui est de la sensibilité aux effets génétiques dus aux époxydes résultant de la métabolisa tion du 1,3-butadiène. Selon certaines données con cernant l’exposition professionnelle mais qui restent toutefois limitées, le 1,3-butadiène serait également génotoxique pour l’Homme et provoquerait des lésions mutagènes et clastogènes dans les cellules somatiques.

Il existe, entre l’exposition au 1,3-butadiène sur le lieu de travail et certaines leucémies, une corrélation qui répond à plusieurs des critères d’une relation de cause à effet. L’étude la plus vaste et la plus exhaustive effectuée jusqu’ici, et qui a porté sur une cohorte de travailleurs employés dans de multiples usines, a mis en évidence un parallélisme entre une augmentation constatée de la mortalité par leucémie et l’exposition cumulée estimative au 1,3-butadiène dans les industries produisant des élastomères styrène-butadiène. La correction pour tenir compte de l’exposition au benzène et au styrène n’a pas fait disparaître cette corrélation, qui se manifestait d’ailleurs le plus fortement dans les sous-groupes potentiellement les plus exposés. On a également observé une corrélation entre l’exposition au 1,3- butadiène et les leucémies lors d’une étude cas-témoins menée indépendamment de l’enquête précitée sur une population de travailleurs en grande partie identique. Par contre, on n’a pas constaté d’augmentation de la mortalité par leucémie chez des travailleurs employés à la production de butadiène monomère mais non exposés simultanément à certaines des autres substances présentes dans l’industrie des élastomères styrène- butadiène, en dépit d’éléments d’appréciation limités concernant la possibilité d’une association avec la mortalité par lymphosarcomes ou réticulosarcomes dans certains sous-groupes.

Au vu des données épidémiologiques et toxicolo giques, le 1,3-butadiène est cancérogène pour l’Homme et pourrait également être génotoxique. Le pouvoir cancérogène (concentration qui entraîne une augmen tation de 1% de la mortalité par leucémie) a été évalué à 1,7 mg/m³ en s’appuyant sur les résultats de l’enquête épidémiologique la plus vaste et la mieux conduite sur des travailleurs exposés. Cette valeur correspond à l’extrémité inférieure de la série de concentrations tumorigènes déterminée par des études sur des rongeurs. Le 1,3-butadiène présente également une toxicité génésique chez les animaux de laboratoire. Ce potentiel toxique peut s’évaluer par la concentration de référence de 1,3-butadiène obtenue dans le cas des effets ovariens et qui est égale à 0,57 mg/m³.

Les effets sanitaires du 1,3-butadiène et le mode d’action de ce composé ont été étudiés en détail, mais d’importants travaux de recherche continuent de lui être consacrés afin de tenter de lever les incertitudes qui subsistent dans la base de données.

RESUMEN DE ORIENTACIÓN

Este CICAD sobre el 1,3-butadieno, preparado por la Dirección General de Higiene del Medio del Minis terio de Salud del Canadá, se basó en la documentación preparada al mismo tiempo como parte del Programa de Sustancias Prioritarias en el marco de la Ley Canadiense de Protección del Medio Ambiente (CEPA). El objetivo de las evaluaciones sanitarias de las sustancias priori tarias en el marco de la CEPA es valorar los efectos potenciales de la exposición indirecta en el medio ambiente general para la salud humana. En este examen se utilizaron los datos obtenidos hasta el final de abril de 1998. La información relativa al carácter del examen colegiado y a la disponibilidad del documento original se presenta en el apéndice 1. La información sobre el examen colegiado de este CICAD figura en el apéndice 2. Este CICAD se aprobó como evaluación internacional en una reunión de la Junta de Evaluación Final, celebrada en Helsinki (Finlandia) del 26 al 29 de junio de 2000. La lista de participantes en esta reunión figura en el apéndice 3. La Ficha internacional de seguridad química (ICSC 0017) para el 1,3-butadieno, preparada por el Programa Internacional de Seguridad de las Sustancias Químicas (IPCS, 1993), también se reproduce en el presente documento.

El 1,3-butadieno (CAS Nº 106-99-0) se produce por una combustión incompleta en procesos naturales y actividades humanas. Es también un producto químico industrial que se utiliza fundamentalmente en la producción de polímeros, en particular polibutadieno, cauchos y látex de estireno-butadieno y cauchos de nitrilo-butadieno. El 1,3-butadieno se incorpora al medio ambiente a partir de las emisiones de los gases de escape de los vehículos con motor de gasolina y diésel, de la combustión de combustibles fósiles no utilizados en el transporte, de la combustión de biomasa y de usos industriales sobre el terreno.

El 1,3-butadieno no es persistente, pero sí ubicuo en el medio ambiente urbano, debido a la presencia generalizada de sus fuentes de combustión. Las concentraciones atmosféricas más altas se han medido en el aire de las ciudades y en las cercanías de fuentes industriales.

La población general está expuesta al 1,3-buta dieno fundamentalmente mediante el aire del ambiente y de los espacios cerrados. En comparación, otros medios, entre ellos los alimentos y el agua de bebida, contrib uyen de manera insignificante a la exposición a él. El humo del tabaco puede producir cantidades significa tivas de1,3-butadieno.

El metabolismo del 1,3-butadieno parece ser cualitativamente semejante en todas las especies, aunque hay diferencias cuantitativas en los metabolitos supues tamente tóxicos que se forman; los ratones parecen oxidar el 1,3-butadieno a monoepóxido y después a diepóxido en mayor medida que las ratas o las personas. Sin embargo, la capacidad metabólica para el 1,3- butadieno puede variar de unas personas a otras, en función del polimorfismo genético de las enzimas correspondientes.

El 1,3-butadieno tiene una toxicidad aguda baja en animales experimentales. Sin embargo, la exposición prolongada de ratones a este producto se asoció con la aparición de atrofia ovárica con todas las concentra ciones utilizadas en los ensayos. En estudios de expo sición más breve se observaron también efectos en los ovarios. En los ratones machos se detectó asimismo atrofia testicular a concentraciones superiores a las asociadas con efectos en las hembras. Sobre la base de los limitados datos disponibles, no hay pruebas conclu yentes de que el 1,3-butadieno sea teratogénico en los animales experimentales tras la exposición materna o paterna o de que induzca una toxicidad fetal importante a concentraciones inferiores a las que provocan toxici dad materna.

El 1,3-butadieno indujo también diversos efectos en la sangre y en la médula ósea de ratones; aunque los datos son limitados, no se han observado efectos seme jantes en las ratas.

El 1,3-butadieno inhalado es un potente carcinó geno en ratones, induciendo tumores en lugares múltiples a todas las concentraciones utilizadas en todos los estudios identificados. Fue también carcinogénico en ratas a todos los niveles de exposición en el único estudio disponible al respecto; aunque en las ratas sólo se ensayaron concentraciones mucho más altas que en los ratones, las ratas parecen ser la especie menos sensible, tomando como base la comparación de los datos sobre la incidencia de tumores. La mayor sensibilidad de los ratones que las ratas a la inducción de estos efectos por el 1,3-butadieno probablemente se debe a diferencias entre las especies con respecto al metabolismo de formación de epóxidos activos.

El 1,3-butadieno es mutagénico en células somáticas tanto de ratones como de ratas, aunque la potencia mutagénica fue mayor en los primeros. De la misma manera, el 1,3-butadieno indujo otros daños genéticos en células somáticas de ratones, pero no en las de ratas. Fue también sistemáticamente genotóxico en células germinales de ratones, pero no en la única valoración identificada en ratas. Sin embargo, no se observaron diferencias aparentes en la sensibilidad de las especies a los efectos genéticos inducidos por los metabolitos epóxidos del 1,3-butadieno. Hay también pruebas limitadas de poblaciones expuestas en el lugar de trabajo de que el 1,3-butadieno es genotóxico para las personas, induciendo daños mutagénicos y clastogénicos en las células somáticas.

La asociación entre la exposición al 1,3-butadieno en el entorno de trabajo y la leucemia cumple varios de los criterios tradicionalmente establecidos para la causalidad. En el estudio más amplio y completo realizado hasta el momento, utilizando una cohorte de trabajadores de diversas instalaciones, la mortalidad por leucemia aumentó con una exposición acumulativa estimada al 1,3-butadieno en la industria del caucho de estireno-butadieno; esta asociación se mantuvo tras el control de la exposición al estireno y al benceno y alcanzó la intensidad máxima en los subgrupos con el mayor potencial de exposición. Análogamente, en un estudio de casos y testigos independiente basado prác ticamente en la misma población de trabajadores se observó una asociación entre la exposición al 1,3-buta dieno y la leucemia. Sin embargo, no se produjo un aumento de la mortalidad a causa de la leucemia en los trabajadores de la producción de monómeros de buta dieno que no estaban simultáneamente expuestos a algunas de las otras sustancias presentes en la industria del caucho de estireno-butadieno, aunque en algunos subgrupos se obtuvieron pruebas limitadas de asociación con la mortalidad por linfosarcoma y reticulosarcoma.

Los datos epidemiológicos y toxicológicos dispon ibles demuestran que el 1,3-butadieno es carcinogénico, y también puede ser genotóxico, para las personas. Se calculó una potencia carcinogénica (concentración que produce un aumento del 1% de la mortalidad por leu cemia) de 1,7 mg/m³, sobre la base de los resultados de la mayor investigación epidemiológica bien realizada en trabajadores expuestos. Este valor es semejante al nivel más bajo de la gama de concentraciones tumorígenas determinadas a partir de los estudios realizados con roedores. El 1,3-butadieno indujo también toxicidad reproductiva en animales experimentales. Como medida de su potencia para inducir efectos reproductivos, se obtuvo una concentración de referencia de 0,57 mg/m³ para la toxicidad ovárica en los ratones.

Aunque se han investigado a fondo los efectos en la salud asociados con la exposición al 1,3-butadieno y el mecanismo de acción para la inducción de estos efectos, se siguen realizando numerosas investigaciones sobre esta sustancia, a fin de tratar de abordar algunas de las incertidumbres asociadas con la base de datos.

FOOTNOTES

1 International Programme on Chemical Safety (1994) Assessing human health risks of chemicals: derivation of guidance values for health-based exposure limits. Geneva, World Health Organization (Environmental Health Criteria 170).

2 Personal communication from L.A. Graham, River Road Environmental Technology Centre, Environment Canada, Ottawa, Ontario, to Commercial Chemicals Evaluation Branch, Environment Canada, Hull, Quebec, 1996.

3 Hazardous Substances Databank, National Library of Medicine’s TOXNET system, searched 10 December 1999.

4 Unpublished data on butadiene levels in Canada from National Air Pollution Surveillance program, provided by T. Dann, River Road Environmental Technology Centre, Environment Canada, Ottawa, Ontario, to Commercial Chemicals Evaluation Branch, Environment Canada, Hull, Quebec, April 1997.

5 Also letter dated 28 August 1996 from P. Steer, Science and Technology Branch, Ontario Ministry of Environment and Energy, to J. Sealy, Health Canada, re. 1,3-butadiene and chloroform data (File No. 1E080149.MEM).

6 Personal communication from H. Michelin, Bayer Inc., Sarnia, Ontario, to Commercial Chemicals Evaluation Branch, Environment Canada, Hull, Quebec, 1997.

7 Also personal communication dated 24 December 1997 from X.-L. Cao to Health Canada, re. method detection limits for 24-h air samples from multimedia exposure pilot study (File No. MDL.XLS).

8 Personal communication from R. Cooper, Department of Biomedical and Environmental Health, School of Public Health, University of California, Berkeley, California, 1989 (cited in CARB, 1992).

9 Also personal communication (correspondence dated 25 March 1998) from J.A. Swenberg, University of North Carolina, Chapel Hill, NC, to Health Canada.

10 The terminology for cancers of the lymphohaematopoietic system is that used by authors of the individual study accounts.

11 SMRs are presented here in the format used by the authors; i.e., SMR = observed/expected or SMR = observed/expected × 100.

12 It is not possible to determine, with any certainty, the size of the population in these earlier studies that was not subsumed in the later investigation by Delzell et al. (1995). One of the small plants of approximately 600 workers included in the Matanoski et al. (1990, 1993) cohort was not examined by Delzell et al. (1995).

13 Also personal communication (letter dated 17 October 1997) from J.B. Ward, Jr., Division of Environmental Toxicology, Department of Preventative Medicine and Community Health, University of Texas Medical Branch, Galveston, TX, to Health Canada.

14 Personal communication (electronic correspondence dated 15 November 1997) from J.B. Ward, Jr., Division of Environ mental Toxicology, Department of Preventative Medicine and Community Health, University of Texas Medical Branch, Galveston, TX, to Health Canada.

15 Also personal communications (letter dated 17 October 1997 and electronic correspondence dated 15 November 1997) from J.B. Ward, Jr., Division of Environmental Toxicology, Department of Preventative Medicine and Community Health, University of Texas Medical Branch, Galveston, TX, to Health Canada.

16 Also personal communication (correspondence dated 30 March 1998) from R.D. Irons, University of Colorado Health Sciences Center, Denver, CO, to Health Canada.

17 Personal communication (electronic correspondence dated 26 June 1998) from B. Davis, National Institute for Environmen tal Health and Safety, National Toxicology Program, Research Triangle Park, NC, to Health Canada.

18 The potency estimate for carcinogenicity is determined by calculating the dose or concentration associated with an increase in cancer incidence or mortality of an appropriate percentage. When based on toxicological data from studies in experimental animals, a 5% increase is generally chosen, as these values usually lie within or close to the observable range (i.e., a TC₀₅ is calculated). When epidemiological data form the basis for derivation of a tumorigenic concentration, the percent increase selected is that which falls within the area of the exposure–response curve that represents the majority of the observable data; this is often less than 5%. In the case of butadiene, the carcinogenic potency calculated on the basis of modelling of epidemiological data (as described herein) was considered to be best defined as a 1% increase in mortality due to leukaemia (i.e., a TC₀₁).

19 Similar to tumorigenic concentrations (TC₀₅s), benchmark concentrations for non-cancer effects (or BMC₀₅s), when based on data in experimental animals, represent the dose or concen tration associated with a 5% increase in the incidence of an effect compared with controls.

20 Unpublished data on butadiene levels in Canada from National Air Pollution Surveillance program, provided by T. Dann, River Road Environmental Technology Centre, Environment Canada, Ottawa, Ontario, to Commercial Chemicals Evaluation Branch, Environment Canada, Hull, Quebec, April 1997.

21 Although it has not been possible to quantitatively character ize uncertainty regarding these estimates of exposure and the impact of this uncertainty upon the estimates of carcinogenic potency, data being collected currently may permit a more quantitative characterization in future (personal communica tion [correspondence dated 20 March 1998] from J. Lynch, Consultant, Rumson, NJ, to Health Canada).

22 The cooperation of the sponsors and researchers for the Delzell et al. (1995) study in the provision of these data is gratefully acknowledged.

23 Epicure is a collection of interactive programs used to fit models to epidemiological data. The specific program used to model the data for this cohort of styrene-butadiene rubber workers is called AMFIT, which is specially designed to model hazard functions for censored cohort survival data. The strength of Epicure lies in its ability to easily allow the background rate to depend on user-specified strata, such as age, calendar period, and race.

24 Mortality data were provided to Health Canada by Statistics Canada. The cooperation of the registrars of vital statistics in the provinces and territories of Canada who make mortality data available to Statistics Canada under federal–provincial agreements is gratefully acknowledged.

    See Also:
       Toxicological Abbreviations