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Benzene - PSL1
2.0 Summary of Critical Supporting Data
2.1 Identity and Physical/Chemical Properties of Substance
Benzene (CAS Registry Number 71-43-2) is a simple cyclic organic compound with molecular formula C6H6. It is a volatile, clear, flammable, colourless liquid at room temperature, and has an aromatic odour. Benzene is miscible in most common organic solvents. It has a relatively high vapour pressure (10.1 to 13.2 kPa at 25 °C), a high water solubility (820 to 2167 mg/L at 25°C), and a low log octanol/water partition coefficient (1.56 to 2.69) (Mackay et al., 1992). Benzene does not appreciably absorb light of wavelengths greater than 260 nm (Bryce-Smith and Gilbert, 1976) or infrared radiation at wavelengths of 7 to 13 µm (Sadtler Research Laboratories, 1982).
2.2 Production and Uses
Benzene can be produced commercially from petroleum, natural gas condensates, or coal. Most isolated (purified) benzene produced in Canada is derived from petroleum sources through catalytic reforming of naphtha, dealkylation of toluene, and separation of pyrolysis gasoline (Hancock, 1975; Allison and Brown, 1977).
A survey of commercial use patterns (CIS, 1991) indicates that 765 kilotonnes of isolated benzene were produced in Canada in 1990 and 131 kilotonnes were imported, for a total Canadian supply of 896 kilotonnes. Of these, 74 kilotonnes were exported, resulting in total domestic consumption of 822 kilotonnes of isolated benzene. Isolated benzene is produced at four industrial plants in the Sarnia/Corunna area in Ontario, at two plants in Alberta, and at two plants in Montreal, Quebec.
Benzene is used extensively in industry as a volatile solvent and as an intermediate in the production of many chemicals including ethylbenzene/styrene, cumene, and maleic anhydride (Jaques, 1990).
Benzene is also a natural component of petroleum (Kirk et al., 1983). In gasoline, benzene acts as an octane-enhancer and an anti-knock agent. An estimated 35 000 megalitres of gasoline were consumed in Canada in 1989 (Statistics Canada, 1989). Based on an average benzene content in premium and regular unleaded gasolines of 2.15% by weight or 1.76% by volume (Madé, 1991), an estimated 540 kilotonnes of benzene are present in the gasoline sold annually in Canada; most of this benzene is burned during normal engine operation. The total yearly consumption of benzene in Canada, including both isolated benzene and benzene as a component of gasoline, is therefore estimated to be 1362 kilotonnes.
2.3 Sources and Releases
Benzene is an organic compound found naturally in the environment in low concentrations. It is a component of crude oil and is formed through incomplete combustion of organic materials. Benzene enters water and soil through petroleum seepage and weathering of exposed coal-containing strata. It enters groundwater from petroliferous rocks, and air from volcanoes, forest fires, and releases of volatile chemicals from plants (Graedel, 1978; Westberg et al., 1981; Whelan et al., 1982; Fishbein, 1984; Slaine and Barker, 1990). The magnitude of emissions from natural sources is not known but, based on concentrations in rural areas, it is believed to be generally low in comparison with anthropogenic sources (Rasmussen and Khalif, 1983; Rudolph et al., 1984).
Benzene can enter the environment from any stage involved in production, storage, use, and transport of isolated benzene, and crude oil and gasoline, including emissions resulting from fuel combustion.
It has been estimated that in 1985, 34 150 tonnes of benzene were released into the atmosphere in Canada (Jaques, 1990). Major sources were combustion of gasoline and combustion of diesel fuels, which together accounted for 76% of total atmospheric releases. Light-duty vehicles accounted for 61% of total releases. Other sources of release to the atmosphere included emissions during benzene production (6.5% of total releases); other chemical production (7.7%); primary iron and steel production (1.0%); solvent uses (1.5%); residential fuel combustion (4.1%); and gasoline marketing (1.9%). Total emissions of benzene to the atmosphere are expected to decline in the future, primarily because of the planned reduction of emissions of volatile organic compounds (VOCs) from light-duty vehicles and the efforts to reduce VOC emissions from a variety of other sources in order to control ground-level ozone (CCME, 1990).
Benzene can enter soil from oil and gasoline spills, leaking underground storage tanks, and seepage from waste disposal sites (U.S. EPA, 1980; Johnson et al., 1989). Contamination of surface water may result from spills of chemicals and petroleum products and from discharges of industrial and municipal effluents (U.S. EPA, 1980; Ontario Ministry of the Environment, 1992). Estimates of total environmental loadings from such sources in Canada are not available.
It is estimated that every year in Canada, 34 kilotonnes of benzene are released into the atmosphere, 1 kilotonne into water, and 0.2 kilotonnes onto soil. These figures are based on proportions of benzene released to air, water, and soil in the United States (Slimak and Delos, 1983) and the Netherlands (National Institute of Public Health and Environmental Hygiene, 1988) and on data on releases to the atmosphere in Canada (Jaques, 1990).
2.4 Environmental Fate and Concentrations
2.4.1 Fate
Mechanisms affecting the environmental fate of benzene include photo-oxidation (Guesten et al., 1981; Tully et al., 1981; Besemer, 1982; Mill, 1982; Atkinson, 1985; Japar et al., 1991), volatilization (Thomas, 1982), advection (Mackay et al., 1992), and biodegradation (Horowitz et al., 1982; Vaishnav and Babeu, 1987). The atmosphere and surface waters should be the major sinks for benzene because of its relatively high vapour pressure, high water solubility, and low octanol/water partition coefficient.
Processes in the atmosphere should play a determining role in benzene's ultimate fate in the environment (Mackay and Paterson, 1991; Mackay et al., 1992).
Photo-oxidation is the major degradation pathway for benzene in air. Benzene is oxidized in reactions with hydroxyl radicals and, to a lesser extent, tropospheric ozone and nitrate radical (NO3). Under typical urban atmospheric conditions, half-lives attributable to reactions with hydroxyl radicals were calculated to be 9 days, more than 235 days with nitrate radical, and more than 470 days with ozone (Finlayson-Pitts and Pitts, 1986). Other estimates for overall half-lives of benzene have ranged from 0.1 to 21 days (Darnall et al., 1976; Atkinson, 1985; Howard et al., 1991). Major products of photo-oxidation include: phenol, nitrophenol, nitrobenzene, glyoxal, butanedial, formaldehyde, carbon dioxide, and carbon monoxide (Nojima et al., 1975; Finlayson-Pitts and Pitts, 1986). Since the atmospheric half-life of benzene is relatively short, long-range transport of benzene is unlikely.
Volatilization and biodegradation are the major processes involved in the removal of benzene from water. The half-life of benzene in water 1 metre deep was estimated to be 4.8 hours as a result of volatilization (Agency for Toxic Substances and Disease Registry, 1989). Reported half-lives of benzene have ranged from 33 to 384 hours for aerobic biodegradation in surface waters (van der Linden, 1978; Tabak et al., 1981; Mills et al., 1982; Vaishnav and Babeu, 1987). For anaerobic biodegradation in deeper waters or in groundwater, half-lives ranged from 28 days to 720 days (Horowitz et al., 1982; Vaishnav and Babeu, 1987; Howard et al., 1991).
The primary mechanisms responsible for loss of benzene from soil are volatilization to the atmosphere and runoff to surface water. Biodegradation also accounts for a small proportion of loss (Scheunert et al., 1985; National Institute of Public Health and Environmental Hygiene, 1988). Benzene released below the soil surface, for example from leaking underground storage tanks, can leach into groundwater. With organic carbon sorption coefficients (KOCs) reported for benzene ranging from 12 to 213, benzene is considered to be moderately to highly mobile in soil (Karickhoff et al., 1979; Rogers et al., 1980; Korte et al., 1982).
Using the Level III Fugacity Modelling developed for southern Ontario (Mackay, 1991), the overall residence time in the environment was predicted to be short (3.5 days, considering both degradation and movement of benzene out of the area) and the reaction residence time was short also (9.7 days, considering loss through degradation reactions only).
Benzene does not bioconcentrate in aquatic biota to a significant degree. Relatively low bioconcentration factors (BCFs) have been reported for aquatic bacteria, algae, macrophytes, and fish. The highest reported value was for Daphnia pulex, with a BCF of 225 (log BCF of 2.35) (Trucco et al., 1983). Once the organisms are removed from contaminated water, benzene is rapidly cleared by the organisms. For Daphnia pulex, 85% of accumulated benzene was removed during the 72 hours following withdrawal from contaminated water (Trucco et al., 1983). The depuration of benzene in fish is also rapid. Half-lives were estimated to be less than 0.5 days in eel, Anguilla japonica (Ogata and Miyake, 1978), and less than 1 day in striped bass, Morone saxatilis (Niimi, 1987).
2.4.2 Concentrations
Mean concentrations of benzene in 586 samples of ambient air in ten Canadian cities surveyed between 1988 and 1990 ranged from 1.2 to 14.6 µg/m3, with a maximum 24-hour average concentration of 41.9 µg/m3; the overall mean concentration was 4.4 µg/m3 (Dann, 1991). Similar levels were reported in a more recent survey of eleven Canadian cities, while mean concentrations of benzene in three rural locations ranged from 0.6 to 1.2 µg/m3 (Dann and Wang, 1992). Airborne concentrations of benzene at the perimeter of gasoline service stations in five Canadian cities averaged 439 µg/m3 (maximum of 6834 µg/m3) in the summer of 1985 (PACE, 1987) and 1383 µg/m3 (maximum of 16 246 µg/m3) in the winter of 1986 (PACE, 1989). Mean short-term (10 to 15 minutes) airborne concentrations during refuelling ranged from 2600 to 4400 µg/m3 (PACE, 1987; 1989).
A major source of benzene in indoor air is cigarette smoke; smoke actually inhaled (mainstream smoke) contains 12 to 48 µg per cigarette, while amounts in smoke emitted from cigarettes (sidestream smoke) are approximately ten times greater (U.S. Department of Health and Human Services, 1986). Based on data obtained in 200 homes in the United States (Wallace, 1989), tobacco smoking is estimated to contribute an additional 3 µg/m3 to the concentration of benzene in residential indoor air. Various household and other products appear to contribute to the concentration of benzene in residential indoor air. The contribution of these products to the benzene content of indoor air, indirectly determined from the differences in reported concentrations in indoor air in the homes of nonsmokers and the corresponding concentrations in ambient air in a survey of homes in the United States (Wallace et al., 1987; Wallace, 1989), has been estimated to be 2 µg/m3 (Holliday and Park, 1989).
Benzene has been measured in Canadian surface waters. In surveys at 10 sites along the Great Lakes and 30 water treatment facilities across Canada, benzene concentrations in untreated water were generally lower than the detection limit (0.1 or 1 µg/L, respectively); the highest reported mean concentration was 2 µg/L (Otson et al., 1982; Otson, 1987). Concentrations of benzene along a 6-km industrialized section of the St. Clair River near Sarnia, Ontario, ranged from below the detection limit (0.1 µg/L) to 4.3 µg/L (Comba and Kaiser, 1987). Benzene levels were below the detection limit upstream from the industrialized section and returned to near or below detection levels about 1 km downstream. A mean benzene concentration of 0.45 µg/L was calculated for the sampling stations along the industrialized section of the river. In Ontario, the highest concentrations of benzene in untreated effluents released into surface water were reported from the organic chemical manufacturing sector; the highest 12-month average concentration of benzene at one outfall was 65.3 µg/L (Ontario Ministry of the Environment, 1992). Benzene was not reported to occur in water at concentrations above the detection limit of 1 µg/L in other surveys of Canadian waters (NAQUADAT, 1991).
Benzene has been measured in groundwater in areas where underground storage tanks containing gasoline have leaked, and near landfill sites. In some sites, benzene concentrations in the groundwater have ranged from below the detection limit (Barker et al., 1988; Intera Technologies Ltd., 1987; Water and Earth Science Associates Ltd., 1988) to 15 mg/L (Jackson et at., 1985).
Few data are available on the concentrations of benzene in drinking water in Canada. Benzene was detected (quantitation limit, 1 µg/L) in 50 to 60% of potable water samples from 30 treatment facilities in a national survey conducted in 1979; mean concentrations of benzene in treated water ranged from <1 to 3 µg/L and the maximum value was 47 µg/L (Otson et al., 1982). Benzene has rarely been detected in provincial monitoring programs at concentrations greater than 1 µg/L (Ayotte, 1987; O'Neill and MacKeigan, 1987a, 1987b, 1987c, 1987d; Ontario Ministry of the Environment, 1989).
Data on the occurrence of benzene in food are very limited. Although it has been detected in individual foodstuffs at concentrations of up to 2100 µg/kg, benzene was not detected in several foods representative of a "typical" U.S. diet, with detection limits ranging up to 0.66 µg/kg (Rose and Chin, 1990).
2.5 Toxicokinetics
While benzene is believed to be readily absorbed from the gastrointestinal tract, it is estimated that approximately 50% of inhaled benzene is absorbed through the lungs, and only very small amounts through the skin (Agency for Toxic Substances and Disease Registry, 1989). Absorbed benzene is distributed throughout the body, with the possibility of some accumulation in adipose tissue. Metabolism of benzene occurs largely in the liver, although some metabolism may take place in the bone marrow. The pathways of benzene metabolism appear to be qualitatively similar in humans and experimental animals, although there may be quantitative differences in the proportion of putatively toxic metabolites in various species. The pathway leading to the formation of the putatively toxic metabolites (benzoquinone and muconaldehyde) appears to be a saturable process at relatively low doses; as a result, the proportion of toxic metabolites formed is greater at low doses than at high doses (Henderson et al., 1989, 1990; Medinsky et al., 1989). The metabolites of benzene are largely excreted in the urine, while unmetabolized benzene is eliminated by exhalation. With absorption of increasing amounts of benzene, a greater proportion is exhaled unchanged than is excreted as metabolites in urine.
2.6 Mammalian Toxicology
Benzene is not highly acutely toxic to experimental animals. Hematological effects similar to those observed in humans have been reported in animals following short-term, subchronic, or chronic exposure to benzene. It has been consistently observed in these studies that lymphocyte levels are depressed most severely and in the shortest time, while granulocytes appear to be the most resistant of the circulating cells, and that anemia does not occur as frequently as lymphocytopenia (Agency for Toxic Substances and Disease Registry, 1989).
In recent studies, benzene has been carcinogenic in two species of experimental animals, inducing a wide variety of tumors following inhalation (Table 1) and ingestion (Table 2). Based on the results of in vitro and in vivo studies in experimental animals, benzene appears to induce clastogenic damage to DNA rather than causing point mutations.
Benzene is not teratogenic in experimental animals, although embryotoxic and fetotoxic effects have been reported at airborne concentrations less than those observed to be toxic to the mothers (as low as 47 ppm or 150 mg/m3 in rats) (Tatrai et al., 1980). Hematological changes have also been noted in mice exposed to 5 ppm (16 mg/m3) benzene in utero (Keller and Snyder, 1986).
Concentrations of benzene as low as 10 ppm (32 mg/m3) have been reported to cause immunological effects (depression of the response of B cells and T cells) in rats (Rozen et al., 1984). Exposure to benzene at concentrations as low as 100 ppm (320 mg/m3) has also been associated with neurological effects and behavioural disturbances in animals similar to those caused by other petroleum hydrocarbons (Agency for Toxic Substances and Disease Registry, 1989; Dempster et al., 1984).
2.7 Effects on Humans
In epidemiological studies, hematotoxic effects have been reported in several populations occupationally exposed to benzene, due to damage or depression of the hematopoietic system. Depression in bone marrow activity results from damage to or destruction of the pluripotential stem cells and/or the early proliferating committed cells. In several studies, workers occupationally exposed to benzene have developed pancytopenia which in more severe cases is referred to as aplastic anemia. Kipen et al. (1988) reported significant decreases in white and red blood cell counts and hemoglobin in workers exposed during the 1940s in the cohort of pliofilm workers studied by Rinsky et al. (1987). Additional work on the hematological effects, particularly during the early years of employment, in workers in this cohort is under way (Cody et al., in press), since it has been suggested that similar decreases in blood cell counts could be found in pre-employment tests, and the correlation with benzene exposure is artifactual (Hornung et al., 1989). Effects on the immune system, including decreases in T lymphocytes (Moszczynski, 1981), alterations in serum immunoglobulins and complement levels, and symptoms of benzene-induced autoimmunity and allergy, have been observed in workers occupationally exposed to benzene (Agency for Toxic Substances and Disease Registry, 1989).
Associations between leukemia and exposure to benzene in occupationally exposed populations have been observed in numerous case studies, and in the majority of epidemiological studies conducted to date (see Tables 3 and 4). In addition, there was a clear exposure-response relationship in the population for which exposure has been most extensively characterized (Rinsky et al., 1987). However, information in only three Studies (Bond et al., 1986; Wong, 1987a, 1987b; Rinsky et al., 1987) is considered sufficient to form the basis of a quantitative assessment of carcinogenic potency, although the numbers of deaths due to leukemia were small in each investigation. The other studies are less relevant owing to limitations that include poor characterization or description of the basis for estimation of exposure, concomitant exposure to substances other than benzene, and/or the low number of observed cases. For example, although 25 deaths due to leukemia were reported in the historical cohort study of workers in various industries by Yin et al., (1987), the published report did not include sufficient information to form the basis for characterization of individual exposure.
| Species | Protocol | Treatment-related Effects | Reference |
|---|---|---|---|
| C57BL/6J mice | 100 and 300 ppm (319 and 958 mg/m3) lifetime | lymphocytic lymphoma (thymic involvement), myeloma, leukemia | Snyder et al., 1980 |
| CD-1 mice and Sprague-Dawley rats | 100 and 300 ppm (319 and 958 mg/m3) lifetime | acute and chronic myelogenous leukemia | Goldstein et al., 1982 |
| Sprague-Dawley rats, breeders and embryos | 200 and 300 ppm (639 and 958 mg/m3), 15 or 104 weeks, beginning at 12 days for 150 weeks | breeders: zymbal gland carcinoma, hepatoma, mammary carcinoma; offspring: zymbal gland carcinoma, nasal carcinoma, hepatoma, leukemia, mammary carcinoma | Maltoni et al., 1985 |
| C57BL/6 mice | 300 ppm (958 mg/m3) for 16 weeks, lifetime observation | thymic lymphoma, unspecified lymphoma | Cronkite et al., 1984 |
| Sprague-Dawley rats | 100 ppm (319 mg/m3) lifetime | zymbal gland carcinoma, liver hemangioma, hepatoma, liver hemangio-endothelioma and fibrosarcoma, chronic granulocytic leukemia, mammary carcinoma | Snyder et al., 1984 |
| CBA/Ca mice | 100 and 300 ppm (319 and 958 mg/m3) for 16 weeks, 900 days observation | leukemia | Cronkite, 1986 |
| C57B1 mice and CD-1 mice | 300 ppm (958 mg/m3) one week in three for life, 1200 ppm 3834 mg/m3) for 10 weeks, observed for life | 300 ppm (958 mg/m3): lung adenoma in CD-1 mice, zymbal gland carcinoma in C57B1 mice; 1200 ppm (3834 mg/m3): lung adenoma and zymbal gland carcinoma in CD-1 mice | Snyder et al., 1988 |
| * Based on the original "author call" of possible or direct correlation, as presented in Agency for Toxic Substances and Disease Registry (1989), with the exception of Snyder et al., (1988) which underwent primary critical review by Health and Welfare Canada staff. | |||
| Species | Protocol | Treatment-related Effects | Reference |
|---|---|---|---|
| Sprague-Dawley rats | 50 and 250 mg/(kg b.w.·day), 52 weeks, observed for 144 weeks; 500 mg/(kg b.w.·day),104 weeks, observed for 141 weeks | zymbal gland carcinoma, carcinoma of oral and nasal cavities, hemolymphoreticular neoplasms, other malignant tumors | Maltoni et al ., 1985 |
| Wistar rats | 500 mg/(kg b.w.·day), 100 weeks | zymbal gland carcinoma, carcinoma of oral cavity, thymoma, other hemolymphoreticular neoplasms | Maltoni et al ., 1985 |
| Swiss mice | 500 mg/(kg b.w.·day), 78 weeks, observed for 100 weeks | zymbal gland carcinoma, pulmonary and mammary tumors | Maltoni et al ., 1985 |
| F344/N rats | 25 to 200 mg/(kg b.w.·day), 2 years | carcinoma of oral cavity, zymbal gland carcinoma, skin carcinoma | National Toxicology Program, 1986 |
| B6C3F1 mice | 25 to 100 mg/(kg b.w.·day), 2 years | zymbal gland carcinoma, malignant lymphoma, alveolar/bronchiolar carcinoma, alveolar/ bronchiolar adenoma, harderian gland adenoma, preputial gland carcinoma, ovarian granulosa cell tumor, mammary gland carcinoma, mammary gland carcinosarcoma | National Toxicology Program, 1986 |
| * Based on the "author call" of possible or direct correlation, as presented by Agency for Toxic Substances and Disease Registry (1989). | |||
| Number and Type of Subjects | Exposure Measure | Standardized Mortality Ratio for Leukemia* | Reference |
|---|---|---|---|
| 38 000 petro chemical industry workers Referent: general population | Potential occupational exposure to > 1% (10 000 ppm) benzene for > 5 years | SMR = 121 | Thorpe, 1974 |
| 594 benzene workers Referent: general population | Cumulative exposure (ppm-months) | SMR = 200 | Ott et al., 1978 |
| 259 chemical workers Referent: general population | Employment at plant that used large quantities of benzene | SMR = 682 | Decouflé et al., 1983 |
| 454 oil refinery workers Referent 1) general population, 2) non-exposed workers | Occupational exposure to benzene < 1 to > 10 ppm (<3.19 to >31.9 mg/m3) | Obs = 0 Exp = 0.42 |
Tsai et al., 1983 |
| 1 361 graphic industry workers | Not specified | SMR = 250 (total leukemia) | Paganini-Hill et al., 1980 |
| 13 570 rubber industry workers | Not specified | SMR = 240 (lymphatic) | Monson and Fing, 1978 |
| 28 460 workers from various industries Referent: 28 257 non-exposed workers | 10 to 1000 mg/m3, 50 to 500 mg/m3, (most areas) | SMR = 574 (total leukemia) | Yin et al., 1987 |
| 33 815 rubber industry workers | "low exposure" | Parkes et al., 1982 | |
| 34 781 petrochemical | "low exposure" | Rushton and | |
| industry workers | Alderson, 1980 | ||
| * Type of leukemia specified where available. | |||
| Observation Period | Number and Type of Subjects | Exposure Measure | Relative Risk for Leukemia* | Reference |
|---|---|---|---|---|
| 1966 to 1969 | 257 leukemia cases 124 hospital controls |
occupational or household exposure to solutions with benzene or toluene | RR = 3.3 (acute leukemia) RR = 4.1 (chronic lymphocytic leukemia) RR = 1.8 (myelocytic leukemia) |
Girard et al., 1970 |
| 1945 to 1967 | 303 adult leukemia cases 303 controls |
potential occupational exposure to benzene or X-rays | RR = 2.5 | Ishimaru et al., 1971 |
| 1955 to 1974 | 138 adult leukemia cases 276 controls |
medical record of benzene exposure | RR = 3.3 | Linos et al., 1980 |
| 1950 to 1975 | Oil refinery workers 36 cases 216 controls |
low, medium, or high occupational benzene exposure | RR = 2.0 (high or medium vs. low) | Rushton and Alderson, 1981 |
| 1964 to 1973 | Rubber workers 15 lymphocytic leukemia cases 30 controls |
primary exposure from job involving direct use of benzene | RR = 4.5 (lymphocytic leukemia) | Arp et al., 1983 |
| 1964 to 1973 | 11 lymphocytic leukemia cases 1 350 controls |
occupational exposure in work areas where benzene was used | RR = 2.5 (lymphocytic leukemia) | Checkoway et al., 1984 |
| * Type of leukaemia specified where available. | ||||
In the cohort of 956 chemical workers studied by Bond et al. (1986), there was a nonsignificant excess of deaths due to leukemia compared to national rates (4 versus 2.1). However, the observed and expected numbers of deaths due to leukemia (observed:expected = 3:1.9, when individuals exposed to arsenic, asbestos, or vinyl chloride were excluded) were small. Although there was a significant excess of deaths due to skin cancer in the cohort excluding those exposed to arsenic, asbestos, or vinyl chloride, all of these cases occurred in the group with the lowest estimated cumulative exposure to benzene. There was no relationship between the excess of deaths due to leukemia and area of work, duration of employment, or cumulative exposure, which may be attributable to the small numbers observed.
In the cohort of 7676 workers at seven chemical plants examined by Wong (1987a, 1987b), there was also an excess of deaths due to leukemia when compared to national rates (not statistically significant). Again, however, the total number was small (observed:expected = 6:4.43 in the continuously exposed group). Mortality due to lymphatic or hematopoietic cancer was significantly increased in the intermittently and continuously exposed groups combined, compared to unexposed workers (19 versus 3), as was the number of deaths due to leukemia (7 versus 0). This may be attributable, however, to a deficit of deaths due to leukemia in the latter group (i.e., 0 observed, 3.4 expected). There was an increasing trend in the standardized mortality ratios (SMRs) for lymphopoietic cancer, leukemia, and non-Hodgkin's lymphopoietic cancer with cumulative exposure in the group that had experienced continuous exposure to benzene. None of the observed leukemias in this cohort was the type most often observed in workers exposed to benzene, i.e., acute myelogenous leukemia.
Rinsky et al. (1987) examined the mortality of a cohort of 1165 pliofilm workers exposed to benzene, which was the only hematotoxic solvent to which employees were exposed in the workplace. Compared to national rates, there was a significant increase in deaths due to all lymphatic and hematopoietic neoplasms (observed:expected = 15:6.6) as well as from leukemia (observed:expected = 9:2.66; seven of the observed cases were acute myelogenous leukemias, one was chronic myelogenous leukemia, and one was an unspecified myelogenous leukemia; see Table 5). There was a strongly positive trend in mortality due to leukemia with increasing cumulative exposure (SMRs of 109, 322, 1186, and 6637, with increasing exposure) but no pattern between exposure and latency period. In a nested, matched case-control analysis, the average duration of exposure was longer for cases than controls (8.7 versus 2.6 years). There were four deaths in the cohort due to multiple myeloma (compared to one expected). Three of these deaths occurred in the lowest exposure group and all had a minimum latency period of 20 years. Although the numbers of observed and expected cases of leukemia in this study were rather small, additional deaths due to this cause have occurred according to the most recent (to December 1987) follow-up of a portion of this cohort, which has not been published (Rinsky, 1991).
Structural and numerical chromosomal aberrations have also been consistently reported in lymphocytes of workers exposed to benzene. Metabolites of benzene have been demonstrated to disrupt microtubule assembly in vitro, and also cause aneuploidy and chromosomal non-disjunction in human lymphocytes. This may be significant in light of the fact that cytogenetic abnormalities involving the loss of all or part of chromosomes 5 and 7 have been associated with therapy-related myelodysplastic syndrome and leukemia (Irons et al., 1984; Lebeau et al., 1986).
| Case Number | Latency* (years) | Cause of Death** | Plant Location; Duration of Employment |
|---|---|---|---|
| 1 | 17 | Monocytic leukemia (204) | Location 1; 1.5 years |
| 2 | 2 | Chronic myelogenous leukemia (204) | Location 1; 1 month |
| 3 | 13.5 | Acute myelocytic leukemia (204) | Location 2; 11.5 years |
| 4 | 15.5 | Acute myelogenous leukemia (204) | Location 2; 14 years |
| 5 | 22 | Di Guglielmo's acute myelocytic leukemia (204) | Location 2; 13 years |
| 6 | 20 | Acute granulocytic leukemia (204) | Location 2; 20 years |
| 7 | 15 | Acute monocytic leukemia (204) | Location 2; 5 years |
| 8 | 3.5 | Myelogenous leukemia (204) | Location 1; 1.5 years |
| 9 | 37 | Acute myeloblastic leukemia (204) | Location 2; 14 years |
| * Latency was defined as the length of time from the date of first exposure until death. ** In parentheses is the International Classification of Disease code as determined by a nosologist from information on the death certificate. |
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Although it has been demonstrated that benzene crosses the placenta in humans, no effects on the fetus (with the exception of chromosomal abnormalities; Funes-Cravioto et al., 1977) and no increase in the incidence of birth defects have been associated with exposure to benzene in a few limited studies (Heath, 1983; Budnick et al., 1984; Olsen 1983a, 1983b; Axelsson et al., 1984). Although some reproductive effects have been reported in women in earlier limited studies, these observations have not been confirmed (Vara and Kinnunen, 1946; Michon, 1965; Pushkina et al., 1968; Mukhanetova and Vozovaya, 1972).
Neurotoxic effects similar to those caused by other petroleum hydrocarbons have been observed in workers exposed to benzene in combination with other industrial chemicals (Agency for Toxic Substances and Disease Registry, 1989).
2.8 Effects on the Environment
The information available on the acute and chronic toxicity of benzene includes data for species from a number of trophic levels, from bacteria and protozoa through to fish and amphibians in the aquatic environment. Information on toxicity to terrestrial species is limited to laboratory studies on plants, invertebrates, and mammals. No field studies on wild species were available.
Acute toxicity studies are available for several species at various trophic levels. The 3-hour EC50 for inhibition of photosynthesis in the alga Chlorella vulgaris was 312 mg/L (Hutchinson et al., 1980). The most sensitive freshwater invertebrates include nymphs of the damselfly, Ischnura elegans, with a 48-hour LC50 of 10 mg/L (Sloof, 1983), and the water fleas, with 48-hour LC50s of 15 mg/L for Daphnia pulex (Trucco et al., 1983) and 31.2 mg/L for Daphnia magna (Bobra et al., 1983). The most sensitive fish species tested were salmonids, including rainbow trout, Oncorhynchus mykiss, with a 96-hour LC50 of 5.3 mg/L for juveniles (DeGraeve et al., 1982), and coho salmon, Oncorhynchus kisutch, with a 96-hour LC50 of 9 mg/L for fry (Moles et al., 1979).
Benzene is toxic to a range of insects following topical or inhalation exposure; lethal effects were reported following exposure to air concentrations of 10 000 to 210 000 mg/m3 (Miller et al., 1976). Acute effects of benzene on terrestrial plants have been reported at atmospheric concentrations greater than 10 000 mg/m3 (Miller et al., 1976).
Black et al. (1982) investigated the chronic toxicity of benzene to the early life stages of rainbow trout, leopard frog (Rana pipiens), and northeastern salamander (Ambystoma gracile). Eggs of each species were exposed continuously to benzene from within 30 minutes of fertilization (embryos) on through to 4 days post-hatch (larvae). This resulted in continuous exposures of 27 days for rainbow trout, 9 days for leopard frog, and 9.5 days for northeastern salamander. The LC50s for continuous exposure were 8.3 mg/L for rainbow trout, 3.7 mg/L for leopard frog, and 5.2 mg/L for northeastern salamander.
Although no data were available on the effects of benzene on wild mammals, the toxicity of benzene to these organisms can be assessed by extrapolation from toxicity studies conducted using laboratory mammals (see Subsection 2.6). Benzene is not highly acutely toxic to mammals exposed by inhalation or ingestion (Agency for Toxic Substances and Disease Registry, 1989). A 4-hour inhalation LC50 of 44 500 mg/m3 has been reported for rats (Drew and Fouts, 1974), while a 7-hour LC50 for rats was 32 500 mg/m3 (NIOSH, 1987). The acute oral LD50s in the rat and mouse are 3306 and 4700 mg/kg b.w., respectively (NIOSH, 1987). As stated in Subsection 2.6, hematological effects have been reported in mice exposed in utero to 16 mg/m3, although these effects may not be strictly relevant to wildlife. Other responses noted in laboratory mammals include immunological effects in rats noted at 32 mg/m3, and neurological effects and behavioural disturbances at 320 mg/m3. No data are available on the effects of exposure to benzene on birds.
Gases involved in enhanced global warming strongly absorb infrared radiation, especially wavelengths of 7 to 13 µm, enabling them to trap and re-radiate a portion of the earth's thermal radiation (Wang et al., 1976; Ramanathan et al., 1985). Since benzene does not absorb at these wavelengths (Sadtler Research Laboratories, 1982), it is not considered to be a direct contributor to global warming. Substances involved in depletion of stratospheric ozone are generally halogenated, insoluble in water, and persistent in the atmosphere allowing movement to the stratosphere. In the stratosphere, they are degraded only by high energy, short wavelength ultraviolet radiation (Firor, 1989). Since benzene is a non-halogenated, water-soluble molecule of low persistence in the atmosphere, it is not associated with depletion of stratospheric ozone.
