General description

Benzene (CAS Registry Number 71-43-2; C₆H₆; molecular weight 78.1 g/mol) is an aromatic compound with a single six-member unsaturated carbon ring. It is a clear, colourless, volatile, highly flammable liquid with a characteristic odour and a density of 874 kg/m³ at 25 °C (1).

At 1 atmosphere of pressure, benzene has a melting point of 5.5 °C, a relatively low boiling point of 80.1 °C and a high vapour pressure (12.7 kPa at 25 °C), causing it to evaporate rapidly at room temperature. It is slightly soluble in water (1.78 g/l at 25 °C) and is miscible with most organic solvents (2). Benzene is soluble in lipids, has a log K octanol–water partition coefficient of 2.14 (1) and a log K soil organic carbon–water partition coefficient of 1.85 at 25 °C. Its Henry's Law constant is 550 Pa.m³/mol at 25 °C, implying that it will have a tendency to volatilize into the atmosphere from surface water (3).

Benzene in air exists predominantly in the vapour phase, with residence times varying between one day and two weeks, depending on the environment, the climate and the concentration of other pollutants. Reaction with hydroxyl radicals is the most important means of degradation, with a rate constant of 1.2 × 10⁻¹² cm³.molecule⁻¹.s⁻¹ at 298 K (4).

Other oxidants such as ozone and nitrate radicals can also contribute to a lesser extent to the degradation of benzene indoors, with rate constants of 2.7 × 10⁻¹⁷ cm³.molecule⁻¹.s⁻¹ at 298 K for nitrate radicals (5) and 1.7 × 10⁻²² cm³. molecule⁻¹.s⁻¹ at 298 K for ozone (1–3,6,7).

Indoor sources

Benzene in indoor air can originate from outdoor air and also from sources indoors such as building materials and furniture, attached garages, heating and cooking systems, stored solvents and various human activities. Indoor concentrations are also affected by climatic conditions and the air exchange rate due to forced or natural ventilation.

Indoor concentrations are affected by outdoor levels owing to the exchange of indoor and outdoor air. Outdoor benzene concentrations are mainly due to traffic sources and are affected by season and meteorology. Other outdoor sources of benzene are petrol stations and certain industries such as those concerned with coal, oil, natural gas, chemicals and steel (8).

Materials used in construction, remodelling and decorating are major contributors to indoor benzene concentrations (9). Certain furnishing materials and polymeric materials such as vinyl, PVC and rubber floorings, as well as nylon carpets and SBR-latex-backed carpets, may contain trace levels of benzene. Benzene is also present in particleboard furniture, plywood, fibreglass, flooring adhesives, paints, wood panelling, caulking and paint remover (3,10,11). Therefore, new buildings or recently redecorated indoor environments have been associated with high concentrations of benzene from materials and furniture. The rate of emission of benzene from materials and furniture will decay and eventually these sources will reach a quasi-steady emission rate in new buildings within weeks or months or up to a year (12).

Attached garages are a potential source of gasoline vapour owing to evaporation and exhaust emissions. In addition to cars, petrol, oil, paint, lacquer and hobby supplies often stored in garages can lead to increased levels of benzene indoors (13). Some 40–60% of benzene indoors may be attributable to the presence of an attached garage (13–16), with indoor benzene concentrations rising to 8 μg/m³ when garages are connected to the main living environment (14).

The use of fuels such as coal, wood, gas, kerosene or liquid petroleum gas (LPG) for space heating and cooking leads to higher concentrations of benzene indoors (17–20).

The problem of indoor pollution from the use of domestic cooking stoves attains greater importance in developing countries owing to poor ventilation and the extensive use of low-efficiency stoves and biofuels. Benzene concentrations of 44–167 μg/m³ have been found to be associated with the use of kerosene stoves (21).

In the past, benzene was widely used as a solvent, mainly in industrial paints, paint removers, adhesives, degreasing agents, denatured alcohol, rubber cements and arts and crafts supplies. The imposition of lower occupational exposure limits led to a reduction in these uses (3) but benzene content may still be an issue in some parts of the world, such as some African countries.

Indoor benzene is also associated with human activities such as cleaning (18), painting (18,22,23), the use of consumer products (24) and mosquito repellents (25), photocopying (26) and printing (27), the storage and use of solvents, and smoking tobacco.

Environmental tobacco smoke (ETS) is considered one of the main indoor sources of benzene. Benzene emissions from cigarette smoking range from 430 to 590 μg per cigarette (28). An increase in benzene concentration of at least 30–70% is expected (3,18,20,29,30) when ETS is present indoors, with increases in some cases of 300% (31) to levels of 16 μg/m³ (18).

To sum up, outdoor benzene provides the baseline for benzene concentrations indoors, upon which will be superimposed benzene given off from building materials and indoor artefacts. The presence of attached garages and combustion sources (especially smoking) and other human activities will be the main determinant of the concentration of benzene indoors.

Pathways of exposure

Inhalation accounts for more than 95–99% of the benzene exposure of the general population, whereas intake from food and water consumption is minimal (3,32). In the United States, daily benzene intake from ambient and indoor air has been calculated to range between 180 and 1300 μg/day, and that in food and water up to about 1.4 μg/day (2). The average daily intake for an adult in Canada was estimated to be 14 μg/day from ambient air, 140 μg/day from indoor air, 1.4 μg/day from food and drinking-water and 49 μg/day from car-related activities, giving a total of about 200 μg/day (33). Wallace (30) estimated the corresponding average intake in the United States to be 320 μg/day.

Cigarette smoking has been found to contribute significantly to the amount of benzene inhaled (34). Exposure to ETS is widespread in most countries (35). A survey conducted in the United States in 2006 found that more than 40% of nonsmoking adults and almost 60% of children aged 3–11 years were exposed to ETS (36). Another survey, conducted among young people in 132 countries, found that 44% had been exposed to ETS at home and 56% in public places, while another survey found that the exposure of young people at home ranged between 30–87% and 53–98% in public places (37). Active smoking may add as much as 400–1800 μg/day (2,34), while inhalation due to passive smoking will represent an additional 14–50 μg/day to the average daily intake (2,38). Driving a car during the rush hour may give a significant additional intake of 20 μg/day (34,39). Fromme (40) calculated the relative intake from food and uptakes from ambient air, indoor air and air inside cars to be 8%, 9%, 53% and 30%, respectively. In a study carried out in Germany in the 1990s, it was found that indoor exposure to ETS and car-related activities (refuelling and time in transit) could account for 20% and 12%, respectively, of personal exposure to benzene (2).

A study carried out in the United Kingdom estimated a daily dose of benzene of 70–75 μg/day for rural non-smokers and 89–95 μg/day for urban non-smokers. The daily dose rose to 116–122 μg/day for urban passive smokers and to over 500 μg/day for urban smokers. Children's daily exposures were estimated to be 15–20 μg/day and 30–40 μg/day for infants and children, respectively, while exposure to ETS led to a daily exposure of 26 μg/day and 59 μg/day for a urban infants and children, respectively (34). Most of the children's exposures were produced in the home (41).

A European study estimated a daily inhaled benzene dose of 102 μg/day, where 36%, 32%, 2% and 30% of the exposure was attributed to indoor home, indoor work, outdoor and in transit, respectively (42). In some Asian cities, where high levels of benzene were reported in homes and offices (25,43), the daily inhalation dose of benzene from indoor sources can be as high as 480–580 μg/day.

Indoor concentrations

Mean indoor concentrations are typically higher than the respective ambient levels and have been consistently shown to be higher in the colder than the warmer seasons (16,44,45). Indoor levels measured in the United States are in the range 2.6–5.8 μg/m³ (13,14,46,47), which are levels similar to those measured in established buildings in Australia (22) and Europe (48).

In European cities, a trend has been observed of increasing indoor concentrations from north to south. Low indoor concentrations (2 μg/m³) were measured in Finnish homes (49), while they ranged from 2 to12 μg/m³ in central European cities (17,44,50–54) and from 10 to 13 μg/m³ in southern cities such as Milan and Athens (48). Indoor levels measured in Turkey were in the range 7–14 μg/m³ (55).

Studies carried out in Asian cities have found much higher indoor benzene concentrations than those reported from cities in the developed world. Houses in India that used kerosene stoves were reported as having average indoor levels of 103 μg/m³ (21). Higher concentrations have been reported from some Chinese cities, with levels as high as 57.4 μg/m³ in Guanzhou (56). On the other hand, indoor levels of benzene in Japan are similar to those found in Australia, Europe and the United States, with arithmetic mean values ranging from 0.7 to 7.2 μg/m³ (45,57–59).

Indoor concentrations in buildings in Singapore were 18.4–35.4 μg/m³ (43), and similar levels of 23–35 μg/m³ were found in the Republic of Korea (25). However, a previous study in the Republic of Korea at the end of the1990s found lower concentrations (8.2 μg/m³ in homes and 12.6 μg/m³ in offices) than those reported in 2003 by Son et al. (60). Another study performed in India reported indoor concentrations of 10.7 μg/m³ (23). The lowest concentrations were reported from the Hong Kong Special Administrative Region of China (Hong Kong SAR), with values of 0.5–4.4 μg/m³ in different indoor environments such as houses, offices and shopping centres (61,62).

Cigarette smoke is an important source of benzene in indoor air, and benzene concentrations measured indoors increase when ETS is present (2). Indoor benzene levels measured in the United States showed arithmetic values of 5.54–10.5 μg/m³ in homes exposed to ETS compared to 3.86–7.0 μg/m³ in ETS-free homes (20,63). A similar situation was reported in Italy, with levels of 32.2 and 18.9 μg/m³ in ETS and ETS-free homes, respectively (64) and in Germany, with levels of 11.0 and 6.5 μg/m³, respectively (40).

Indoor concentrations measured in offices are generally higher than those measured in residential buildings, owing to the presence of sources such as photocopiers and printers. The mean office level in eight European countries was 14.6 μg/m³, while 87.1 μg/m³ was measured inside an office in Singapore (43). A recent study in United Kingdom offices reported lower benzene levels in the range of 0.4–4.0 μg/m³ (1.3 μg/m³ arithmetic mean) (53).

Benzene levels measured in restaurants ranged from 1.1 to 22.7 μg/m³, while higher levels of 5.1–78.8 μg/m³ were reported in pubs (18,53,60,61,65,66), with discotheques/clubs being the locations with the highest mean concentrations (193 μg/m³) in a study carried out in Germany (66). Benzene concentrations measured in several public indoor spaces such as shopping centres, libraries and cinemas ranged from 0.7 to 15.5 μg/m³ (18,53,62).

Benzene concentrations measured in vehicles are generally higher than those outdoors. Levels of benzene measured in vehicles in Europe ranged from 13 to 42 μg/m³ (65,67), while lower levels of 1.3–3.8 μg/m³ were measured in a recent United Kingdom study (53). Benzene levels measured in Mexico and the United States ranged from 1.7 to 42 μg/m³ (68,69) and a similar range (0.5–47 μg/m³) was found in several Asian cities (61,70). The highest in-vehicle benzene levels were measured in Italy in the early 2000s, with geometric means ranging from 17 to 101 μg/m³ (64).

Relatively high benzene concentrations indoors have been attributed to sources such as incense burning, with benzene concentrations peaking at up to 117 μg/m³ (48); new buildings (e.g. up to 30 μg/m³) (22); attached garages (e.g. 16–19 μg/m³); tobacco smoke (e.g. 16–193 μg/m³) (18,23,66); cleaning (e.g. 13 μg/m³) (18); painting (e.g. 9–13 000 μg/m³) (18,23) and using a kerosene stove (e.g. 166 μg/m³) (21).

Indoor–outdoor relationship

Indoor concentrations of benzene are normally higher than those in outdoor air (9) as a consequence of the entry and accumulation of benzene from outdoor sources and the presence of dominant benzene sources indoors. Viewed across published studies, indoor concentrations of benzene ranged from 0.6 to 3.4 (arithmetic mean 1.8) times the outdoor concentrations and are greatly influenced by those outdoors. This occurs in part because there are numerous indoor sources of benzene and because the relatively low rates of ventilation typically used in residences and offices prevent the rapid dispersal of airborne contaminants (9).

Indoor–outdoor ratios close to unity (i.e. 0.96–1.10) have been reported in some Asian countries where outdoor air concentrations were particularly high (25–35 μg/m³) (25,55,60,71). High indoor–outdoor ratios have been traditionally associated with strong indoor sources such as attached garages (ratio > 3) (13,14), combustion sources such as kerosene stoves (ratio 3.3) (21), gas and charcoal cooking (ratio 2) (60) or ETS (ratio 1.6–2) (23,48,60).

Kinetics and metabolism

The toxicity of benzene is dependent on its metabolism, as shown by its lower toxicity (a) in the presence of toluene, an inhibitor of benzene metabolism; (b) in animals that have had a partial hepatectomy; and (c) in mice that lack the enzyme CYP2E1 (72). Many studies have been completed in animals and to some extent in humans to determine the metabolism of benzene and its toxicokinetics.

Toxicokinetics

Absorption

Following inhalation exposure, the fraction absorbed is concentration-dependent, with a higher fraction absorbed at lower concentrations. In rats exposed for six hours to 11 or 130 ppm benzene, approximately 95% of the inhaled benzene was absorbed, while only 52% was absorbed after exposure to 930 ppm benzene (73).

Two studies in humans indicate that 50% of the quantity of inhaled benzene is absorbed (74,75). Cigarette smoke is a source of benzene exposure; the benzene concentration in the blood of 14 smokers was significantly higher (median 493 ng/l) than that in 13 non-smokers (median 190 ng/l) (76). Absorption of benzene is also rapid via the oral and dermal routes. Rats absorb and rapidly metabolize oral doses of benzene up to approximately 50 mg/kg. However, after an oral dose of 150 mg/kg, about 50% of the dose is exhaled as non-metabolized benzene (73).

Distribution

After entry into the human organism, benzene is distributed throughout the body and, owing to its lipophilic nature, accumulates preferentially in fat-rich tissues, especially fat and bone marrow. In humans, benzene crosses the blood–brain barrier and the placenta and can be found in the brain and umbilical cord blood in quantities greater than or equal to those present in maternal blood (77,78).

Elimination

Following all routes of exposure in rats and mice, absorbed benzene is rapidly metabolized (mostly within 48 hours), mainly by the liver, and approximately 90% of the metabolites are excreted in the urine (72). Elimination of non-metabolized benzene is by exhalation.

Metabolism

Qualitatively, the metabolism and elimination of benzene appear to be similar in humans and laboratory animals (79). Benzene is metabolized mainly in the liver but also in other tissues, such as the bone marrow. A diagram of benzene metabolism is presented in Fig. 1.1 (80).

Fig. 1.1

Metabolism of benzene. Source: Agency for Toxic Substances and Disease Registry (80). Note: ADH: alcohol dehydrogenase; ALDH: aldehyde dehydrogenase; CYP 2E1: cytochrome P450 2E1; DHDD: dihydrodiol dehydrogenase; EH: epoxide hydrolase; GSH: glutathione; (more...)

The first step consists in oxidation to benzene oxide and benzene oxepin (formation in equilibrium). This step is mainly catalysed by the enzyme CYP2E1 (81). There are three major pathways by which benzene oxide is further metabolized. It can go through a series of ring-breakage reactions to form t,t-muconaldehyde, which is further oxidized to the acid; it can go through a series of reactions to form a conjugate with glutathione, which is eventually excreted in the urine as phenyl mercapturic acid; or it can rearrange non-enzymatically to form phenol (82).

Phenol can be excreted in the urine directly or it can be further oxidized by CYP2E1 to catechol or hydroquinone. Catechol can be oxidized to trihydroxy-benzene, and hydroquinone can be oxidized to the highly reactive bipolar benzoquinone. All of the phenolic compounds can form conjugates (glucuronides or sulfates) and be excreted in the urine (72,83–86). The enzyme myeloperoxidase (MPO), which is present in bone marrow, can also oxidize phenolic compounds into quinones (79,87–89).

The metabolites responsible for benzene toxicity are not yet fully understood. The key toxic metabolites for cytotoxicity and the induction of leukaemia are thought to be benzoquinone, benzene oxide and muconaldehyde (1,90–95). The genotoxic activity of benzene metabolites is thought to be clastogenic (causing chromosomal damage) rather than acting through point mutations (see section on mechanism of action below). Benzoquinone and muconaldehyde are both reactive, bipolar compounds known to be clastogenic and the pathways leading to their formation are favoured at low concentrations in both mice and humans (72,96,97).

Two enzymes are active in the detoxification of benzene metabolites (98). One is NAD(P)H:quinone oxidoreductase (NQO1), which reduces the quinone metabolites to the less toxic diols (87,99); the other is the microsomal epoxide hydrolase, which hydrolyzes the epoxide group on benzene oxide.

There are species differences in the metabolism of benzene. Rats convert most of the benzene to phenol, a marker of a detoxification pathway, while mice form greater amounts of hydroquinone, hydroquinone glucuronide and muconic acid, all markers of toxification pathways. Human metabolism resembles that of mice, the species more sensitive to benzene toxicity (79,100–102).

Biomarkers of exposure

In the past, urinary phenol was commonly used as a biological exposure index in industrial settings to evaluate the exposure of workers to benzene. However, phenol is a good marker only of high-level benzene exposure and, with increased regulation of exposures, urinary phenol is no longer sensitive enough to be useful. More sensitive than phenol are urinary S-phenyl mercapturic acid and t,t-muconic acid, but the most sensitive exposure biomarker studied so far is the parent compound, benzene, in the urine (103,104).

Polymorphisms

Polymorphisms in genes involved in benzene metabolism are thought to influence individual susceptibility to various levels of benzene exposure. Lin et al. (105) concluded that, among the GST genotypes investigated, only the GSTT1 genotype was related to the level and dose-related production of S-phenyl mercapturic acid.

NQO1 also exists in polymorphic form. The wild NQO1*1 allele encodes the normal enzyme NQO1, whereas the NQO1*2 allele encodes a mutated NQO2 enzyme presenting negligible activity. Approximately 5% of the Caucasian and Afro-American population, 15% of the American–Mexican population and 20% of the Asian population are homozygotes for the NQO1*2 allele (106,107). Rothman et al. (108) demonstrated that workers in which the enzymatic activity of NQO1 was negligible presented a higher risk of benzene poisoning. The same is true for those expressing a rapid cytochrome CYP2E1 activity. Workers who simultaneously had a negligible NQO1 activity and a rapid CYP2E1 had a sevenfold higher risk of benzene poisoning than workers not presenting this dual polymorphism. Deletion of the glutathione S-transferase T1 (GSTT1) gene also showed a consistent quantitative 35–40% rise in DNA single strand break (DNA-SSB) levels.

Health effects

Carcinogenic effects

Genotoxicity

The genotoxic effect of benzene has been shown to be mainly clastogenic rather than the induction of point mutations. Numerous studies have demonstrated that benzene and its primary metabolites cause chromosomal aberrations (hypodiploidy and hyperdiploidy, deletion and breaking) in humans after chronic exposure (162–177). These chromosomal aberrations were observed in workers exposed to benzene concentrations high enough to induce dyscrasia. They are frequently localized in the peripheral blood lymphocytes and bone marrow. The main limitations of these studies lie in the lack of precise data concerning measurement of exposure, possible co-exposure to other substances and the absence of a suitable control group. Analysis of peripheral lymphocytes in workers exposed to benzene vapour (mean 30 ppm) revealed a significant increase in monosomy for chromosomes 5, 7 and 8, as well as an increase in trisomy or tetrasomy for chromosomes 1, 5, 7 and 8 (176,177).

A significant increase in hyperploidy for chromosomes 8 and 21, along with an increase in translocations between chromosomes 8 and 21, have been observed in workers exposed to benzene at a mean concentration of 31 ppm (100.75 mg/m³) (173). Kim et al. (178) showed around a twofold increase in micronuclei and chromosomal aberrations. In contrast, studies showed a decreased level of t(14;18) chromosome translocation in workers (179). Lichtman (180) did not find any chromosomal band damage.

The studies by Liu et al. (118) and Nilsson et al. (119) suggested that the metabolites of benzene activate oxygenated radical species, which can lead to DNA changes and the formation of hydroxylated bases such as 8-hydroxy-2-deoxyguanosine. Navasumrit et al. (181) showed a significant twofold increase of leukocyte 8-hydroxy-2-deoxyguanosine and DNA strand breaks in temple workers. Buthbumrung et al. (182) reported a similar result in schoolchildren exposed to benzene.

The genotoxic capacities of benzene are due to its metabolites. Pandey et al. (183) showed with the micronucleus assay that metabolites of benzene, especially p-benzoquinone, produce significant DNA damage. Keretetse et al. (184) showed DNA damage with the comet assay. Galván et al. (185) showed that the WRN gene protects against DNA damage. For the first time, Shen et al. (186) reported an association between benzene exposure and increased mitochondrial DNA copy number.

Carcinogenesis

Animals. Chronic exposure of both rats and mice to benzene leads to an increased incidence of tumours, though mice are more sensitive (100–102). The tumours formed include hepatomas, Zymbals gland tumours, lymphomas and tumours of the lung and ovary. However, there is no animal model for the induction of acute myeloid leukaemia, the major neoplastic lesion in humans. A study by Ross (88,99) in mice deficient in some detoxification enzymes showed that the genetically modified mice developed myeloid cell hyperplasia. Animal studies also showed that intermittent lifetime exposures to benzene at 980 mg/m³ were more tumorigenic than short-term high-level exposures at 3900 mg/m³ (187).

Humans. Epidemiological studies have clearly demonstrated a causal relationship between exposure to benzene or solvents containing benzene in the workplace and the development of acute myeloid leukaemia (123,124,188–191).

Rinsky et al. (190) studied a cohort of 1165 male workers employed in the Pliofilm¹ manufacturing industry between 1940 and 1965 up to 1981. The control data were the mortality data of American individuals of the same age as those studied in the cohort. An increase in mortality due to leukaemia was observed (9 cases observed instead of 2.7 expected), i.e. an SMR (standardized mortality ratio) of 3.37 (95% CI 1.54–6.41), along with an increase in mortality linked to multiple myeloma (4 cases observed, 1 case expected (SMR 4.09; 95% CI 1.10–10.47). The same evaluation repeated 15 years later reported a reduction in SMR for both leukaemia (SMR 2.56; 95% CI 1.43–4.22) and multiple myeloma (SMR 2.12; 95% CI 0.69–4.96) (189). A significant increase in leukaemia, including myeloid leukaemia but not multiple myeloma, was observed with an increase in cumulative exposure to benzene (200 ppm-years,² i.e. 650 mg/m³-years) (190,192). An analysis of 4417 workers did not clearly reveal an increased risk of acute non-lymphocytic leukaemia, multiple myeloma or other types of lymphohaematopoietic cancers, with a low cumulative exposure to benzene, i.e. between 1 and 72 ppm-years (3.25 and 234 mg/m³-years).

Kirkeleit et al. (193) performed a historical cohort study of workers employed in Norway's petroleum industry exposed to crude oil and other products containing benzene. Workers in the job category “upstream operator offshore”, having the most extensive contact with crude oil, had an excess risk of haematological neoplasms (blood and bone marrow) (rate ratio (RR) 1.90; 95% CI 1.19–3.02). This was ascribed to an increased risk of acute myeloid leukaemia (RR 2.89; 95% CI 1.25–6.67) (190). A peak exposure number of more than 100 ppm (325 mg/m³) benzene over 40 days or more therefore appears to be a better indicator of the risk of leukaemia and multiple myeloma than long-term exposure to benzene (194,195).

Within the most recently updated Pliofilm cohort, Paxton et al. (196,197) conducted an extended regression analysis with exposure description for the 15 leukaemia cases and 650 controls. They used all three exposure matrices, which gave estimates of 0.26–1.3 excess cancer cases among 1000 workers at a benzene exposure of 1 ppm (3.2 mg/m³) for 40 years.

A study resulting from collaboration between the National Cancer Institute (NCI) and the Chinese Academy of Preventive Medicine (CAPM) analysed different types of haematopoietic disease, malignant or otherwise (development of the disease and mortality rate linked to the disease), in a cohort of 74 828 workers exposed to benzene. A group of 35 805 workers not exposed to benzene were used as a control. All the workers included in the study came from 672 factories in 12 Chinese cities (189,198–201). The workers were employed from 1972 to 1987 for a mean duration of 12 years. A significant increase in the relative risk of haematological malignancies was observed (RR 2.6; 95% CI 1.4–4.7) as well as the risk for all leukaemias (RR 2.5; 95% CI 1.2–5.1), acute non-lymphocytic leukaemia (RR 3.0; 95% CI 1.0–8.9) and the combination of acute non-lymphocytic leukaemia and precursor myelodysplastic syndromes (RR 4.1; 95% CI 1.4–11.6) (189). Analysis of these risks as a function of atmospheric benzene concentrations (< 10 ppm, 10–24 ppm and ≥ 25 ppm) or cumulative exposure to benzene per year (< 40 ppm-years, 40–99 ppm-years and ≥100 ppm-years) indicated that the risk for all haematological malignancies was increased significantly at benzene concentrations of less than 10 ppm (32.5 mg/m³) and at cumulative benzene concentrations of less than 40 ppm-years (less than 130 mg/m³-years). The risk of acute non-lymphocytic leukaemia and the combination of acute non-lymphocytic leukaemia and precursor myelodysplastic syndromes was significant for a benzene concentration of between 10 and 24 ppm (32.5 and 78 mg/m³) and for cumulative exposures of between 40 and 99 ppm-years (130 and 322 mg/m³-years). Some criticisms may limit the utility of these data to develop a risk model. Limitations include the possibility of concurrent chemical exposures and a lack of reliable exposure data (124).

Analysis of these results on the basis of exposure duration (< 5 years, 5–9 years and ≥ 10 years) demonstrated that the risk does not increase with exposure duration, irrespective of the disease studied. Analysis on the basis of different factories and sectors demonstrated that the risks are similar irrespective of the factory's activity, suggesting that the risks calculated are indeed attributed to benzene and not to other pollutants that may be found in the factories. A study conducted in workers employed in a shoe-making factory in Italy demonstrated the same results as the Chinese study (202,203). The cohort was monitored from 1950 to 1999 and included 891 men and 796 women exposed to benzene concentrations of 0–92 ppm (0–299 mg/m³). The cumulative mean exposure was 71.8 ppm-years (233 mg/m³-years) for men and 43.4 ppm-years (141 mg/m³-years) for women. A significant increase in the risk of leukaemia was observed in both sexes for the highest benzene concentration among the four concentration categories. This increase was more apparent in men. For cumulative exposures divided into the following four categories: < 40 ppm-years, 40–99 ppm-years, 100–199 ppm-years and > 200 ppm-years (<130, 130–322, 325–647 and > 650 mg/m³-years), the SMR values for men were, respectively, 1.4 (95% CI 0.2–5), 3.7 (95% CI 0.1–20.6), 3.0 (95% CI 0.4–10.9) and 7.0 (95% CI 1.9–18.0). The type of leukaemia was not indicated.

A meta-analysis was conducted on 19 cohorts of workers in the petrochemical sector in the United Kingdom and the United States (204). The overall cohort included 208 741 workers. Mean exposures and mean cumulative exposures to benzene, for the most exposed posts, were 1 ppm and 45 ppm-years (3.25 mg/m³ and 233 mg/m³-years), respectively. No increase in mortality due to acute myeloid leukaemia, chronic myeloid leukaemia, acute lymphocytic leukaemia and chronic lymphocytic leukaemia was observed in this study.

Recently, Richardson (205) evaluated data from a cohort of 1845 rubber hydrochloride workers. He reported an association between leukaemia mortality and benzene exposure at greatest magnitude in the 10 years immediately after exposure: RR 1.19 (95% CI 1.1–1.29). The association was smaller in the period 10–20 years after exposure.

Recent data indicate that benzene exposure is haematotoxic at less than 1 ppm. A decrease in circulating lymphocytes has been observed in workers exposed for six months to a mean exposure concentration of less than 1 ppm (3.25 mg/m³) (143–145). For leukaemia, the studies by Hayes et al. (189,198,199) and Yin et al. (200,201) in a cohort of approximately 75 000 workers and 36 000 controls indicated that the risk for acute myeloid leukaemia and precursor myelodysplastic syndromes increased at between 10 and 24 ppm (32.5 and 78 mg/m³) and, for cumulative exposures, between 40 and 99 ppm-years (130 and 322 mg/m³-years).

The study by Rinsky et al. (190) described above demonstrates an increase in mortality related to the development of multiple myeloma in 1165 male workers followed up for one year. However, this result was not demonstrated in the other cohort studies (200,201,203,206,207). A follow-up analysis by Rinsky et al. (191) indicated an increased but non-significant risk of multiple myeloma, with no evidence of an exposure-response relationship. In addition, case-control studies conducted in hospital populations indicate that exposure to benzene was probably not related to an increased risk of developing multiple myeloma (208–213). Kirkeleit et al. (193) reported an increase in RR for multiple myeloma (RR 2.49; 95% CI 1.21–5.13) in workers exposed to crude oil and other products containing benzene employed in Norway's upstream petroleum industry.

The results of studies on non-Hodgkin's lymphoma appear to be less clear (191,194,195). In a meta-analysis of 25 occupational cohorts, no association of non-Hodgkin's lymphoma was found (207,214). A possible link between exposure to benzene and the development of non-Hodgkin's lymphoma was suggested by analysis of the results of the Chinese (NCI/CAPM) cohort described above (189). The relative risk of mortality linked to non-Hodgkin's lymphoma in the overall cohort was 3 (95% CI 0.9–10.5). This increase was not statistically significant. However, the risk of non-Hodgkin's lymphoma increased significantly at the highest benzene concentration and for the longest exposure duration. For mean exposure to benzene concentrations < 10 ppm, 10–24 ppm and ≥ 25 ppm, the relative risks for non-Hodgkin's lymphoma were, respectively, 2.7 (95% CI 0.7–10.6), 1.7 (95% CI 0.3–10.2) and 4.7 (95% CI 1.2–18.1) (P = 0.04). For cumulative benzene exposures of < 40, 40–99 and ≥ 100 ppm-years, the relative risks for non-Hodgkin's lymphoma were, respectively, 3.3 (95% CI 0.8–13.1), 1.1 (95% CI 0.1–11.1) and 3.5 (95% CI 0.9–13.2) (P = 0.02). In addition, the risk of developing non-Hodgkin's lymphoma increases significantly with an increase in benzene exposure duration. The relative risks are, respectively, 0.7 (95% CI 0.1–7.2), 3.3 (95% CI 0.7–14.7) and 4.2 (95% CI 1.1–15.9) for workers exposed for less than 5 years, for between 5 and 9 years and for more than 10 years (P = 0.01). The other cohort studies did not reveal any positive relationship between exposure to benzene and an increase in mortality due to non-Hodgkin's lymphoma (191,194,206). Kirkeleit et al. (193) reported no statistical differences between the groups in respect to non-Hodgkin's lymphoma.

Recently, Steinmaus et al. (215) conducted a meta-analysis of cohort and case-control studies of benzene exposure and non-Hodgkin's lymphoma and a meta-analysis of non-Hodgkin's lymphoma and refinery work. Results for the 22 studies indicated that the summary relative risk for non-Hodgkin's lymphoma was 1.22 (95% CI 1.02–1.47) (P = 0.01). When the authors excluded unexposed subjects in the “exposed group” (9 studies), the RR increased to 1.49. When studies based solely on self-reported work history were excluded (7 studies), the RR rose to 2.12 (95% CI 1.11–4.02). In refinery workers, the summary RR for non-Hodgkin's lymphoma in all 21 studies was 1.21 (95% CI 1.00–1.46) (P = 0.02). When adjusted for the healthy worker effect, this RR estimate increased to 1.42 (95% CI 1.19–1.69). These results suggest that effects of benzene on non-Hodgkin's lymphoma might be missed in occupational studies if these biases are not accounted for.

In addition, a recent review by IARC concluded that there is limited evidence of an association between benzene exposure and acute lymphocytic leukaemia or non-Hodgkin's lymphoma (216).

Table 1.1 collates studies on carcinogenic effects linked to human exposure to benzene, along with significant causal relationships between cancer and benzene exposure (subchronic and chronic).

Table 1.1

Review of SMR and RR values identified in the literature for chronic human exposure to benzene (occupational and environmental studies) for carcinogenic effects.

In conclusion, the different studies available (in humans, in animals and in vitro) have demonstrated that benzene metabolites trigger chromosomal aberrations (translocation, monosomy, trisomy). The carcinogenic mechanism of action of benzene is linked to its genotoxic effects and the critical health outcomes are blood dyscrasias and leukaemia, particularly acute myeloid leukaemia.

Health risk evaluation

Guidelines

Guidelines on exposure levels are needed for indoor air because indoor air is a significant source of benzene exposure and inhalation is the main pathway of human exposure to benzene. Benzene is present in both outdoor and indoor air. However, indoor concentrations are generally higher than concentrations in outdoor air owing to the infiltration of benzene present in outdoor air and to the existence of many other indoor sources. Typically, indoor concentrations are below the lowest levels showing evidence of adverse health effects. Considering benzene is present indoors and taking into account personal exposure patterns, which are predominantly indoors, indoor guidelines for exposure are needed.

Benzene is a genotoxic carcinogen in humans and no safe level of exposure can be recommended. The risk of toxicity from inhaled benzene would be the same whether the exposure were indoors or outdoors. Thus there is no reason that the guidelines for indoor air should differ from ambient air guidelines.

Previous WHO benzene guidelines for ambient air were calculated using the Pliofilm cohort studies (220). Since these studies, new data have become available, such as those on the large Chinese workers cohort (189). However, the unit risks and risk assessment analysis based on these data are still not available. Hence we recommend continuing to use the same unit risk factors calculated from the Pliofilm cohort studies. The geometric mean of the range of the estimates of the excess lifetime risk of leukaemia at an air concentration of 1 μg/m³ is 6 × 10⁻⁶. The concentrations of airborne benzene associated with an excess lifetime risk of 1/10 000, 1/100 000 and 1/1000 000 are 17, 1.7 and 0.17 μg/m³, respectively.

As noted above, there is no known exposure threshold for the risks of benzene exposure. Therefore, from a practical standpoint, it is expedient to reduce indoor exposure levels to as low as possible. This will require reducing or eliminating human activities that release benzene, such as smoking tobacco, using solvents for hobbies or cleaning, or using building materials that off-gas benzene. Adequate ventilation methods will depend on the site of the building. In modern buildings located near heavy traffic or other major outdoor sources of benzene, inlets for fresh air should be located at the least polluted side of the building.

The guidelines section was formulated and agreed by the working group meeting in November 2009.

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Footnotes

1

Pliofilm is a trade name. It is a plastic, derived from rubber, that is impermeable to water and used to package or store equipment or food, for example.

2

Cumulative benzene exposure over a one-year period.