General description

The term polycyclic organic matter (POM) defines a broad class of compounds that generally includes all organic structures containing three or more fused aromatic rings. These structures can contain the elements carbon, hydrogen, oxygen, nitrogen and sulfur.

POM containing up to seven fused rings has been identified, and theoretically millions of POM compounds could be formed; however, only about 100 species have been identified and studied. The most common subclass of POM is the polycyclic aromatic hydrocarbons (PAHs). These compounds contain only carbon and hydrogen (1).

PAHs are a large group of organic compounds with two or more fused aromatic (benzene) rings (2). Low-molecular-weight PAHs (two and three rings) occur in the atmosphere predominantly in the vapour phase, whereas multi-ringed PAHs (five rings or more) are largely bound to particles. Intermediate-molecular-weight PAHs (four rings) are partitioned between the vapour and particulate phases, depending on the atmospheric temperature (3). Particle-bound PAHs are considered to be very hazardous to human health. Benzo[a]pyrene (B[a]P) is often used as a marker for total exposure to carcinogenic PAHs, as the contribution of B[a]P to the total carcinogenic potential is high (in one study reported as being in the range 51–64%) (4).

B[a]P (CAS Registry Number, 50-32-8; C₂₀H_12; molecular weight = 252.31 g/mol) is a pale yellow monoclinic crystal with a faint aromatic odour. It has a melting point of 179 °C, a high boiling point of 496 °C at 1 atm, a Henry's Law constant of 4.8 × 10⁻⁵ kPa.m³/mol and a low vapour pressure of 7.3 × 10⁻⁷ Pa at 25 °C. As a consequence of these physical properties, B[a]P is predominantly particle phase rather than gas phase.

PAHs have a relatively low solubility in water (e.g. solubility in water of B[a]P at 25 °C is 3.8 μg/l) but are highly lipophilic (e.g. B[a]P log K_ow = 6.04)¹ and are soluble in most organic solvents. Once adsorbed on to soil, PAHs have low mobility (e.g. B[a]P log K_oc = 6.6–6.8).² Therefore, once released into the environment and owing to their low aqueous solubility, elevated octanol–water and organic carbon coefficients as well and high melting and boiling points, PAHs have a tendency to be associated with particulate matter, soils and sediments (2,5).

In the atmosphere, PAHs may be subject to direct photolysis, although adsorption to particulates can retard this process. PAHs can also react with pollutants such as ozone, hydroxyl radicals, nitrogen dioxide and sulfur dioxide, yielding diones, nitro- and dinitro-PAHs, and sulfonic acids, respectively (2). PAHs may also be degraded by some fungi and microorganisms in the soil and can be metabolized by a wide variety of terrestrial and aquatic organisms (7), although they are expected to bioconcentrate in organisms (aquatic and terrestrial) that cannot metabolize them (2,8).

Sources and pathways of exposure

Sources

PAHs are widespread environmental pollutants that are formed in the combustion process of carbonaceous materials at high temperature (10). Indoor air is contaminated by PAHs, which come not only from infiltration or intrusion of outdoor air but also from indoor emission sources such as smoking, cooking, domestic heating with fuel stoves and open fireplaces, as well as from incense and candle emissions (11–16).

For lower-molecular-weight PAHs, the impact of house characteristics and indoor activities tends to be greater than the influence of the penetrating outdoor air. On the other hand, while indoor sources may exist for PAHs with two or three rings, outdoor air may contribute significantly to the indoor PAHs, especially those with four or more rings (17).

Emissions from traffic have been found to be the main outdoor source for the indoor PAH concentration at urban and suburban locations in many industrialized countries (18). Motor vehicle emissions account for around 46–90% of the mass of individual PAHs in ambient air particles in urban areas (19), while domestic heating can account for some 16% of PAHs in outdoor air in the United States, 29% in Sweden and 33% in Poland, as reported in the early 1980s (20). Other outdoor sources of PAHs are industrial plants, power generation plants, waste incinerators and open burning. The age of a house or building, since it reflects its condition, affects PAH concentrations indoors. For example, the older a house the higher the PAH concentrations will be, as outdoor sources have a greater impact owing to higher air exchange through such routes as poorly fitting windows (21).

In industrialized countries, ETS appears to have the greatest impact on the total PAH concentration indoors and it is identified as the single largest source of PAHs in the indoor environment, with significant emission factors associated with smoking (22). Although reductions in the emission of PAHs in mainstream cigarette smoke have been reported (Table 6.1), the concentration of B[a]P in a room extremely polluted with cigarette smoke could still be as high as 22 ng/m³(23). In smokers' homes, more than 87% of the total PAHs may be attributable to this source. On the other hand, background sources are the largest contributor to PAHs in non-smokers' homes (24).

Table 6.1

Benzo[a]pyrene emission factors.

Cooking and heating with solid fuels such as dung, wood, agricultural residues or coal, especially in unvented or flueless stoves, is likely to be the largest source of indoor air pollution globally owing to the high level of use of these fuels in developing countries. More than 75% of people in China, India and nearby countries, and 50–75% of people in parts of Africa and South America, use solid fuel for cooking (25).

Concerning data reported on emission factors for B[a]P (Table 6.1) and PAHs (Table 6.2) from different fuels, these can be ranked as briquettes < wood < wood/ root-fuel mixtures according to their polluting potential, and natural gas < coal < briquettes < wood according to their B[a]P and PAH emission factors, respectively. However, caution should be exercised, as different studies report different ranges of compounds, which might not be comparable. Data on emission factors from burning candles show that this source emits less than cigarettes and fuels (16).

Table 6.2

PAH emission factors.

Wood burning in fireplaces and wood/solid fuel stoves is used as the main source of heating in developing countries and as a secondary heating source in countries with a cold winter climate. The burning of fossil fuel, solid fuel and biomass has been recognized as an important source of airborne PAHs as it releases a wide range of air pollutants, including PAHs, which are emitted to the indoor atmosphere in unvented or flueless combustion and also to the outdoor air (34). Even in airtight stoves with a flue, elevated indoor levels of PAHs can result from intrusion of outdoor air and/or leakage from wood-burning appliances (35).

High concentrations of particulate PAH compounds have been reported in indoor environments during the burning of fossil fuels and biofuel for cooking (36), generally in unvented stoves, suggesting that exposure during the cooking period is 2–10 times higher than ambient exposure (37). Concentrations of PAHs and B[a]P indoors, using different types of cooking fuel, increased in the order LPG < kerosene < coal < wood < dung cake/wood mixture < dung cake as reported in Tables 6.3 and 6.4, respectively. Transient high concentration peaks were reported in measurements performed during cooking (38).

Table 6.3

Indoor PAH concentrations associated with different sources.

Table 6.4

Indoor benzo[a]pyrene concentrations associated with different sources.

Apart from cooking fuel being a source of PAHs, generated particularly in unvented stoves, cooking practice (e.g. charring meat, deep frying) is another source of PAHs generated during cooking. The emissions from cooking practice depend greatly on the cooking method used, the fat content of the food and the quantity of food being cooked. Food with a higher fat content emits more PAHs than low-fat food (41). Also, an increase in cooking temperature generally increases the production of most PAHs (3) because there is an increase firstly in the evaporation of PAHs from heated oils into the air and secondly there is an increase in the PAHs generated by pyrolysis from partially cracked organic compounds in food and cooking oils (3,42). A comparative study of cooking practices showed that boiling produced the least PAHs, while broiling and frying produced most PAHs (43).

Varying amounts of PAHs are present in creosote, which has been traditionally used as a wood preservative in the foundations of buildings, in fences and in the manufacture of garden furniture and outdoor recreational facilities in parks. B[a]P levels of 58–749 μg/g were found in creosote-impregnated wood products (29).

The EU restricts creosote applications inside buildings (44) and Japan restricts the B[a]P content in creosote (45), but creosote-treated wood might be an indoor source in other parts of the world.

Finally, mothball storage is associated with significant levels of naphthalene (39,46), acenaphthalene, phenanthrene and fluorene indoors (46).

Routes of exposure

Humans are exposed to PAH through several routes, namely inhalation of air and re-suspended soil and dust, consumption of food and water, and dermal contact with soil and dust (65). All these sources are relevant to global human exposure. However, while soil contact generally occurs outdoors and food and water consumption is usually indoors, inhalation leads to exposure both indoors and outdoors. Yet people spent 80–93% of their time indoors, and hence indoor air would be the most relevant source contributing to the inhalation route (66).

Air

The potential doses of carcinogenic PAHs³ were estimated using the standard EPA recommendation for an individual's respiration rate (67) and applying this factor to the range of concentrations reported in the section of this chapter dealing with indoor levels and their relation to outdoor levels (page 301). The recommended value for the average inhalation rate of the general population is 11.3 m³/ day for women and 15.2 m³/day for men (67). Considering the different B[a]P indoor air concentrations reported, and using the adult male inhalation rate as a worst-case scenario, the daily intake dose due to inhalation spans the range of 0.15–32 ng/day. However, higher daily levels of inhaled B[a]P can be experienced during exposure to specific indoor sources such as cooking with different fuels (91–2523 ng/day) or using non-airtight stoves for heating (30–7448 ng/day) (36) (Table 6.5).

Table 6.5

Benzo[a]pyrene inhalation daily dose.

ETS is an important contributor to the inhalation source of PAHs. Using the same methodology as describe above,⁴ daily inhalation of B[a]P in indoor environments would range from 4 to 15 ng/day in ETS-polluted compared with 1.3–6.7 ng/day in homes not exposed to ETS. The daily (24-hour) inhalation can be as high as 26–62 ng/day in pubs and discotheques. Children's daily exposures, expressed as urinary cotinine levels (a biomarker of tobacco smoke) were 8.1 μg/l urine in ETS-exposed children compared to 2.7 μg/l in children not exposed to ETS (68).

Drinking-water

Several studies performed in the United States reported values of carcinogenic PAHs for drinking-water in the range 0.1–61.6 ng/l, although most of the values fell between 1 and 10 ng/l. In the case of B[a] P, all the values were below the limit of detection (0.1 ng/l) (65,69). Similarly, the examination of a number of drinking-water supplies for six PAHs (fluoranthene, benzo[b]fluoranthene, benzo[k]fluoranthene, B[a] P, benzo[g,h,i]perylene and indeno[1,2,3-c,d]pyrene) indicated that the collective concentrations generally did not exceed 100 ng/l. The concentrations of these six PAHs were between 1 and 10 ng/l in 90% of the samples and higher than 110 ng/l in 1% (23,70).

As regards the concentrations of the 16 PAHs, these span the range of 106.5– 150.3 ng/l in several European and Canadian cities (71), while lower values of 85.2–94.6 ng/l have been reported in Taiwan, China (72). Studies performed in Europe have reported levels of B[a]P in the range of < 1 ng/l in Germany (73) to 10 ng/l in Poland (74). Values of B[a]P in the same range (1.4–2.5 ng/l) have also been reported in Taiwan, China (72).

Assuming an average drinking-water consumption of 2 l/day, the potential dose of carcinogenic PAHs via drinking-water ranged from 0.2 to 123 ng/day, 170–300 ng/day for the 16 PAHs and < 2–20 ng/day for B[a]P.

Food

PAHs are found in substantial quantities in some foods, depending on the method of cooking, preservation and storage, and intake is influenced by personal eating habits (75). PAHs are detected in a wide range of meats, fish, vegetables and fruits, fluoranthene and B[a]P being the two PAHs detected at highest levels in food with fluoranthene levels exceeding those of B[a]P (76,77). Food groups that tend to have the highest levels of PAHs and B[a]P include charcoal-broiled or smoked meats, fats and oils, and some leafy vegetables and grains. For these food groups, concentrations of 16 PAHs were typically in the tens of micrograms per kilogram (Table 6.6) (78–81). However, the PAH load on leafy vegetables and grains can be removed by washing. As regards B[a] P, recent studies report that food containing fat show the highest levels of B[a] P, with maximum levels of 60 μg/kg (Table 6.7) (65,75,82). Lower levels of B[a]P in the range of hundreds of nanograms per kilogram have been reported in more recent studies for fruits and vegetables, sweets, dairy products, beverages, bread, cereals, grains and seafood (83,84).

Table 6.6

The 16-PAH content of foods.

Table 6.7

Benzo[a]pyrene content of foods.

A Dutch “market basket” study of dietary components for 18-year-old males, involving the determination of 17 different PAHs,⁵ revealed that all of these compounds were detected. The most frequently occurring were benzo[b] fluoranthene, fluoranthene and benzo[k]fluoranthene in 59%, 48% and 46% of the samples, respectively. The highest concentration of a single PAH was found for chrysene, at 36 μg/kg in the commodity group “sugar and sweets”. The mean daily intake of the total PAH fraction (17 PAHs) analysed ranged between 5 and 17 μg/day. The intake of the carcinogenic PAH fraction was roughly half of these amounts. The largest contribution to the daily PAH intake came from sugar and sweets, cereals, oils, fats and nuts (85).

For the average American diet, the intake of carcinogenic PAHs was estimated to be 1–5 μg/day, with unprocessed grains and cooked meats the greatest sources of the compounds (65). This is lower than in a recent study in Spain, where the dietary intake of carcinogenic PAHs⁶ ranged from 723 to 969 ng/day and the 16 PAHs ranged from 8.57 to 13.81 μg/day (78,86).

The dietary intake of B[a]P ranged between 0.002 and 1.1 μg/day in the United States in the late 1980s (69). However, lower levels were reported in a recent study, similar to those reported in Asia and Europe, ranging from 4.2 to 320 ng/day (Table 6.8). The lowest daily intake for B[a]P and 16 PAHs has been reported in Yemen (1.7 and 167 ng/day, respectively) based on the fish consumption of the Yemeni population (81).

Table 6.8

Benzo[a]pyrene daily dietary intake dose.

Soil

Carcinogenic PAHs are found in all surface soils (65). Typical concentrations in forest soil range from 5 to 100 μg/kg (Table 6.9). Substantial amounts of PAHs are transferred to forest soil from vegetative litter because the compounds are adsorbed from air onto organic matter such as leaves and pine needles. Rural soil contains carcinogenic PAHs at levels of 10–100 μg/kg, originating mainly from atmospheric fallout. For both forest and rural soil, values as high as 1000 μg/kg may occasionally be found (65,91,92).

Table 6.9

PAH content of soils.

Metropolitan areas have higher PAH concentrations than forest and agricultural areas because of the many sources of fossil fuel combustion. The majority of urban soil concentrations fall in the 600–3000-μg/kg range (65,93,94). Higher values near areas of heavy transportation and industrialization range from 8 to 336 mg/kg (65,95). Values in the order of 1000–3000 μg/kg are regarded as being in the upper range.

As regards B[a]P levels in topsoil (Table 6.10), the lowest concentrations are found in tropical rural and urban soils (0.3–5.5 μg/kg) and the highest in arable and forest areas in temperate latitudes (18–39 μg/kg) (96). The highest concentrations were found in urban areas, with values ranging from 5.5 to 379 μg/kg (96–100) and from 971 to 1600 μg/kg in large United Kingdom cities and Chicago (97). Levels in industrialized areas across the world ranged between 18 and 360 μg/kg (99,101,102).

Table 6.10

Benzo[a]pyrene content of soils.

Incidental ingestion of soil by adult males was estimated to be of the order of a few milligrams per day. Soil ingestion rates of the order of 100 mg/day are more typical for small children (103). Therefore, the potential dose of carcinogenic PAHs for urban populations ranged from 0.2 to 96 ng/day (median 7 ng/day).

Toxicokinetics and metabolism

The kinetics and metabolism of PAH(s) have been addressed previously in several WHO documents (2,8,70). The emphasis below is on aspects relevant particularly to exposure in indoor air. Moreover, the most recent data on PAHs are reviewed.

Health effects

Human health effects

While the risks of occupational exposure to PAHs are not the focus of this chapter, indoor exposure to smoke from biomass and coal burning for the population in developing countries could be comparable to the levels of pollutants in the occupational setting (234). For example, indoor particulate matter (< 10 μm) levels in solid-fuel-burning households reach up to several milligrams per cubic metre. An estimated half of the global population depends on solid fuel for cooking and heating, often in inadequately ventilated spaces (234). Women and children are particularly vulnerable because of the longer time spent at home. Timing of exposure in children could influence disease risk owing to the sensitive window of development, as well as exposure levels that are often higher relative to their body size. Indoor smoke exposure in such settings remains one of the top ten risks in the global burden of disease (235).

Identification of studies

PubMed was searched in English, the search being restricting to human studies. For non-carcinogenic effects, the following search terms were used: “polycyclic aromatic hydrocarbons AND indoor AND birth weight”, “polycyclic aromatic hydrocarbons AND smoke”, “polycyclic aromatic hydrocarbons AND biomass”, “polycyclic aromatic hydrocarbons AND coal burning”, “polycyclic aromatic hydrocarbons AND indoor AND intrauterine growth restriction”, “polycyclic aromatic hydrocarbons AND indoor AND low birth weight”, “polycyclic aromatic hydrocarbons AND indoor AND small-for-gestational age”, “polycyclic aromatic hydrocarbons AND indoor AND birth length”, “polycyclic aromatic hydrocarbons AND indoor AND birth head circumference”, “polycyclic aromatic hydrocarbons AND indoor AND fetal growth”, “polycyclic aromatic hydrocarbons AND indoor AND neurodevelopment”, “polycyclic aromatic hydrocarbons AND indoor AND PAH-DNA adducts”, “polycyclic aromatic hydrocarbons AND indoor AND bronchitis” and “polycyclic aromatic hydrocarbons AND indoor AND asthma”.

For carcinogenic risk, the following search terms were used: “polycyclic aromatic hydrocarbons AND 1-hydroxypyrene”, “polycyclic aromatic hydrocarbons AND PAH-DNA adducts”, “polycyclic aromatic hydrocarbons AND DNA”, “polycyclic aromatic hydrocarbons AND chromosom*”, “polycyclic aromatic hydrocarbons AND cancer”, “polycyclic aromatic hydrocarbons AND occupational”, “polycyclic aromatic hydrocarbons AND ischaemic heart disease”, “polycyclic aromatic hydrocarbons AND cognitive” and “polycyclic aromatic hydrocarbons AND neurodevelopment*”.

The search identified 455 papers. Moderate to large population-based prospective cohort studies using a quantitative assessment of PAH exposure were given first priority. Studies that did not adjust for known and potential confounders were not considered. Clinical trials, risk assessments, reviews of the literature, future studies, case reports, diagnostic guidelines and studies that lacked quantitative assessment of exposure were also excluded from the review. Subsequently, 178 papers were chosen for full review and 56 papers are included in the present report. In addition, earlier reviews by WHO (2,8), IARC (70,92,260) and other authors (130) were considered.

Non-carcinogenic effects

Intrauterine growth restriction. Intrauterine growth restriction has been operationalized as low birth weight (< 2500 g), low birth weight at full term and small for gestational age (SGA), defined as < 10th percentile of population mean weight at a given gestational age and gender.

Prenatal exposure to PAHs has been associated with reduction in birth weight, an increased likelihood of low birth weight in Europe and the United States (236) and SGA (237,238) in a dose-responsive manner, after controlling for region-and cohort-specific sets of confounders. In Teplice and Prague (Czech Republic), PAHs isolated from respirable particulate matter during winter induced the highest genotoxicity and embryotoxicity (239). High ambient concentrations of PAHs, PM₁₀ and PM_2.5 during the first month of gestation were associated with a significantly elevated risk of SGA in the industrial city of Teplice (237). A study in Poland showed that neonates with high levels of PAH-DNA adducts in the leucocytes had significantly lower birth weight, length and head circumference (240).

Two parallel prospective cohort studies in Krakow and New York City enrolled non-smoking pregnant women with no known risks of adverse birth outcomes and monitored their personal PAH exposure concentrations (241). Mean personal PAH exposures to eight carcinogenic PAHs¹² differed more than 10-fold between the two cities (Krakow mean 39.0 ng/m³, range 1.8–272.2 ng/m³; New York mean 3.3 ng/m³, range 0.3–36.5 ng/m³). In the Krakow cohort, prenatal exposure to the summed eight carcinogenic PAHs was significantly associated with reduced birth weight (68.75 g),¹³ birth length (0.48 cm) and birth head circumference (0.21 cm) (241). In the New York cohort, however, prenatal exposure to PAHs was associated with reduced birth weight (177.57 g) among New York African Americans but not among New York Dominicans (241). Furthermore, a natural log-unit increase in prenatal PAH exposure was associated with a two-fold increase in the likelihood of being born SGA among African Americans but not Dominicans in New York (238). SGA is one of the most clinically predictive markers of fetal growth impairment. SGA has been associated with a significantly greater risk of delayed neurodevelopment (242,243), shorter stature, cardiovascular disease, insulin resistance and diabetes during adulthood (244,245). Among the New York African Americans, a natural log-unit of prenatal exposure to PAHs¹⁴ was associated with a five-fold greater risk of preterm delivery (< 37 weeks at birth) (238). While residual confounding remains a possible alternative explanation, African American neonates and children appear to be more vulnerable.

Bronchitis, asthma and asthma-like symptoms. In the same New York City cohort, prenatal exposure to the particle-bound PAHs may increase the risk of asthma symptoms by the age of 1–2 years (246). In Teplice and Prachatice in the Czech Republic, unit exposure to ambient 100 ng/m³ PAHs, based on a 30-day average, and a unit exposure to 25 μg/m³ PM_2.5 were respectively associated with 56% (95% CI 22–100) and 23% (95% CI −6 to −62) increases in the risk of bronchitis in children between the ages of 2 and 4½ years (247).

Fatal ischemic heart disease. A multinational cohort of male asphalt workers was followed for fatal ischemic heart disease (IHD) for 17 years (SD = 9 years). Mean personal exposure to B[a]P for the cohort was 273 ng/m³. For those exposed to ≥ 273 ng/m³ B[a]P, the risk of IHD mortality was 1.64-fold greater (95% CI 1.13– 2.38) than in those exposed to ≤ 68 ng/m³(248). The risk increased in a dose-responsive manner. At the highest PAH exposure category, cigarette smoking by the workers did not explain the significant increase in IHD mortality risk, thus supporting the etiological role of B[a] P.

Neurodevelopmental index

In the same New York City birth cohort, those neonates with a higher than median prenatal PAH exposure (range 4.16–36.47 ng/m³) had a significantly lower Bayley Mental Development Index as well as a greater likelihood of cognitive developmental delay at the age of three years compared to children exposed to 0.27–4.15 ng/m³, controlling for ethnic background, gender, gestational age, level of nurturing provided at home, ETS and chlorpyrifos exposure (249).

In a cross-sectional investigation in Tongliang, China, babies born close to a coal-fired power plant in 2002 were associated with 0.32 ± 0.14 B[a]P DNA adducts per 10⁸ nucleotides (250). Such a level of B[a]P adducts at birth was associated with a decreased motor development quotient at two years of age (250). A 0.1-unit increase in B[a]P DNA adducts per 10⁸ nucleotides measured at birth was associated with two-times greater likelihood of developmental delay in motor dimension at the age of two years (250). However, following the shut-down of the Tongliang power plant in 2004, lower (0.20 per 10⁸ nucleotides) B[a]P adduct levels in cord blood were observed in a new cohort of children (251). Subsequently, in 2005, cord blood B[a]P adducts were not associated with a reduction in developmental status at two years of age (251).

Human carcinogenic risk

Early biological effects of PAH exposures have been examined in children in a small number of studies. A group of rural Indian children had significantly higher blood PAH concentrations (125.55 ± 26.99 ppb) than those from an urban region (23.96 ± 13.46 ppb), consistent with higher home exposure to the burning of wood, coal, cow dung and kerosene (252). The extent of lipid peroxidation in whole blood was positively, albeit weakly, correlated with their total PAH and carcinogenic PAH levels in whole blood (252). Among school-age children in Bangkok, the major source of exposure was road traffic.

When a group of Bangkok children were compared to a group of rural children, those from Bangkok had 3.5-fold higher exposure to B[a]P-equivalent compounds (mean 1.50 ± 0.12 vs 0.43 ± 0.05 ng/m³). The Bangkok children's corresponding urinary 1-hydroxypyrene levels were significantly higher than those of the rural children. Furthermore, the levels of lymphocyte PAH-DNA adducts in the Bangkok children were 5-fold higher than those for the rural schoolchildren (0.45 ± 0.03 vs 0.09 ± 0.00 adducts/10⁸ nucleotides). The frequency of DNA strand breaks was significantly higher, while the DNA repair capacity was significantly impaired in the Bangkok children compared to the rural children (253).

In Prague, the mean personal exposure to 1.6 ng/m³ of B[a]P during winter was associated with an elevated level of PAH-DNA adducts in a group of policeman (254). Compared to the non-smoking controls (mean B[a]P = 0.8 ng/m³), the frequency of chromosomal translocations for the policemen was significantly higher, based on fluorescence in situ hybridization (FISH) (255). Within each individual subject, the level of “B[a]P-like” DNA adducts was a significant predictor of genomic frequency of translocations as detected in terms of FISH, percentage of aberrant cells and aberrations per cell (255). In other groups of city policemen in Prague, DNA adducts were also positively correlated with genomic frequency of translocations (254,256). Corresponding personal exposure to B[a]P was 1.6 ng/m³ vs 0.4 ng/m³ (in controls) (254,256). Personal B[a]P exposure > 1.0 ng/m³ also increased the micronuclei measured by automated image analysis (257) as well as DNA fragmentation in sperm (268) in the group of city policemen. All these results from biomonitoring studies indicate that exposure to B[a]P concentrations > 1.0 ng/m³ induce DNA damage.

Carcinogenic effects of exposure to PAHs

In 2005, an IARC working group evaluated a number of occupational exposures to PAH-containing complex mixtures for their potential to induce cancer in humans. Occupational exposures during chimney sweeping, aluminium production, coal gasification, coke and steel production, coal-tar distillation, and paving and roofing with coal-tar pitch have been classified as carcinogenic to humans (Group 1). Occupational exposures to PAH-containing complex mixtures have been strongly associated with lung cancer and some of these exposures have a highly suggestive association with bladder and urinary tract tumours (259). In addition, B[a]P was upgraded to a human carcinogen, based on sufficient evidence of carcinogenic activity in animals and strong evidence that the mechanisms of carcinogenicity in animals operate in humans (259). The working group also updated a list of probable carcinogens to humans (Group 2A) and possible carcinogens to humans (Group 2B), as shown in Table 6.11. Cigarette smoke contains PAHs and cigarette smoking has also been classified as carcinogenic to humans (260).

Table 6.11

IARC Classification of agents and occupations.

Mutagenic and carcinogenic risk of PAH-DNA adducts in adults

PAH-DNA adduct formation represents one of the key first steps in carcinogenesis (261). PAH/aromatic DNA adducts have been associated with increased lung cancer risk in some molecular epidemiological studies. In a nested case-control study, healthy volunteers with detectable adduct levels in their leukocytes at the outset of the study were at two-fold greater risk of lung cancer than those with non-detectable levels (262). In particular, never-smokers with a detectable adduct level were at four-fold higher lung cancer risk than those with a non-detectable level (262). When the exposure was dichotomized at the median (0.6 DNA adducts per 10⁸ nucleotides), the non-smokers with greater than median adduct levels were at seven-fold greater risk of lung cancer than those with non-detectable adducts (262). In a small cross-sectional case-control study, those within the highest quartile of leukocyte PAH-DNA adduct levels (> 1.52 PAH-DNA adducts per 10⁸ nucleotides) were at three-fold greater risk of colorectal adenoma than the lowest quartile of the adducts (≤ 0.71 adducts per 10⁸ nucleotides) (263).

In population-based case-control studies, the detectable level of PAH-DNA adducts has been associated with a 29–35% elevation in breast cancer risk (261). Several genetic variants have been examined for their role in cancer development. A variant allele (GA or AA) in the FAS1377 gene was associated with a 36% increase in breast cancer risk among those with detectable PAH-DNA adduct levels (264). However, variants of other genes, including GSTA1, GSTM1, GSTP1 and GSTT1, did not modify the risk (265), while any dose–response relationship was consistently absent in several molecular epidemiological investigations (261,266,267). In a large population-based cohort, detectable PAH-DNA adducts, measured at the time of the patient's breast cancer diagnosis, were not associated with subsequent all-cause or breast-cancer-specific mortality (268). In a prospective follow-up study of prostate cancer, the risk of biochemical recurrence one year after surgical removal of the tumour was twofold greater in those with higher than the median level of prostate-specific adducts (269).

Quantitative estimates of carcinogenic effects at occupational exposure range

In a combined meta-analysis of the aluminium production, coal gasification, coke production, iron and steel, coal tar, carbon black and carbon electrode production industries, cumulative B[a]P exposure concentrations over the working years ranged from 0.75 to 805 μg/m³-years, equivalent to a concentration range of 0.04–40 μg/m³(270). The mean relative risk of lung cancer increased by between 20% and 168% per 100 μg/m³-years (270,271). The mean risk remained robust when study design, smoking status and exposure measurement were accounted for (270). While risk sizes were consistent within each occupation, large variability in lung cancer risk was observed across the occupations. Such heterogeneity in relative risk was attributed to variability in occupation-related exposure range.

Another meta-analysis estimated the industry-specific relative risk of respiratory and urinary tract cancers (272). Risk of lung cancer was significantly elevated for all examined industries (272). In particular, within the aluminium smelting industry, 100 μg/m³-years exposure to B[a]P (equivalent to 3.3 μg/m³ for 30 years) was associated with a 2.68-fold greater likelihood of developing lung cancer (271). While a significant departure from linearity was observed in this analysis, indicating that the unit risk might be smaller at the highest exposure range, this pooled estimate sufficiently accounted for confounding by smoking (271).

The risk of bladder cancer was significantly elevated only for those involved in aluminium production.

Conclusions

• Sources of airborne PAHs indoors are infiltration by PAHs in outdoor air and indoor emissions from smoking, domestic cooking and heating with fuel-burning stoves, open fireplaces, and incense and candle burning.
• In the absence of indoor sources, indoor concentrations of PAHs are lower than those outdoors.
• When indoor sources are present, indoor concentrations are likely to exceed those outdoors.
• In industrialized countries, ETS appears to have the greatest impact on indoor PAH concentrations, while in developing countries it is cooking and heating with solid and biomass fuels.
• In industrialized countries, inhalation is a minor route of exposure for non-smokers compared with dietary ingestion.
• In developing countries, inhalation is at least as important a route of exposure as dietary exposure.
• A sufficient body of evidence supports the causal role of PAH/aromatic DNA-adducts in lung cancer occurrence among non-smokers.
• B[a]P at levels above 1.0 ng/m³ predicted greater genomic frequency of translocations, micronuclei and DNA fragmentation.
• Prenatal exposure to PAHs is associated with an increase in the likelihood of low birth weight.
• B[a]P and many other PAHs induce cancer by a mutagenic mechanism that involves metabolic activation to reactive diol-epoxides that covalently bind to DNA. These PAH-DNA adducts have been detected in tissues from experimental animals exposed to PAHs, and B[a]P-DNA adducts have been found in human lung tissues.
• PAH-DNA adducts are converted into mutations after cell replication, and mutations in critical tumour oncogenes and tumour suppressor genes have been identified in lung tumours from both experimental animals and humans exposed to PAHs or PAH-containing mixtures.
• Sufficient evidence exists of a link between the prenatal exposure to mixtures of carcinogenic PAHs and intrauterine growth restriction in humans.
• There is a robust body of evidence supporting a strong association between occupational exposure to PAH-containing mixtures and lung cancer in humans.

Health risk evaluation

Guidelines

Some PAHs are potent carcinogens and, in air, are typically attached to particles. The primary exposure to carcinogenic PAHs found in air occurs via inhalation of particles. PAHs occur in indoor air as complex mixtures, the composition of which may vary from site to site. Experimental data on metabolism, gene expression and DNA adducts suggest that interactions between PAHs in mixtures may be complex and highly unpredictable for various PAH compositions (inhibitory, additive, synergistic).

In view of the difficulties in developing guidelines for PAH mixtures, B[a]P was considered to represent the best single indicator compound. Its toxicology is best known, most single PAH concentration data in ambient and indoor air are for B[a] P, and B[a]P has widely been used as an indicator compound for exposure in epidemiological studies.

The health evaluation data suggest that lung cancer is the most serious health risk from exposure to PAHs in indoor air. B[a]P is one of the most potent carcinogens among the known PAHs.

In its evaluation of PAHs as ambient air pollutants in 2000, WHO (2) expressed a unit cancer risk as a function of the concentration of B[a]P taken as a marker of the PAH mixture. Use of the same unit risk factor for indoor air implies that B[a]P represents the same proportion of carcinogenic activity of the PAH mixture as in the occupational exposure used to derive the unit risk. This assumption will not always hold, but the associated uncertainties in risk estimates are unlikely to be large.

Reducing exposure to B[a]P may also decrease the risk of other adverse health effects associated with PAHs.

Based on epidemiological data from studies on coke-oven workers, a unit risk for lung cancer for PAH mixtures is estimated to be 8.7 × 10⁻⁵ per ng/m³ of B[a] P. This is the guideline for PAH in indoor air. The corresponding concentrations for lifetime exposure to B[a]P producing excess lifetime cancer risks of 1/10 000, 1/100 000 and 1/1 000 000 are approximately 1.2, 0.12 and 0.012 ng/m³, respectively.

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

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Footnotes

1

The octanol–water partition coefficient (K_ow) is a measure of the hydrophobicity of a compound. It is a measure of the distribution of a compound between water and an organic (octanol) with which is in contact (6).

2

The organic carbon coefficient (K_oc) is a measure of a chemical compound's mobility in soil and the prevalence of leaching from soil (6).

3

Carcinogenic PAHs include benz[a]anthracene, benzo[a]pyrene, benzo[b]fluoranthene, benzo[k]fluoranthene, chrysene, dibenzo[a,h]anthracene, benzo[ghi]perylene and indeno[1,2,3-cd]pyrene.

4

Calculated using the concentrations reported by Mitra & Ray 1995 and Fromme et al. 1995 and applying the USEPA 1997 male individual's breathing rate.

5

The 17 PAHs comprise naphthalene, acenaphthylene, acenaphthene, fluorene, phenanthrene, anthracene, fluoranthene, pyrene, benz[a]anthracene, chrysene, benzo[b]fluoranthene, benzo[j]fluoranthene benzo[k]fluoranthene, benzo[a]pyrene, indeno[1,2,3-cd]pyrene, dibenz[a,h])anthracene, benzo[ghi] perylene.

6

Carcinogenic PAHs from this study include benz[a]anthracene, benzo[a]pyrene, benzo[b]fluoranthene, benzo[k]fluoranthene, chrysene, dibenzo[a,h]anthracene and indeno[1,2,3-cd]pyrene.

7

Calculated using the concentrations reported by Traynor et al. (49) and applying the USEPA (67) male individual's breathing rate.

8

Calculated using the concentrations reported by Zhu & Wang (41) and applying the USEPA (67) male individual's breathing rate.

9

Calculated using the concentrations reported by Raiyani et al. (36,38) and applying the USEPA (67) male individual's breathing rate.

10

The 12 PAHs comprised naphthalene, acenaphthene, fluorene, phenanthrene, anthracene, fluoranthene, pyrene, benz[a]anthracene, chrysene, benzo[k]fluoranthene, benzo[e]pyrene and benzo[a] pyrene.

11

The 12 PAHs comprised naphthalene, acenaphthylene, acenaphthene, fluorene, phenanthrene, anthracene, fluoranthene, pyrene, benz[a]anthracene, chrysene, benzo[k]fluoranthene and benzo[a]pyrene.

12

Sum of benz[a]anthracene, chrysene/isochrysene, benzo[b]fluoranthene, benzo[k]fluoranthene, B[a] P, indeno[1,2,3-c,d]pyrene, dibenz[a,h]anthracene and benzo[g,h,i]perylene.

13

Per natural-log unit of the carcinogenic PAHs.

14

Sum of benz[a]anthracene, chrysene/isochrysene, benzo[b]fluoranthene, benzo[k]fluoranthene, B[a]P, indeno[1,2,3-c,d]pyrene, dibenz[a,h]anthracene and benzo[g,h,i]perylene.