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WHO Guidelines for Indoor Air Quality: Selected Pollutants. Geneva: World Health Organization; 2010.

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WHO Guidelines for Indoor Air Quality: Selected Pollutants.

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6Polycyclic aromatic hydrocarbons

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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; C20H12; 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 Kow = 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 Koc = 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).

Conversion factors

At 760 mmHg and 20 °C, 1 ppm of B[a]P = 10.494 mg/m3 and 1 mg/m3 = 0.095 ppm; at 25 °C, 1 ppm of B[a]P = 10.318 mg/m3 and 1 mg/m3 = 0.097 ppm (9).

Sources and pathways of exposure


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 (1116).

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.

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.

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 6.3

Indoor PAH concentrations associated with different sources.

Table 6.4. Indoor benzo[a]pyrene 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).


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.

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).


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.


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) (7881). 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 6.6

The 16-PAH content of foods.

Table 6.7. Benzo[a]pyrene 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.

Table 6.8

Benzo[a]pyrene daily dietary intake dose.


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.

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 (96100) 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.

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).

Relative importance of different routes of exposure

Human exposure will be from both inhalation of contaminated air and consumption of contaminated food and water. Especially high exposure will occur through the smoking of cigarettes and the ingestion of certain foods (e.g. smoked and charcoal-broiled meats and fish) (2). Food ingestion is likely to be a larger route of exposure compared to inhalation for a large section of the general population exposed to PAHs. Drinking-water and soil are generally minor sources of these compounds in the daily intake dose (65).

In an earlier American study, diet was reported to make a substantial contribution (generally more than 70% in non-smokers) to PAH intake other than occupational PAH exposure. For a non-smoking reference male (70 kg body weight), a mean carcinogenic PAH intake of 3.12 μg/day was estimated, of which dietary intake contributed 96.0%, air 1.6%, water 0.2% and soil 0.4% (65). In the early 1990s, the potential dose of carcinogenic PAHs for American adult non-smoking males was estimated to be 3 μg/day up to a maximum of 15 μg/day. Smokers of unfiltered cigarettes might have had a potential dose twice that of non-smokers (65).

Recent studies conducted on human exposure to B[a]P for non-smokers in developed countries revealed that nowadays, the range and magnitude of dietary exposures (0.5–320 ng/day) (87) are generally larger than for inhalation (0.15–26 ng/day). In certain cases where indoor air contains high concentrations of PAHs, however, air could be a major contributing source. This could be the case if a person spent the day in an ETS environment (4–62 ng/day) or in microenvironments fitted with non-airtight stoves (30–7448 ng/day)7 or cooked food in the Chinese style (91–365 ng/day).8

In developing countries where biomass is generally used for cooking in homes without a flue or with a deficient flue, the contribution of inhalation to the B[a]P exposure could be as high as 138–3320 ng/day9 and therefore inhalation would be the main contributor to the total daily intake.

Indoor concentrations and their relation to outdoor concentrations

About 500 PAHs and related compounds have been detected in air, but most measurements have been made on B[a]P (2). Indoor levels have been generally found to be influenced by seasonal variations, with higher levels in winter than in summer (39,63). The levels of B[a]P in United States homes were found to be between 0.05 and 0.44 ng/m3, which were within the range of B[a]P in European homes (0.01–0.65 ng/m3) (as shown in Table 6.4). The highest B[a]P levels (0.09–25.52 ng/m3) were found in Polish homes (59).

The levels of B[a]P in Asian cities ranged between 0.21 and 3.4 ng/m3(14,6064). Higher levels of B[a]P were found in Chinese domestic kitchens. The average concentration of 12 PAHs10 in Chinese domestic kitchens was 7.6 μg/m3 and was dominated mainly by 3- and 4-ring PAHs. The B[a]P levels in domestic kitchens were 6–24 ng/m3, which was associated with conventional Chinese cooking methods (41). Lower concentrations were found in domestic kitchens in other Asian cities (12,14,106).

The use of non-airtight stoves was found to increase the levels of B[a]P by up to 2–490 ng/m3(49), while the mean indoor level of B[a]P in homes with airtight wood-burning stoves was 0.63 ng/m3(49,50), which in turn is higher than those levels recorded in non-wood-burning homes (34). High levels of B[a]P (70 ng/m3) and other PAHs have been measured in traditional rural houses with unvented fireplaces in Burundi (51).

High levels of 12 PAHs11 (164.2 ng/m3 geometric mean) have also been measured when kerosene stoves were used in Indian homes, with B[a]P geometric mean levels of 6.9 ng/m3(37). However, the highest PAH levels were measured when using other solid fuels such as coal, wood and cattle dung, with B[a]P levels ranging from 33 to 166 ng/m3. Homes in industrialized countries with ETS presented higher B[a]P levels (0.23–1.7 ng/m3) than homes without the presence of ETS (0.01–0.58 ng/m3) (22,24,39,47).

The sum of all 16 gaseous and particle-bound PAHs measured in pubs, restaurants and discotheques varied between 22 and 840 ng/m3 (B[a]P 1.45–4.1 ng/m3) (39,40,107), with discotheques/clubs the locations with the highest mean concentrations in a study carried out in Germany (40).

PAH concentrations measured in public indoor spaces ranged from 0.4–0.6 ng/m3 in hospitals, libraries and coffee shops and 1.2–1.4 ng/m3 in food courts and shopping malls in the United States (108) to 2.1–18.2 ng/m3 inside Czech kindergartens (109).

Indoor: outdoor ratios

The concentrations of low-molecular-weight PAHs (two and three rings) are usually higher indoors than outdoors, whereas those of high-molecular-weight PAHs (four rings and larger) are normally higher outdoors than indoors (63), suggesting that the indoor concentrations of the high-molecular-weight PAHs are dominated by outdoor sources (53). However, a study found that the 95th percentile of the I : O ratios of several four-ring PAHs were much higher than unity (> 3) (53), suggesting than in some homes the influence of indoor sources (generally tobacco smoking, heating or cooking sources) may be considerable (110).

The I : O ratios of individual PAHs varied from 0.3 to 10.5. Similarly, the I : O ratio of B[a]P ranged from 0.09 to 3.34 (13,17,47,53,56,6063,108,111). This variety in I : O concentration ratios suggest that the ratios fluctuate substantially across different settings, particularly those with smokers or indoor combustion sources and cooking activities (61). The differences in I : O ratios are affected by variables such as differences in combustion sources and heating systems, climatic conditions and ventilation habits. Nevertheless, smoking is generally the most relevant factor in determining I : O ratios in homes in industrialized countries (58).

Several studies found I : O ratios for PAH species of 1.4 ± 0.6 (B[a]P 1.6) in non-smokers' homes and much greater than unity (4.3 ± 3.3 for PAHs, B[a]P 5.5) in smokers' homes (13,22,47).

The I : O levels in homes using kerosene stoves for cooking and heating were 4.5 for 12 PAHs (see footnote 11) and 7.6 for B[a] P. These high values for the I : O ratio show the impact of indoor combustion sources on indoor levels of PAHs (37).

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.

Identification of studies

Studies on pharmacokinetics, metabolism and toxicology were identified by hand searching references in former reviews by WHO (2,8,70) and other authors (112) and by electronic searches in PubMed and the ISI Web of Science. As to the description of toxic effects, the focus was on in vivo studies but for the mechanisms of toxicity and metabolism all relevant studies were reviewed. Altogether, 320 original papers were selected from the literature searches with wide scope for review of contents, and 114 papers were included as relevant references for this work.


Owing to differences in the physicochemical properties of PAHs, their toxicokinetics differ widely. In this section, the focus is on the kinetics of lipophilic high-molecular-weight PAHs, such as B[a] P, because they cause the main health concern.


The major route of exposure to PAHs in the indoor environment is through the lungs and respiratory tract after inhalation of PAH-containing aerosols and particles. Data on the fate of PAHs in lungs are mainly based on animal and in vitro studies.

After deposition in the airways, the structure of the PAH and the dimensions and the chemical nature of the particles define the fate of the PAH. PAHs may dissolve from particles, the remainder in particles may be eliminated by bronchial mucociliary clearance of particles (to be swallowed), or the PAH in particles may remain in the lungs for a longer time.

B[a]P is rapidly absorbed in the lungs from solutions. After intratracheal instillation of radiolabelled B[a]P in rats, the peak concentration in the liver was attained in 10 minutes (113). The B[a]P-associated radioactivity was cleared from lungs with elimination half-lives of 5 and 116 minutes, respectively.

PAH in particles follows biphasic absorption kinetics in the lungs. The absorption kinetics depends on the site of deposition in the respiratory tract. A fraction of B[a]P in diesel particles was quickly desorbed and absorbed into circulation through type I epithelial cells in the alveolar region (114116) and systemically rapidly metabolized (116). The fraction deposited in the tracheobronchial region was more slowly absorbed into circulation and intensely locally metabolized (116). The release rates of B[a]P from particles decreased drastically after the initial burst and a notable fraction of B[a]P (up to 30%) remained unaffected on the surface of particles in lungs and in lymph nodes for several months (116).

In perfused rat lung, the absorption kinetics of B[a]P is dose-dependent (117). At low exposure levels, absorption of B[a]P in the mucosa followed the first-order kinetics with substantial local metabolism. At high exposure levels, the capacity of epithelium to dissolve and metabolize B[a]P became saturated and the absorption rate turned constant (zero-order kinetics). In the indoor environment, human exposure most likely follows the low-dose, first-order kinetics.

No data are available on exact quantitative estimates of PAH absorption in human lungs.

The kinetics of lipophilic PAHs in lungs suggest that, after deposition in lungs, (a) there is a rapid systemic exposure to B[a]P after inhalation of PAH-containing particles, (b) the intracellular B[a]P is higher in the tracheobronchial region than the alveolar region and in the epithelium lining the airways, and (c) there is a sink of B[a]P in particles to cause long-term exposure in lungs and local lymph nodes after inhalation exposure.

B[a]P and other PAHs (phenanthrene and pyrene) efficiently penetrate the skin in animals. Absorption of up to 84% of the B[a]P-associated radioactivity has been observed in mice (118) and 46% in rats (119). Absorption through human skin may be less efficient than in animals.

PAHs are ingested in house dust (as a non-dietary source) and swallowed in particles that are transported by mucociliary transport from the lungs. PAHs are readily absorbed in the gastrointestinal tract by passive diffusion (120). The composition of the diet may increase or decrease the absorption (8). Bioavailability from particles limits the absorption. From soil particles, up to 50% of total PAHs was absorbed from the gastrointestinal tract in vitro (121). The absorption was highest for small-molecule PAHs (naphthalene, acenaphthene, anthracene). The bioavailability from particles, however, probably varies depending on the content of organic carbon in dust particles.


PAHs are rapidly and widely distributed in the body. Lipophilic compounds easily pass biological membranes. Detectable levels of B[a]P can be observed in most tissues in minutes to hours after exposure, irrespective of the exposure route. PAHs undergo hepatobiliary clearance (122) and high concentrations of PAHs and their metabolites are detectable in the gastrointestinal tract (8,122).

PAHs do not accumulate in the body. Fat tends to contain more PAHs than other tissues (8). Fat and PAH contents, however, did not correlate well in lungs (123).

PAHs are generally detectable in most human tissues, typically at the sub-μg/kg level (8). The reactive metabolites are bound covalently to proteins and nucleic acids and the turnover rate of adducts defines the half-life in tissues.

Particles may cause high concentrations of PAHs in lungs. A 100-fold higher radioactivity occurred in lungs of rats after inhalation of labelled B[a]P adsorbed on carbon black particles than after inhalation of pure B[a] P. The half-time of decline also lengthened from 6 weeks to 34 weeks (124).

B[a]P and other PAHs can readily cross the placental barrier (8). The concentrations in animal embryo tissues have been, however, at one to two orders of magnitude lower than in maternal organs (125127). PAHs, including B[a] P, are detectable in maternal milk (128).


The faeces are the main route of excretion of high-molecular-weight PAHs and their metabolites (8). Biliary secretion and enterohepatic circulation are significant (122,129) and increase the concentrations of metabolites and parent compounds in the gastrointestinal tract. PAHs in bile are nearly completely present as metabolites. Less than 1% was detected as B[a]P in bile after intravenous administration of B[a]P to mice (122).

Urine is the other main excretion route. Some 4–12 % of B[a]P was excreted in urine in rats (122) compared with 60% of pyrene as metabolites (130). The role of urine as an excretion route is compound-specific; for large-molecule PAHs, it is a minor route.


Metabolism is crucial for toxicity of PAHs. Reactive intermediates and metabolites are formed that cause the toxicity and carcinogenicity. The metabolism pathways of B[a]P are best known (2,8,70,112). Most other large-molecule PAHs probably follow the same metabolism patterns (131) but the metabolic activation of sterically nonalternant PAHs, such as benzo[b]fluoranthene, may differ (2).

Three principal pathways activate PAHs for toxic intermediates and further metabolism: that via (dihydro)diol-epoxide formation, that via radical cation formation, and the o-quinone pathway (112,131). Several enzymes interplay in the metabolism.

The key enzymes in PAH metabolism are CYPs (cytochrome P450s) and epoxide hydrolase. CYPs activate PAH to optically active oxides, which rearrange to phenols. Epoxide hydrolase converts the oxides (epoxides) to optically active dihydrodiols (diols) (8,112). CYPs also metabolize PAHs to a series of quinones. For B[a] P, three quinones have been identified in vitro and in vivo: B[a]P-1,6-quinone, B[a]P-3,6-quinone and B[a]P-6,12-quinone (132). The diols can be converted to four optically active isoforms of diol-epoxides by CYPs. The diol-epoxides are highly reactive towards DNA and form a series of stable DNA adducts (112). The (+)-anti-B[a]P-7,8-diol-9,10-epoxide (anti-B[a]PDE) is suggested to be the ultimate carcinogenic form of B[a]P (112,131).

The catalytic property, mode of regulation and tissue specificity of CYPs vary and there are species differences. One or more members of the CYP family are capable of metabolizing one or more PAHs. The highest metabolism capacity is in the liver, followed by the lung, intestinal mucosa, skin and kidneys (8). Toxic metabolites producing CYPs are expressed and induced in a number of other tissues, including cardiovascular tissues (133,134). The key enzymes for PAH metabolism are CYP1A1 and CYP1B1 but several other CYPs (CYP1A2, CYP2B, CYP2C and CYP3A) also metabolize PAHs (8,112).

PAHs, especially B[a]P (135), stimulate their own metabolism by inducing CYP enzymes (8). CYP1A and CYP1B are induced via the Ah-receptor (8). Enzyme induction results in lower tissue levels of PAHs and more rapid excretion of PAHs as metabolites.

The site of induction is important for toxicity. Strong induction of the metabolism in the liver decreases PAH levels in peripheral tissues and levels of toxic metabolites by local CYP metabolism. Clear differences in PAH toxicity have been demonstrated in mice strains of different CYP induction capacity (8,136138). PAHs also inhibit CYP enzymes, and even their own metabolism (139). On the basis of toxicokinetics, PAHs may be expected to be relatively more toxic through inhalation and dermal exposure (owing to focal toxicity at the site of entry) than after oral exposure, because inhalation and dermal exposure bypass the first-pass metabolism in the liver.

The diol-epoxides have been regarded as principal toxic metabolites (70) but recent data suggest that two other routes of PAH metabolism produce toxic metabolites. In the radical cation metabolism pathway, radical cations are formed from PAH by CYPs or peroxidases and these form depurinating DNA adducts (140). In the o-quinone pathway, o-PAH diols are converted by aldo-keto reductases to catechols, which autoxidize to o-quinones. These o-quinones undergo redox cycling and form reactive oxygen species (131,141). Other B[a]P quinones have also been associated with reactive oxygen species and mutagenesis. In vivo, both mice and rats metabolize B[a]P to B[a]P-1,6-quinone, B[a]P-3,6-quinone and B[a]P-6,12-quinone and these quinones redox cycle and induce mutations (132,142). Reactive oxygen species have been associated with carcinogenesis (131,141).

Although B[a]P-diol epoxides, B[a]P-radical cations and B[a]P-o-quinones can form DNA adducts in vitro, only B[a]P-diol epoxide- and B[a]P-depurinating-DNA adducts have been measured in vivo in experimental animals and in humans (131,140,143,144). The relative importance of each activation pathway of metabolism depends on several factors, including the tissue level and stability of each activated form and the levels of expression of the activation and detoxification enzymes. For B[a] P, based on the wealth of data, the diol epoxide metabolic activation mechanism seems to be the dominant mechanism in the induction of lung carcinogenesis in rodents and humans. This conclusion is based on toxicological and mechanistic data obtained from experimental animals and from the many human biomarker studies.

PAHs and their reactive metabolites are finally converted to more polar and detoxified metabolites for excretion by the phase II metabolism enzymes, including glutathione S-transferase, UDP-glucuronosyltransferase, sulfotransferase, NAD(P)H-quinone oxidoreductase 1 and aldo-keto reductase (112). Though some of them may also be induced by PAHs, the induction is not as strong as CYP induction (145).

Genetic polymorphism may contribute to capacity to metabolize PAHs and affect toxicity. Genetic polymorphism has been described in CYP1A1, CYP1A2, CYP1B1, some CYP2C and CYP3A (8) and phase II detoxification enzymes (112,146).

Metabolism in the respiratory tract has particular relevance for toxicity of inhaled PAHs. Macrophages are actively metabolizing cells of PAHs in the lung (8). Macrophages can engulf PAH-containing particles and transport them to bronchi. It has been hypothesized that ultimate carcinogenic metabolites released from macrophages contribute to cancer development in the lung (8).

Health effects

DNA adducts

The formation of DNA adducts is a key event in mutagenicity and carcinogenicity by PAHs. Owing to the many stereoisomeric forms of B[a]P-diol epoxides (BPDE), their reactivity to covalently bind to nitrogen atoms on guanine (and to a lesser extent on adenine) bases, and epoxide ring opening yielding both cis and trans adducts, a potential total of eight unique B[a]P-diol epoxide stereoisomeric DNA adducts can be formed for each site on the nucleic acid base (131). However, far fewer stable DNA adducts are observed in vitro or in vivo. Only one diol epoxide B[a]P-DNA adduct (anti-B[a]PDE-deoxyguanosine) was observed in the lungs of mice treated with B[a]P (147) and the same adduct was found in human diploid lung fibroblasts in vitro (148) and in mononuclear white blood cells from exposed coke oven workers (149).

In heavily PAH-exposed workers, the anti-B[a]PDE-DNA adducts in peripheral blood lymphocytes were associated with increased micronuclei in cells (150). Radical cations produce a series of B[a]P adducts on guanine and adenine that are unstable (depurinating) and cleave from the DNA (131,140). o-Quinones, another metabolite of B[a] P, also form both stable and unstable adducts in vitro (131,144). PAH-DNA adduct formation blocks DNA replication and induces base and nucleotide excision repair activities (151). Errors in DNA replication (misreplication) and in DNA repair (misrepair) can create mutations that are fixed after cell division.

DNA adducts display tissue- and compound-specific qualitative and quantitative differences (152,153). B[a]P formed DNA adducts in rat lungs and liver in a dose- and time-dependent way (153). In rats and mice, the adducts reach maximal levels in tissues within a few days after a single dose, after which they gradually decrease but persist for several weeks (147,154156). In the rat lung, two adducts predominated equally (adduct with B[a]P-diol epoxide and 9-OH-B[a]P-derived adduct, about 40% of each) after intraperitoneal administration in the liver, the B[a]P-diol epoxide adduct dominated. These same adducts have also been detected in the lungs of mice, B[a]P-diol epoxide predominating (157). A comparative study with different PAHs in A/J mice indicated that the formation and persistence of DNA adducts determined the potency to induce adenomas in lungs after a single intraperitoneal administration (147).

DNA adducts have been observed postnatally in thymocytes and splenocytes of pups after in utero exposure of mice to B[a]P (158), indicating rather long persistence of the DNA adducts and vulnerability of pups to gestational exposure to B[a] P. PAH-DNA adducts have been detected in human fetal umbilical cord blood and maternal blood after exposure to ambient air PAHs at different levels (159). Prenatal exposure may increase the cancer risk of PAHs.


A number of PAHs are mutagenic and genotoxic, and induce DNA adduct formation in vitro and in vivo (8).

The potential to cause mutations is compound-dependent. Dibenzo[a,l] pyrene-diol-epoxide was over 60-fold more reactive towards DNA, induced over 200 times more mutations and yielded a fourfold higher yield of mutations per adduct than B[a]P-diol-epoxide in V79-derived XEM2 cells (160). Moreover, dibenzo[a,l]pyrene-diol-epoxide-induced adducts were less efficiently repaired.

Some PAHs probably cause mutations in a number of genes that contribute to cancer development. The anti-diol-epoxide of B[a]P ((±)-anti-BPDE) causes adducts at several hotspots of the p53 gene (161), especially in codons 157, 248 and 273 (162). The mutations by this epoxide are predominantly G to T transversions (163,164). Diol epoxides of several other PAHs cause adducts in these and other codons (165). PAH o-quinones have more potently caused similar p53 mutations in yeast reporter gene assay than (±) anti-BPDE (161). PAH-induced DNA damage stimulates cellular p53 accumulation and up-regulates the p21 protein (148,166) as typical cellular responses to DNA damage. In experimental animals, tumours induced by a series of PAHs have harboured mutations in K-ras (lung tumours) and H-ras oncogenes (skin, liver and mammary tumours). B[a]P has induced K-ras codon 12 mutations in mouse lung tumours almost exclusively at guanine, consistent with the detection of anti-BPDE-deoxyguanosine-DNA adducts in the lung tissues (167).

In human studies, lung tumours from non-smokers exposed to PAH-rich coal combustion emissions had mutations at guanine in K-ras codon 12 and p53 genes (168).

In addition to base-pair substitutions, PAHs cause other mutations to a lesser extent (exon deletions, frame-shift mutations) (163,169).

Ambient air particles have variably caused genotoxicity in vitro and DNA adduct formation (170,171). PAHs, especially B[a]P (170,172) and nitro- and oxy-PAHs, are major active components (171). B[a]P levels have indicated well the presence in ambient air of compounds causing DNA adducts (172).

In a limited data set, PAHs were assessed to contribute 3–23% of the mutagenicity in settled house dust (173). On the mass basis, settled house dust was considered on average more mutagenic than contaminated soils but less mutagenic than suspended particles in indoor and outdoor air (173).

Because some PAHs cause mutations and genotoxicity, they may generally be regarded as genotoxic carcinogens. However, PAHs also promote tumour development (see below).


B[a]P and a number of 4- to 7-ring PAHs are carcinogenic in experimental animals (8,70). Several small-molecule PAHs, such as anthracene, perylene and fluorene, have not been carcinogenic and the carcinogenicity of some compounds (acenaphthene, phenanthrene, pyrene) is, as yet, questionable (8). Inhalation of naphthalene has induced respiratory tract tumours in mice and rats at high cytotoxic concentrations but not at non-cytotoxic concentrations (174,175).

Carcinogenic PAHs such as B[a]P have induced tumours through dermal, oral, intraperitoneal, intramamillary and respiratory tract routes (8). The species that have developed tumours after exposure to PAHs include mice, rats, rabbits, hamsters and monkeys (8). Tumour induction is not restricted to the site of administration. After oral exposure to PAHs, tumours have been observed typically in the liver, forestomach, lungs and mammary glands (8). PAHs painted onto skin have caused skin papillomas and carcinomas but also lung and liver tumours (8,70). Administration of B[a]P into the respiratory tract has consistently caused lung tumours in mice, rats, hamsters and monkeys (8). Fewer data exist for other PAHs after exposure via the respiratory tract, but acenapthene, benzo[b]fluoranthene, benzo[j]fluoranthene, benzo[k]fluoranthene, chrysene, dibenz[a,h]anthracene and indeno[1,2,3-c,d]pyrene have caused lung tumours in rats and dibenz[a,h]anthracene and dibenzo[a,i]pyrene in hamsters (8).

The potencies of the carcinogenic PAHs differ by three orders of magnitude (8). In comparative studies with PAHs, B[a]P was repeatedly a potent carcinogen after dermal application (8) but in an initiation–promotion model in SENCAR mice, dibenzo[a,l]pyrene and 7,12-dimethyl-benz[a]anthracene were more potent initiators of skin tumorigenesis than B[a]P (176,177). Based on the administered dose, dibenzo[a,l]pyrene was also more potent than B[a]P in inducing mouse lung adenomas (178). Since dibenzo[a,l]pyrene was also the most potent in inducing mammary tumours after intramamillary injection (176), it may be regarded as the most potent carcinogenic PAH known.

B[a]P (8) and dibenzo[a,l]pyrene (179,180) have been shown to be transplacental carcinogens in mice, causing lung and liver tumours (B[a]P and dibenzo[a,l]pyrene) and lymphoma (dibenzo[a,l]pyrene) in progeny after in utero exposure.

Limited data are available on the potency of specific PAHs to induce lung cancer following inhalation exposure. Data are inadequate for ranking the potency of specific PAHs to induce lung cancer.

The Ah-receptor-mediated pathways are crucial for carcinogenicity of PAHs: Ah-receptor-deficient mice are resistant to B[a]P- and dibenzo[a,l]pyreneinduced skin cancer (181,182). Organic extracts of airborne particulate matter, where PAHs are suggested to be the primary carcinogens, have not caused lung tumours in AhR -/- mice either (183).

Although genotoxic effects (mutations in cancer genes and DNA damage) are likely to be the primary events in PAH-induced carcinogenesis, in vitro studies have indicated that PAHs have also non-genotoxic effects that could contribute to carcinogenesis. B[a]P-diol epoxide induces gene promoter hypermethylation in immortalized bronchial epithelial cells (184) and several PAHs inhibit gap-junctional intercellular communication (185), a typical mechanism of tumour promotion. Several small-molecule PAHs have been more potent in inhibiting gap-junctional intercellular communication in liver cells than high-molecular-weight PAHs (185). Anti-B[a]P-7,8-diol-9,10-oxide has been shown to increase cell proliferation in human cell lines, including lung cancer cells (186) and to induce apoptosis in H460 human lung cancer cells (187). B[a]P has induced apoptosis in human lung fibroblast MRC-5 cells via the JNK1/FasL and JNK1/p53 signals (188). It is possible that PAHs with low genotoxic potential also promote tumour development by non-genotoxic mechanisms.

PAHs have induced the expression of a number of genes in cells in vitro with compound-specific profiles (189191). Little can be interpreted as yet, however, about the mechanisms of toxicity from the gene and protein expression data. In addition to induction of PAH-metabolizing CYP genes, only oxidative stress pathway genes were induced by carcinogenic PAHs in rat liver slices (190). In human mammary carcinoma-derived cells (MCF-7), cytoskeletal proteins, heat-shock proteins, DNA-associated proteins and glycolytic and mitochondrial proteins were altered (191). Dibenzo[a,l]pyrene and B[a] P, two potent carcinogenic PAHs, have consistently displayed different gene expression patterns (189191). Diol epoxides of carcinogenic PAHs, but not the parent PAHs, have increased intracellular Ca2+ in human small-airway epithelial cells in vitro (192). Increased intracellular Ca2+ is likely to be one mechanism contributing to the toxicity of diol epoxides.

General toxicity

There are limited data on the toxicity of individual PAHs in experimental animals. Data on mixtures containing PAHs as principal toxic components (coal tar, coal tar pitch and creosote) complement the information. In general, the acute toxicity of PAHs in animals is low to moderate (8). B[a]P causes eye irritation and skin sensitization in animals (8) and PAH mixtures are phototoxic in both skin and eyes (193).

On repeated exposure, the main target organs of toxicity are the liver or kidneys in animals, depending on the PAH (8). Typically, the weight of the liver increases owing to enzyme induction. Nephropathy and decreased kidney weight have been caused by pyrene in mice (8).

High oral doses of B[a]P have caused bone marrow depression in mice, decreasing especially proliferating haematopoietic cells (137). B[a]P has also impaired in vitro proliferation and differentiation of human haematopoietic CD34+ stem cells and caused their apoptosis (194). By destroying these cells, PAHs may down-regulate cell lineages (lymphocytes, macrophages, neutrophils) important for immunoresponse.

PAHs are immunotoxic and cause immunosuppression. Immunotoxicity of PAHs has been demonstrated in several different cells in vitro (194197). Immunotoxicity requires focal PAH metabolic activation in cells (194,197). Immunosuppression in B6C3F1 mice has followed the structure–activity relationship observed for the carcinogenicity of PAHs, with benz[a]anthracene, B[a] P, dibenz[a,c]anthracene and dibenz[a,h]anthracene being more potent than anthracene, chrysene, benzo[e]pyrene and perylene (136). The greatest immunosuppression was observed with 3-methylcholanthrene and 7,12-dimethylbenz[a] anthracene.

PAHs, including B[a] P, dibenz[a,h]anthracene, dibenz[a,c]anthracene and 7,12-dimethylbenz[a]anthracene, have accelerated atherosclerosis plaque formation in Ah-responsive mice, chickens and pigeons (198,199). The atherogenic effect of PAHs may not be associated with their mutagenic and carcinogenic capacity (200). PAHs are thought to cause adverse cardiovascular effects by metabolites through activation of the Ah-receptor: by increased production of reactive oxygen species, induction of inflammatory-mediated (201) and hypertrophic genes, and increased disruption of endogenous substances such as prostaglandins (134). Therefore, induction of inflammation might be a crucial process in PAH-enhanced atherogenesis.

B[a]P has decreased fertility and caused embryotoxicity. It increased primordial oocyte destruction, decreased the number of corpora lutea, caused resorptions, decreased the number of pups and decreased fetal weight in rats and mice (8). Sterility associated with alterations in gonadal tissues has also been observed in mice after prenatal exposure, in both females and males (202). In male rats, B[a]P reduced testis weight (203) and decreased the testosterone level in the blood and sperm motility (203,204), probably contributing to reproduction toxicity.

PAHs are also teratogenic in mice and rats (8,205). Reproduction toxicity has also been noted by the dermal and parenteral exposure routes. Inhalation exposure of rats to B[a]P during pregnancy decreased plasma estrogen, progesterone and prolactin levels in dams in association with decreased pup survival and development, thus partly explaining the effects (206). B[a]P exposure before conception caused Ah-receptor-dependent fetal intrauterine growth restriction in mice, which was associated with altered vasculature in the placenta and a decreased placental cell death rate (207). Maternally toxic doses of B[a]P and 7,12-dimethylbenz[a]anthracene caused necrosis in the placenta and haemorrhages in fetuses in rats (208), suggesting that the vascular system in general is one target of PAHs.

B[a] P, benz[a]anthracene and fluoranthene displayed weak estrogenic activity in rat immature uterotropic assay (209), with 3–4 orders of magnitude lower potency than endogenous estrogen. Some PAHs (including B[a]P), and especially their hydroxylated metabolites, interact with estrogen receptors (210,211). PAHs have also indicated anti-estrogenic effects (212). Contrasting effects may be explained by a complex crosstalk between Ah-receptor and estrogen receptors and induction of estrogen metabolism by PAHs (209).

PAH mixtures

In indoor air, exposure to PAHs occurs mostly in mixture. Since the toxicity of PAHs depends nearly exclusively on their biotransformation to toxic metabolites, interactions at the level of key metabolism enzymes are highly relevant to the associated health risk. Induction of metabolism by one PAH may enhance the toxicity of another; inhibition of the metabolism may decrease the toxicity.

Gene expression data with B[a]P and dibenzo[a,h]anthracene in binary mixture with dibenzo[a,l]pyrene, benzo[b]fluoranthene and fluoranthene in precision-cut rat liver slices in vitro indicated that (a) each of them induced an altered expression of genes, (b) the genes affected considerably by the combination of PAHs were slightly different from those altered by either of the constituents, (c) total altered expression of the genes by a mixture was less than that induced by individual PAHs and (d) the interactions were mostly antagonistic, leading to decreased altered expression of genes by the mixture compared to single PAHs (213). Fewer DNA adducts were formed by the mixtures than by individual PAHs. In contrast, in human hepatoma cells (HepG2), equimolar and equitoxic mixtures of these same PAHs have shown an additive effect on apoptosis and cell cycle blockage, an additive or antagonistic effect on gene expression, and a synergistic effect on DNA adduct formation (214). Since the interactions depend on PAH composition, concentrations and cell types, contrasting results may be expected. However, the majority of in vitro studies have shown an antagonistic effect on DNA adduct formation by a mixture of PAHs (172,189,215).

The PAH metabolites have been shown to interfere with each other. The major B[a]P metabolite 3-hydroxybenzo(a)pyrene inhibits both the mutagenic and tumorigenic activity of the carcinogenic metabolite anti-B[a]PDE (216). When applied topically to mouse skin, different binary and tertiary mixtures of carcinogenic PAHs formed DNA adducts at levels that were additive, less than additive and greater than additive relative to the levels formed when the compounds were applied individually (217). Complex PAH mixtures have also decreased skin tumorigenicity of B[a]P (218) and that of dibenzo[a,l]pyrene in mice (219). Decreased tumorigenesis was associated with decreased DNA adduct formation. In contrast, mixtures of five PAHs both enhanced and inhibited mouse lung tumorigenesis, depending on the composition of the PAH mixture and the dose (220).

Altogether, the experimental data indicate that mixture effects of PAHs may be complex in vitro and in vivo. Attenuation of the toxicity rather than synergism has been observed in several studies. These observations imply that the effects of a PAH mixture may not be reliably predicted from single PAH components.

Biomarkers for evaluation of exposure

Internal exposure to PAHs has been assessed mostly by urine 1-hydroxypyrene or aromatic bulky DNA adducts in peripheral lymphocytes in humans (8).

1-Hydroxypyrene, a metabolite of non-carcinogenic pyrene detectable in urine, may be used as a general biomarker of exposure to PAH mixtures (8,221). Because urinary 1-hydroxypyrene displays all possible sources of PAHs (including food and ambient air) and all exposure routes, it indicates total exposure to pyrene-containing PAH mixtures. A good correlation between the PAH concentration in air and urine 1-hydroxypyrene has been observed in several occupational environments (8). In single studies, a significant correlation has also been observed with exposure to residential indoor air sources of PAHs, such as ETS, cooking practices and the burning of coal for heating (222,223). A review on environmental exposure to PAHs in ambient air indicated that urine 1-hydroxypyrene may serve as a qualitative indicator of excess exposure to PAHs at the group level but a large inter-individual variation limits its use for personal exposure (224).

The levels of DNA adducts reflect not only the exposure to PAHs but also the body's ability to metabolize them. In general, exposures that have led to increased excretion of 1-hydroxypyrene have led to elevated adduct levels (8). This is demonstrated best on occupational exposure to PAHs. However, the inter-individual variation in DNA adducts has been large (up to 50–100-fold) and the correlation between measured/estimated exposure to PAHs and adduct levels variable (from clear to no correlation, for example (225)). The correlation may be better when exposure is measured personally (226). As to the general population, elevated DNA adduct levels in blood leucocytes have been detected in populations living in industrialized areas (227229). Consumption of charcoal-grilled food increases their levels (230,231), as does high indoor air exposure (232). Total PAH-DNA adduct levels and BPDE-DNA adduct levels were significantly higher in smokers than among non-smokers (233). Altogether, PAH-DNA adducts can be used as a qualitative biomarker of exposure to combustion emissions, most reliably on a group basis. DNA adducts are considered to be a less sensitive parameter for exposure assessment than excretion of 1-hydroxypyrene in urine (8,224).

Environmental exposure to PAHs may not be assessed reliably on the basis of exposure biomarkers such as urinary 1-hydroxypyrene concentration or aromatic bulky PAH-DNA adducts in blood cells. This is particularly true in indoor environments, where assessment at individual level is often needed. There are no proper risk functions, the individual variation in biomarkers is large and the biomarkers measure exposure to all possible sources of PAHs (including ambient air and food).

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 PM2.5 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 PM2.5 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.

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.


  • 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

Critical health outcomes

Biomarkers of exposure and effects

A growing body of evidence suggests that exposure to B[a]P at levels over 1.0 ng/m3 induces DNA damage. Personal exposure to B[a]P over 1.0 ng/m3 predicted greater genomic frequency of translocations, micronuclei and DNA fragmentation in sperm. In children in various developing countries, a number of markers for cytotoxic and oxidative stress have been positively correlated with either monitored personal PAH concentration or carcinogenic PAH levels in whole blood. Further, elevated exposure to B[a]P has been associated with higher levels of PAH-DNA adducts, urinary 1-hydroxypyrene, DNA strand breaks and impaired DNA repair capacity.

Intrauterine growth restriction

Intrauterine growth restriction has been defined in terms of low birth weight (< 2500 g), low birth weight at full-term (≥ 37 weeks) or SGA (< 10th percentile of population mean weight at a given gestational age and gender). A strong body of data demonstrates a significant role of prenatal exposure to particle-bound PAHs in reduced or low birth weight in Europe and the United States (236). The direction and size of birth weight reduction are consistent overall. In addition, prenatal exposure to several PAHs induced severe fetotoxic effects in several animal species. Thus, it is concluded that sufficient evidence exists of a relationship between prenatal exposure to mixtures of carcinogenic PAHs and intrauterine growth restriction in humans.

Lung cancer

Occupational exposure to complex mixtures containing PAHs has been strongly associated with lung cancer. Studies on experimental systems have shown many PAHs to be genotoxic and carcinogenic. A detectable level of leukocyte PAH/ bulky DNA adducts in non-smokers was correlated with an increased risk of lung cancer compared to those with a non-detectable level. When coded as a continuous variable, each unit increase in DNA adducts led to a 25 % increase in the risk of lung cancer, without a threshold. In occupational settings, with B[a]P exposure ranging between 0.04 and 40 μg/m3, the risk of lung cancer increased by 20-168% per 100 μg/m3 B[a]P-years. As a result, sufficient evidence of a causal relationship is considered to exist between exposure to mixtures of airborne PAHs containing B[a]P and lung cancer. There is a sufficient body of evidence to support the critical role of PAH/bulky DNA adducts, mutations in tumour oncogenes and tumour suppressor genes in the development of lung cancer following exposure to PAHs.

Bladder cancer

Risk of bladder cancer has often been examined along with that of lung cancer in investigations of occupational exposures to PAHs. The risks were significantly elevated for some industries involving exposure to complex mixtures containing PAHs, namely aluminium production, paving and roofing, and chimney sweeping. Experimentally, PAHs have not been shown to induce bladder cancer. It is concluded that insufficient evidence exists of a relationship between exposure to mixtures of airborne PAHs containing B[a]P and bladder cancer.

Breast cancer

Risk of breast cancer has been examined mostly in terms of detectable levels of PAH-DNA adducts. A modest (29–35%) elevation in the likelihood of breast cancer diagnosis has been observed for those with detectable levels of PAH-DNA adducts compared to non-detectable levels, without an apparent dose–response relationship. However, when the follow-up was continued to the point of death, the detectable PAH-DNA adducts did not increase the risk of all-cause or breast cancer mortality. It is concluded that insufficient evidence exists for a relationship between exposure to mixtures of airborne PAHs containing B[a]P and breast cancer.

Fatal ischaemic heart disease

In experimental animals, PAHs including B[a] P, dibenz[a,h]anthracene, dibenz[a,c]anthracene and 7,12-dimethylbenz[a]anthracene have accelerated atherosclerosis plaque formation (198,199). Following several inconclusive studies, a multinational cohort of male asphalt workers suggested that risk of ischemic heart disease (IHD) was associated with B[a]P exposure in a dose–responsive manner. However, the authors could still not rule out the possibility of confounding by exposure to fine particulates in the occupational setting (248). It is concluded that limited evidence exists for a relationship between exposure to mixtures of airborne PAHs containing B[a]P and IHD. If these findings are corroborated, the IHD mortality from PAH exposure would be higher than that for lung cancer.

Health relevance of indoor exposure

Since there is sufficient evidence that some PAHs, including B[a] P, are genotoxic carcinogens, no threshold can be determined and all indoor exposures are considered relevant to health.

Indoor airborne levels of PAHs are influenced not only by infiltration of outdoor PAHs but also by lifestyle-related indoor emissions. In particular, children and women in developing countries are exposed to multiple and season-dependent sources of PAHs, including domestic burning of wood, coal, cow dung and kerosene, as well as industrial coal-burning and road traffic. Indoor B[a]P levels in homes that use biomass and coal for heating and cooking range from 33 to 186 ng/m3 (range of geometric means) compared to B[a]P levels generally less than 1 ng/m3 in non-smoking homes in developed countries (typically based on a 24hour mean). The indoor B[a]P concentrations in developing countries increase the inhalation doses (105–2523 ng/day; range of geometric means). Hence, in these situations, the inhalation of particle-bound PAHs is at least as important a route of exposure as dietary exposure.

Conclusions of other reviews

IARC (273) concluded, based on occupational studies, that there is sufficient evidence that coal gasification, soot (as found in occupational exposure of chimney sweeps), aluminium production, coal tar pitch (as encountered in paving and roofing), iron and steel founding and coke production cause human lung cancer (259). There is sufficient evidence that aluminium production causes bladder cancer in humans.

There is limited evidence in humans for a causal association of soot and coal tar pitch with bladder cancer (259). Indoor emissions from household combustion of coal are carcinogenic to humans (Group 1), inducing lung cancer. Indoor emissions from household combustion of biomass fuel (mainly wood) are probably carcinogenic to humans (Group 2A), inducing lung cancer. Emissions from high-temperature frying are probably carcinogenic to humans (Group 2A) (259).

B[a]P was reclassified by IARC (260) as a human carcinogen (Group 1) based on sufficient evidence of carcinogenic activity in animals and strong evidence that the mechanisms of carcinogenicity in animals operate in humans (259). Cigarette smoke contains PAHs and ETS has also been classified as carcinogenic to humans (Group 1) (260).

WHO concluded in 2000 (2) that occupational epidemiology data should serve as the basis for the risk estimate. Based on epidemiological data from studies in coke-oven workers, a unit risk for B[a]P as an indicator in ambient air constituents was estimated to be 8.7 × 10−5 per ng/m3, which is the same as that established by WHO in 1987 (23).


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.

Summary of main evidence and decision-making in guideline formulation

Critical outcome for guideline definition

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.

Source of exposure–effect evidence

There is sufficient evidence that some PAHs, including B[a] P, are genotoxic carcinogens. Based on epidemiological data from studies in coke-oven workers, a unit risk for B[a]P as an indicator of PAH in ambient air was estimated to be 8.7 × 10−5 per ng/m3(2,23).

Supporting evidence

Studies on early biological effects of PAH exposure based on biomarkers in general populations of children and adults (252257), on carcinogenic effects in the occupational setting (259) and on mutagenic and carcinogenic risk of PAH-DNA adducts (261269).

Results of other reviews

IARC: B[a]P and PAH-containing indoor emissions from household combustion of coal have been classified in Group 1 (human carcinogens) (259,260).


No threshold can be determined and all indoor exposures are considered relevant to health.

Unit risk for lung cancer for PAH mixtures is estimated to be 8.7 × 10−5 per ng/m3 of B[a]P.

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.


B[a]P is taken as a marker of the PAH mixture. Use of the B[a]P unit risk factor assumes that B[a]P represents the same proportion of carcinogenic activity of the PAH mixture in all indoor environments as in the occupational setting. This assumption will not always hold, but the associated uncertainties in risk estimates are unlikely to be large.


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The octanol–water partition coefficient (Kow) 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).


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


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.


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


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.


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.


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


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


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


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.


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


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.


Per natural-log unit of the carcinogenic PAHs.


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.

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