TOXICOLOGICAL PROFILE FOR
 

POLYCYCLIC AROMATIC HYDROCARBONS
 

U.S. DEPARTMENT OF HEALTH AND HUMAN SERVICES
 

Public Health Service
 

Agency for Toxic Substances and Disease Registry
 

August 1995 
PAHs ii 

DISCLAIMER 

The use of company or product name(s) is for identification only and does not imply endorsement by 

the Agency for Toxic Substances and Disease Registry. 
PAHs iii 

UPDATE STATEMENT
 

A Toxicological Profile for Polycyclic Aromatic Hydrocarbons was released in December 1990. This 

edition supersedes any previously released draft or final profile. 

Toxicological profiles are revised and republished as necessary, but no less than once every three 

years. For information regarding the update status of previously released profiles, contact ATSDR at: 

Agency for Toxic Substances and Disease Registry
 

Division of Toxicology/Toxicology Information Branch
 

1600 Clifton Road NE, E-29
 

Atlanta, Georgia 30333
 















PAHs 1 

1. PUBLIC HEALTH STATEMENT 

This statement was prepared to give you information about polycyclic aromatic hydrocarbons 

(PAHs) and to emphasize the human health effects that may result from exposure to them. 

The Environmental Protection Agency (EPA) has identified 1,408 hazardous waste sites as the 

most serious in the nation. These sites make up the National Priorities List (NPL) and are the 

sites targeted for long-term federal clean-up activities. PAHs have been found in at least 

600 of the sites on the NPL. However, the number of NPL sites evaluated for PAHs is not 

known. As EPA evaluates more sites, the number of sites at which PAHs are found may 

increase. This information is important because exposure to PAHs may cause harmful health 

effects and because these sites are potential or actual sources of human exposure to PAHs. 

When a substance is released from a large area, such as an industrial plant, or from a 

container, such as a drum or bottle, it enters the environment. This release does not always 

lead to exposure. You can be exposed to a substance only when you come in contact with it. 

You may be exposed by breathing, eating, or drinking substances containing the substance or 

by skin contact with it. 

If you are exposed to substances such as PAHs, many factors will determine whether harmful 

health effects will occur and what the type and severity of those health effects will be. These 

factors include the dose (how much), the duration (how long), the route or pathway by which 

you are exposed (breathing, eating, drinking, or skin contact), the other chemicals to which 

you are exposed, and your individual characteristics such as age, sex, nutritional status, family 

traits, lifestyle, and state of health. 

1.1 WHAT ARE POLYCYCLIC AROMATIC HYDROCARBONS? 

PAHs are a group of chemicals that are formed during the incomplete burning of coal, oil, 

gas, wood, garbage, or other organic substances, such as tobacco and charbroiled meat. There 

are more than 100 different PAHs. PAHs generally occur as complex mixtures (for example, 

as part of combustion products such as soot), not as single compounds. PAHs usually occur 
PAHs 2 

1. PUBLIC HEALTH STATEMENT 

naturally, but they can be manufactured as individual compounds for research purposes; 

however, not as the mixtures found in combustion products. As pure chemicals, PAHs 

generally exist as colorless, white, or pale yellow-green solids. They can have a faint, 

pleasant odor. A few PAHs are used in medicines and to make dyes, plastics, and pesticides. 

Others are contained in asphalt used in road construction. They can also be found in 

substances such as crude oil, coal, coal tar pitch, creosote, and roofing tar. They are found 

throughout the environment in the air, water, and soil. They can occur in the air, either 

attached to dust particles or as solids in soil or sediment. 

Although the health effects of individual PAHs are not exactly alike, the following 17 PAHs 

are considered as a group in this profile: 

• acenaphthene 

• acenaphthylene 

• anthracene 

• benz[a]anthracene 

• benzo[a]pyrene 

• benzo[e]pyrene 

• benzo[b]fluoranthene 

• benzo[g,h,i]perylene 

• benzo[j]fluoranthene 

• benzo[k]fluoranthene 

• chrysene 

• dibenz[a,h]anthracene 

• fluoranthene 

• fluorene 

• indeno[ 1,2,3-c,d]pyrene 

• phenanthrene 

• pyrene 

These 17 PAHs were chosen to be included in this profile because (1) more information is 

available on these than on the others; (2) they are suspected to be more harmful than some of 

the others, and they exhibit harmful effects that are representative of the PAHs; (3) there is a 

greater chance that you will be exposed to these PAHs than to the others; and (4) of all the 

PAHs analyzed, these were the PAHs identified at the highest concentrations at NPL 

hazardous waste sites. 
 

 

PAHs 3 

1. PUBLIC HEALTH STATEMENT 

More information can be found on the chemical and physical properties of PAHs in Chapter 

3 and on their use and disposal in Chapter 4. 

1.2 WHAT HAPPENS TO 	 POLYCYCLIC AROMATIC HYDROCARBONS WHEN

 THEY ENTER THE ENVIRONMENT? 

PAHs enter the environment mostly as releases to air from volcanoes, forest fires, residential 

wood burning, and exhaust from automobiles and trucks. They can also enter surface water 

through discharges from industrial plants and waste water treatment plants, and they can be 

released to soils at hazardous waste sites if they escape from storage containers. The 

movement of PAHs in the environment depends on properties such as how easily they 

dissolve in water, and how easily they evaporate into the air. PAHs in general do not easily 

dissolve in water. They are present in air as vapors or stuck to the surfaces of small solid 

particles. They can travel long distances before they return to earth in rainfall or particle 

settling. Some PAHs evaporate into the atmosphere from surface waters, but most stick to 

solid particles and settle to the bottoms of rivers or lakes. In soils, PAHs are most likely to 

stick tightly to particles. Some PAHs evaporate from surface soils to air. Certain PAHs in 

soils also contaminate underground water. The PAH content of plants and animals living on 

the land or in water can be many times higher than the content of PAHs in soil or water. 

PAHs can break down to longer-lasting products by reacting with sunlight and other 

chemicals in the air, generally over a period of days to weeks. Breakdown in soil and water 

generally takes weeks to months and is caused primarily by the actions of microorganisms. 

For more information on what happens to PAHs in the environment see Chapter 5. 

1.3 	 HOW MIGHT I BE EXPOSED TO POLYCYCLIC AROMATIC

 HYDROCARBONS? 

PAHs are present throughout the environment, and you may be exposed to these substances at 

home, outside, or at the workplace. Typically, you will not be exposed to an individual PAH, 

but to a mixture of PAHs. 
PAHs 4 

1. PUBLIC HEALTH STATEMENT 

In the environment, you are most likely to be exposed to PAH vapors or PAHs that are 

attached to dust and other particles in the air. Sources include cigarette smoke, vehicle 

exhausts, asphalt roads, coal, coal tar, wildfires, agricultural burning, residential wood 

burning, municipal and industrial waste incineration, and hazardous waste sites. Background 

levels of some representative PAHs in the air are reported to be 0.02-1.2 nanograms per 

cubic meter (ng/m3; a nanogram is one-millionth of a milligram) in rural areas and 

0.15-19.3 ng/m3 in urban areas. You may be exposed to PAHs in soil near areas where coal, 

wood, gasoline, or other products have been burned. You may be exposed to PAHs in the 

soil at or near hazardous waste sites, such as former manufactured-gas factory sites and 

wood-preserving facilities. PAHs have been found in some drinking water supplies in the 

United States. Background levels of PAHs in drinking water range from 4 to 24 nanograms 

per liter (ng/L; a liter is slightly more than a quart). 

In the home, PAHs are present in tobacco smoke, smoke from wood fires, creosote-treated 

wood products, cereals, grains, flour, bread, vegetables, fruits, meat, processed or pickled 

foods, and contaminated cow’s milk or human breast milk. Food grown in contaminated soil 

or air may also contain PAHs. Cooking meat or other food at high temperatures, which 

happens during grilling or charring, increases the amount of PAHs in the food. The level of 

PAHs in the typical U.S. diet is less than 2 parts of total PAHs per billion parts of food 

(ppb), or less than 2 micrograms per kilogram of food (μg/kg; a microgram is one-thousandth of a 

milligram). 

The primary sources of exposure to PAHs for most of the U.S. population are inhalation of 

the compounds in tobacco smoke, wood smoke, and ambient air, and consumption of PAHs in 

foods. For some people, the primary exposure to PAHs occurs in the workplace. PAHs have 

been found in coal tar production plants, coking plants, bitumen and asphalt production plants, 

coal-gasification sites, smoke houses, aluminum production plants, coal tarring facilities, and 

municipal trash incinerators. Workers may be exposed to PAHs by inhaling engine exhaust 

and by using products that contain PAHs in a variety of industries such as mining, oil 

refining, metalworking, chemical production, transportation, and the electrical industry. PAHs 

have also been found in other facilities where petroleum, petroleum products, or coal are used 
PAHs 5 

1. PUBLIC HEALTH STATEMENT 

or where wood, cellulose, corn, or oil are burned. People living near waste sites containing 

PAHs may be exposed through contact with contaminated air, water, and soil. For more 

information on human exposure to PAHs, see Chapter 5. 

1.4 	 HOW CAN POLYCYCLIC AROMATIC HYDROCARBONS ENTER AND LEAVE

 MY BODY? 

PAHs can enter your body through your lungs when you breathe air that contains them 

(usually stuck to particles or dust). Cigarette smoke, wood smoke, coal smoke, and smoke 

from many industrial sites may contain PAHs. People living near hazardous waste sites can 

also be exposed by breathing air containing PAHs. However, it is not known how rapidly or 

completely your lungs absorb PAHs. Drinking water and swallowing food, soil, or dust 

particles that contain PAHs are other routes for these chemicals to enter your body, but 

absorption is generally slow when PAHs are swallowed. Under normal conditions of 

environmental exposure, PAHs could enter your body if your skin comes into contact with 

soil that contains high levels of PAHs (this could occur near a hazardous waste site) or with 

used crankcase oil or other products (such as creosote) that contain PAHs. The rate at which 

PAHs enter your body by eating, drinking, or through the skin can be influenced by the 

presence of other compounds that you may be exposed to at the same time with PAHs. 

PAHs can enter all the tissues of your body that contain fat. They tend to be stored mostly in 

your kidneys, liver, and fat. Smaller amounts are stored in your spleen, adrenal glands, and 

ovaries. PAHs are changed by all tissues in the body into many different substances. Some 

of these substances are more harmful and some are less harmful than the original PAHs. 

Results from animal studies show that PAHs do not tend to be stored in your body for a long 

time. Most PAHs that enter the body leave within a few days, primarily in the feces and 

urine. More information on how PAHs enter and leave your body can be found in Chapters 2 

and 6. 
PAHs 6 

1. PUBLIC HEALTH STATEMENT 

1.5	 HOW CAN POLYCYCLIC AROMATIC HYDROCARBONS AFFECT MY

 HEALTH? 

PAHs can be harmful to your health under some circumstances. Several of the PAHs, 

including benz[a]anthracene, benzo[a]pyrene, benzo[b]fluoranthene, benzo[j]fluoranthene, 

benzo[k]fluoranthene, chrysene, dibenz[a,h]anthracene, and indeno [1,2,3-c,d]pyrene, have 

caused tumors in laboratory animals when they breathed these substances in the air, when 

they ate them, or when they had long periods of skin contact with them. Studies of people 

show that individuals exposed by breathing or skin contact for long periods to mixtures that 

contain PAHs and other compounds can also develop cancer. 

Mice fed high levels of benzo[a]pyrene during pregnancy had difficulty reproducing and so 

did their offspring. The offspring of pregnant mice fed benzo[a]pyrene also showed other 

harmful effects, such as birth defects and decreased body weight. Similar effects could occur 

in people, but we have no information to show that these effects do occur. 

Studies in animals have also shown that PAHs can cause harmful effects on skin, body fluids, 

and the body’s system for fighting disease after both short- and long-term exposure. These 

effects have not been reported in people. 

The Department of Health and Human Services (DHHS) has determined that 

benz[a]anthracene, benzo[b]fluoranthene, benzo[j]fluoranthene, benzo[k]fluoranthene, 

benzo[a]pyrene, dibenz[a,h]anthracene, and indeno[ 1,2,3-c,d]pyrene are known animal 

carcinogens. The International Agency for Research on Cancer (IARC) has determined the 

following: benz[a]anthracene and benzo[a]pyrene are probably carcinogenic to humans; 

benzo[b]fluoranthene, benzo[j]fluoranthene, benzo[k]fluoranthene, and indeno[ 1,2,3-c,d]pyrene 

are possibly carcinogenic to humans; and anthracene, benzo[g,h,i]perylene, benzo[e]pyrene, 

chrysene, fluoranthene, fluorene, phenanthrene, and pyrene are not classifiable as to their 

carcinogenicity to humans. EPA has determined that benz[a]anthracene, benzo[a]pyrene, 

benzo[b]fluoranthene, benzo[k]fluoranthene, chrysene, dibenz[a,h]anthracene, and 

indeno[ 1,2,3-c,d]pyrene are probable human carcinogens and that acenaphthylene, anthracene, 
PAHs 7 

1. PUBLIC HEALTH STATEMENT 

benzo[g,h,i]perylene, fluoranthene, fluorene, phenanthrene, and pyrene are not classifiable as 

to human carcinogenicity. Acenaphthene has not been classified for carcinogenic effects by 

the DHHS, IARC, or EPA. More information on the health effects associated with exposure 

to PAHs can be found in Chapter 2. 

1.6 	 IS THERE A MEDICAL TEST TO DETERMINE WHETHER I HAVE BEEN

 EXPOSED TO POLYCYCLIC AROMATIC HYDROCARBONS? 

In your body, PAHs are changed into chemicals that can attach to substances within the body. 

The presence of PAHs attached to these substances can then be measured in body tissues or 

blood after exposure to PAHs. PAHs or their metabolites can also be measured in urine, 

blood, or body tissues. Although these tests can show that you have been exposed to PAHs, 

these tests cannot be used to predict whether any health effects will occur or to determine the 

extent or source of your exposure to the PAHs. It is not known how effective or informative 

the tests are after exposure is discontinued. These tests to identify PAHs or their products are 

not routinely available at a doctor’s office because special equipment is required to detect 

these chemicals. More information on tests used to determine the presence of PAHs in your 

body is presented in Chapters 2 and 6. 

1.7 	 WHAT RECOMMENDATIONS HAS THE FEDERAL GOVERNMENT MADE TO

 PROTECT HUMAN HEALTH? 

The federal government has set regulations to protect people from the possible health effects 

of eating, drinking, or breathing PAHs. EPA has suggested that taking into your body each 

day the following amounts of individual PAHs is not likely to cause any harmful health 

effects: 0.3 milligrams (mg) of anthracene, 0.06 mg of acenaphthene, 0.04 mg of 

fluoranthene, 0.04 mg of fluorene, and 0.03 mg of pyrene per kilogram (kg) of your body 

weight (one kilogram is equal to 2.2 pounds). Actual exposure for most of the United States 

population occurs from active or passive inhalation of the compounds in tobacco smoke, 

wood smoke, and contaminated air, and from eating the compounds in foods. Skin contact 
PAHs 8 

1. PUBLIC HEALTH STATEMENT 

with contaminated water, soot, tar, and soil may also occur. Estimates for total exposure in 

the United States population have been listed as 3 mg/day. 

From what is currently known about benzo[a]pyrene, the federal government has developed 

regulatory standards and guidelines to protect people from the potential health effects of 

PAHs in drinking water. EPA has provided estimates of levels of total cancer-causing PAHs 

in lakes and streams associated with a risk of human cancer development. If the following 

amounts of individual PAHs are released to the environment within a 24-hour period, EPA 

must be notified: 1 pound of benzo[b]fluoranthene, benzo[a]pyrene, or dibenz[a,h]anthracene; 

10 pounds of benz[a]anthracene; 100 pounds of acenaphthene, chrysene, fluoranthene, or 

indeno[ 1,2,3-c,d]pyrene; or 5,000 pounds of acenaphthylene, anthracene, 

benzo[k]fluoranthene, benzo[g,h,i]perylene, fluorene, phenanthrene, or pyrene. 

PAHs are generally not produced commercially in the United States except as research 

chemicals. However, PAHs are found in coal, coal tar, and in the creosote oils, oil mists, and 

pitches formed from the distillation of coal tars. The National Institute for Occupational 

Safety and Health (NIOSH) concluded that occupational exposure to coal products can 

increase the risk of lung and skin cancer in workers. NIOSH established a recommended 

occupational exposure limit, time-weighted average (REL-TWA) for coal tar products of 

0.1 milligram of PAHs per cubic meter of air (0.1 mg/m3

) for a 10-hour workday, within a 

40-hour workweek. The American Conference of Governmental Industrial Hygienists 

(ACGIH) recommends an occupational exposure limit for coal tar products of 0.2 mg/m3

 for 

an 8-hour workday, within a 40-hour workweek. The Occupational Safety and Health 

Administration (OSHA) has established a legally enforceable limit of 0.2 mg/m3

 averaged 

over an 8-hour exposure period. 

Mineral oil mists have been given an IARC classification of 1 (sufficient evidence of 

carcinogenicity). The OSHA Permissible Exposure Limit (PEL) for mineral oil mist is 

5 mg/m3 averaged over an 8-hour exposure period. NIOSH has concurred with this limit, and 

has established a recommended occupational exposure limit (REL-TWA) for mineral oil mists 
PAHs 9 

1. PUBLIC HEALTH STATEMENT 

of 5 mg/m3 for a 10-hour work day, 40-hour work week, with a 10 mg/m3 Short Term 

Exposure Limit (STEL). 

More information on rules and standards for exposure to PAHs can be found in Chapter 7. 

1.8 WHERE CAN I GET MORE INFORMATION? 

If you have any more questions or concerns, please contact your community or state health or 

environmental quality department or: 

Agency for Toxic Substances and Disease Registry 

Division of Toxicology 

1600 Clifton Road NE, E-29 

Atlanta, Georgia 30333 

(404) 639-6000 

This agency can also provide you with information on the location of occupational and 

environmental health clinics. These clinics specialize in the recognition, evaluation, and 

treatment of illness resulting from exposure to hazardous substances. 

PAHs 11 

2. HEALTH EFFECTS 

2.1 INTRODUCTION 

The primary purpose of this chapter is to provide public health officials, physicians, toxicologists, and 

other interested individuals and groups with an overall perspective of the toxicology of polycyclic 

aromatic hydrocarbons (PAHs). It contains descriptions and evaluations of toxicological studies and 

epidemiological investigations and provides conclusions, where possible, on the relevance of toxicity 

and toxicokinetic data to public health. 

A glossary and list of acronyms, abbreviations, and symbols can be found at the end of this profile. 

2.2 DISCUSSION OF HEALTH EFFECTS BY ROUTE OF EXPOSURE 

To help public health professionals and others address the needs of persons living or working near 

hazardous waste sites, the information in this section is organized first by route of 

exposure-inhalation, oral, and dermal; and then by health effect-death, systemic, immunological, 

neurological, reproductive, developmental, genotoxic, and carcinogenic effects. These data are 

discussed in terms of three exposure periods-acute (14 days or less), intermediate (15-364 days), and 

chronic (365 days or more). 

Levels of significant exposure for each route and duration are presented in tables and illustrated in 

figures. The points in the figures showing no-observed-adverse-effect levels (NOAELs) or 

lowest-observed-adverse-effect levels (LOAELs) reflect the actual doses (levels of exposure) used in 

the studies. LOAELs have been classified into “less serious” or “serious” effects. “Serious” effects 

are those that evoke failure in a biological system and can lead to morbidity or mortality (e.g., acute 

respiratory distress or death). “Less serious” effects are those that are not expected to cause significant 

dysfunction or death, or those whose significance to the organism is not entirely clear. ATSDR 

acknowledges that a considerable amount of judgment may be required in establishing whether an end 

point should be classified as a NOAEL, “less serious” LOAEL, or “serious” LOAEL, and that in some 

cases, there will be insufficient data to decide whether the effect is indicative of significant 

dysfunction. However, the Agency has established guidelines and policies that are used to classify 

these end points. ATSDR believes that there is sufficient merit in this approach to warrant an attempt 
PAHs 12 

2. HEALTH EFFECTS 

at distinguishing between “less serious” and “serious” effects. The distinction between “less serious” 

effects and “serious” effects is considered to be important because it helps the users of the profiles to 

identify levels of exposure at which major health effects start to appear. LOAELs or NOAELs should 

also help in determining whether or not the effects vary with dose and/or duration, and place into 

perspective the possible significance of these effects to human health. 

The significance of the exposure levels shown in the Levels of Significant Exposure (LSE) tables and 

figures may differ depending on the user’s perspective. Public health officials and others concerned 

with appropriate actions to take at hazardous waste sites may want information on levels of exposure 

associated with more subtle effects in humans or animals (LOAELs) or exposure levels below which 

no adverse effects (NOAELs) have been observed. Estimates of levels posing minimal risk to humans 

(Minimal Risk Levels or MRLs) may be of interest to health professionals and citizens alike. 

Levels of exposure associated with carcinogenic effects (Cancer Effect Levels, CELs) of PAHs are 

indicated in Tables 2-1, 2-2, and 2-3 and Figures 2-l and 2-2. Because cancer effects could occur at 

lower exposure levels, Figure 2-2 also shows a range for the upper bound of estimated excess risks, 

ranging from a risk of 1 in 10,000 to 1 in 10,000,000 (10-4

 to 10-7), as developed by EPA. 

Estimates of exposure levels posing minimal risk to humans (Minimal Risk Levels or MRLs) have 

been made for PAHs. An MRL is defined as an estimate of daily human exposure to a substance that 

is likely to be without an appreciable risk of adverse effects (noncarcinogenic) over a specified 

duration of exposure. MRLs are derived when reliable and sufficient data exist to identify the target 

organ(s) of effect or the most sensitive health effect(s) for a specific duration within a given route of 

exposure. MRLs are based on noncancerous health effects only and do not consider carcinogenic 

effects. MRLs can be derived for acute, intermediate, and chronic duration exposures for inhalation 

and oral routes. Appropriate methodology does not exist to develop MRLs for dermal exposure. 

Although methods have been established to derive these levels (Barnes and Dourson 1988; EPA 1990), 

uncertainties are associated with these techniques. Furthermore, ATSDR acknowledges additional 

uncertainties inherent in the application of the procedures to derive less than lifetime MRLs. As an 

example, acute inhalation MRLs may not be protective for health effects that are delayed in 

development or are acquired following repeated acute insults, such as hypersensitivity reactions, 
PAHs 13 

2. HEALTH EFFECTS 

asthma, or chronic bronchitis. As these kinds of health effects data become available and methods to 

assess levels of significant human exposure improve, these MRLs will be revised. 

A User’s Guide has been provided at the end of this profile (see Appendix A). This guide should aid 

in the interpretation of the tables and figures for Levels of Significant Exposure and the MRLs. 

PAHs are a group of chemicals that are formed during the incomplete burning of coal, oil, gas, wood, 

garbage, or other organic substances, such as tobacco and charbroiled meat. PAHs can either be 

synthetic or occur naturally. Most of these chemicals as individual compounds (i.e., not as part of a 

combustion product) have no known use except for research purposes. A few PAHs are used in 

medicines and to make dyes, plastics, and pesticides. Others are contained in asphalt used in road 

construction. They are found throughout the environment in the air, water, and soil. There are more 

than 100 different PAH compounds and the health effects of the individual PAHs are not exactly alike. 

Fifty-four PAHs have been identified at one or more NPL hazardous waste sites. These 54 are 

acenaphthene, acenaphthylene, 2-acetoaminofluorene, anthracene, 9, 10-anthracenedione, 

benz[a]anthracene, benzo[a]pyrene, benzo[e]pyrene, benzo[a]fluoranthene, benzo[b]fluoranthene, 

benzo[b]fluorene, benzofluoranthene, benzo[j]fluoranthene, benzo[k]fluoranthene, 

benzo[g,h,i]fluoranthene, benzoperylene, benzo[g,h,i]perylene, benzophenanthrene, benzopyrene, 

benzothiophene, benzo[b]thiophene, chrysene, 4H-cyclopenta[d,e,f]phenanthrene, dibenz[a,j]anthracene, 

dibenz[a,h]anthracene, 7,12-dimethylbenz[a]anthracene, 2,7-dimethylbenzo[b]thiophene, 

1,4-dimethoxyanthracene, dimethyl phenanthrene, 2,5dimethyl phenanthrene, 

dodecachlorodecahydrotrim, fluoranthene, fluorene, indeno[ 1,2,3-c,d] pyrene, 

12-methylbenz[a]anthracene, methyl anthracene, 9-methylanthracene, 3-methylcholanthrene, 

methylfluorene, methylphenanthrene, 2-methylphenanthrene, 1-methylphenanthrene, 

4-methylphenanthrene, methylpyrene, phenanthrene, phenanthridine, phenanthroline, pyrene, perylene, 

6,7-tetrahydropyrene, tetramethylphenanthrene, 3,4,5,6-tetramethylphenanthrene, and 

trimethylphenanthrene. 

However, only 17 PAHs are discussed in this profile. These 17 PAHs are: 

• acenaphthene 

• acenaphthylene 

• anthracene 
PAHs 14 

2. HEALTH EFFECTS 

• benz[a]anthracene 

• benzo[a]pyrene 

• benzo[e]pyrene 

• benzo[b]fluoranthene 

• benzo[j]fluoranthene 

• benzo[g,h,i]perylene 

• benzo[k]fluoranthene 

• chrysene 

• dibenz[a,h]anthracene 

• fluoranthene 

• fluorene 

• indeno[l,2,3-c,d]pyrene 

• phenanthrene 

• pyrene 

These 17 PAHs were selected using the following four criteria: 

(1) toxicity 

(2) potential for human exposure 

(3) frequency of occurrence at NPL hazardous waste sites 

(4) extent of information available. 

The 17 PAHs were combined into one profile to avoid repetition across multiple profiles on the 

individual PAHs since these chemicals often occur together in the environment and many have similar 

toxicological effects, environmental fate, etc. Instances in which it is known that the various PAHs 

differ with regard to toxicological effects or environmental fate will be pointed out. For example, 

PAHs can be classified as “alternant” (e.g., benzo[a]pyrene, benz]a]anthracene, chrysene, 

dibenz[a,h]anthracene) or “nonalternant” (e.g., fluoranthene, benzo[k]fluoranthene, 

benzol[j]fluoranthene, indeno[ 1,2,3-c,d]pyrene). This distinction is based on the electron density 

associated with the molecule. Altemant PAHs have an equally distributed electron density, whereas 

nonalternant PAHs behave almost as if they were two different molecules because of an uneven 

distribution of electron density from one portion of the molecule to another. The toxicological 

significance of this difference is that alternant and nonalternant PAHs appear to behave differently, for 

example, with regard to how they are metabolized to ultimate carcinogens (see Section 2.3.3, 

Metabolism). 

Reliable health-based and environmental information exists on only a few of the 17 PAHs discussed in 

this profile, and the potential health effects of the other less well-studied PAHs must be inferred from 

this information. By combining all 17 PAHs in one profile, these comparisons and inferences can 
 

PAHs 15 

2. HEALTH EFFECTS 

easily be made. Although a large toxicity database exists on complex mixtures that contain PAHs 

(such as crude oils, various high boiling point distillates, complex petroleum products, coal tars, 

creosote, and the products of coal liquification processes), these data generally have not been used in 

this profile. It is difficult to ascertain the toxicity of the component PAHs in these mixtures because 

of the potential interactions that could occur and the presence of other toxic substances in the 

mixtures. Furthermore, ATSDR has developed a profile on one of these complex mixtures, creosote, 

and the reader is referred to this profile for information on this complex mixture (ATSDR 1994). 

However, most of the available information on the health effects of PAHs in humans must be inferred 

from studies that reported the effects of exposure to complex mixtures that contain PAHs. Several 

epidemiologic studies have shown increased mortality due to lung cancer in humans exposed to coke 

oven emissions, roofing-tar emissions, and cigarette smoke. Each of these mixtures contains 

benzo[a]pyrene, chrysene, benz[a]anthracene, benzo[b]fluoranthene, and dibenz[a,h]anthracene as well 

as other potentially carcinogenic PAHs and other carcinogenic and potentially carcinogenic chemicals, 

tumor promoters, initiators, and co-carcinogens such as nitrosamines, coal tar pitch, and creosote. It is 

thus impossible to evaluate the contribution of any individual PAH to the total carcinogenicity of these 

mixtures in humans because of the complexity of the mixtures and the presence of other carcinogens. 

Despite these limitations, reports of this nature provide qualitative evidence of the potential for 

mixtures containing PAHs such as benzo[a]pyrene, chrysene, benz[a]anthracene, benzo[b]fluoranthene, 

and dibenz[a,h]anthracene to cause cancer in humans. For this reason, and also because of the lack of 

data on the effects of individual PAHs in humans, such information has been included in this profile on 

PAHs. 

2.2.1 Inhalation Exposure 

2.2.1.1 Death 

No studies were located regarding death in humans following inhalation exposure to any of the 

17 PAHs discussed in this profile. However, a dose-related decrease in survival was noted in hamsters 

after 60 weeks of inhalation exposure to 46.5 mg/m3

 benzo[a]pyrene for 109 weeks (Thyssen et al. 

1981). The authors attributed this reduced survival in part to toxic and carcinogenic effects induced 

by this PAH (e.g., tumors in the pharynx and larynx that could have inhibited food intake). 
PAHs 16 

2. HEALTH EFFECTS 

2.2.1.2 Systemic Effects 

No studies were located regarding cardiovascular, gastrointestinal, hematological, musculoskeletal, 

hepatic, dermal, or ocular effects in humans or animals following inhalation exposure to any of the 

17 PAHs discussed in this profile. The systemic effects observed after inhalation exposure are 

discussed below. 

The highest NOAELs for respiratory and renal effects in each species and duration category are 

recorded in Table 2-1 and plotted in Figure 2-1. 

Respiratory Effects. Only one study was located regarding respiratory effects in humans 

following inhalation exposure to PAHs, specifically, benzo[a]pyrene. The respiratory health of 

667 workers in a rubber factory was investigated (Gupta et al. 1993). Respiratory health was 

evaluated and examined for correlations to length of employment at the factory. In addition, total 

suspended particulate matter and benzo[a]pyrene concentrations were monitored in various parts of the 

factory and examined for possible correlation with the respiratory health of the workers in the same 

area of the factory. Statistically significant decrements in ventilatory function occurred following 

prolonged exposure as assessed by duration of employment. When different sections of the factory 

were considered, workers in the compounding section were the most affected, which was associated 

with the highest exposure to particulate matter and benzo[a]pyrene. Workers in the compounding 

section exhibited radiographic abnormalities including patch opacities, prominent bronchiovascular 

markings, and pleural effusions. Other symptoms included bloody vomit, breathing problems, chest 

pains, chest irritation, throat irritation, and cough. Workers in other areas of the plant exposed to 

lower levels of particulate matter and benzo[a]pyrene were similarly affected although to a lesser 

degree and in fewer numbers (Gupta et al. 1993). No attempt was made to separate the effects of 

exposure to benzo[a]pyrene and particulate matter, or to identify possible simultaneous exposure to 

other toxic chemicals. 

Groups of 40 Fischer-344/Crl rats/sex were exposed nose-only to an aerosol of benzo[a]pyrene 

(7.7 mg/m3

) 2 hours/day, 5 days/week, for 4 weeks (Wolff et al. 1989a). Nasal and left lung sections 

were examined histopathologically. No treatment-related lesions were noted in the lungs or nasal 

cavities of the animals exposed to benzo[a]pyrene. Although this was a well-conducted inhalation 

toxicity study, it is not appropriate for use in risk assessment because only one concentration was 


 

PAHs 19 

2. HEALTH EFFECTS 

studied (thereby precluding the assessment of a dose-response relationship); no adverse 

treatment-related effects were observed; and the only parts of the respiratory tract examined 

histopathologically were the lungs and nose. 

Renal Effects. No studies were located regarding renal effects in humans following inhalation 

exposure to any of the 17 PAHs discussed in this profile. 

Groups of 40 Fischer-344/Crl rats/sex were exposed nose-only to an aerosol of benzo[a]pyrene 

2 hours/day, 5 days/week, for 4 weeks (Wolff et al. 1989a). Kidney sections were examined 

histopathologically. No treatment-related lesions were noted in the kidneys of the animals exposed to 

benzo[a]pyrene. 

2.2.1.3 Immunological and Lymphoreticular Effects 

Humoral immunity was monitored in male iron foundry workers in Poland (Szczeklik et al. 1994). 

Coke oven workers (199) were compared to cold-rolling mill workers (76). The groups were similar 

with respect to age, length of employment, and smoking habits. The results showed that coke oven 

workers, exposed to high concentrations of atmospheric PAHs, including fluoranthene, perylene, 

pyrene, benzo[a]pyrene, chrysene, benz[a]anthracene, dibenz[a,h]anthracene, and benzo[g,h,i]perylene, 

had reduced levels of serum immunoglobins. The workers most exposed to PAHs worked at the 

topside area of the coke ovens. Benzo[a]pyrene exposure was used as a reference point. Coke oven 

workers, exposed to 0.0002-0.50 mg/m3

 benzo[a]pyrene, were compared to cold-rolling mill workers, 

whose exposure to benzo[a]pyrene was 3-5 orders of magnitude less. Average length of employment 

was 15 years. IgG, IgA, IgM, and IgE concentrations were measured. Coke oven workers exhibited a 

marked depression of mean serum IgG and IgA, compared to mill workers. IgM tended to decrease, 

whereas IgE tended to increase in the coke oven workers. The biological significance of this finding 

is unclear and is not addressed by the authors. However, the authors suggest that serum 

immunoglobulin levels may be a useful biomarker for PAH exposure. The authors note, however, that 

the coke oven workers were exposed to higher levels of sulfur dioxide and carbon monoxide than were 

the cold-rolling mill workers, and they suggest that this additional exposure may have potentiated the 

effects of the PAH exposure. The potential contribution of the smoking habits of the subjects was not 

investigated. 
 

 

 

 

PAHs 20 

2. HEALTH EFFECTS 

No studies were located regarding the following effects in humans or animals following inhalation 

exposure to any of the 17 PAHs discussed in this profile: 

2.2.1.4 Neurological Effects 

2.2.1.5 Reproductive Effects 

2.2.1.6 Developmental Effects 

2.2.1.7 Genotoxic Effects 

Becher et al. (1984) evaluated urine and blood samples from 15 aluminum plant workers (average age, 

29 years; average years employed, 3.8) exposed to an estimated total PAH concentration of 1 mg per 

8-hour work shift. The main PAH components identified by air sampling and also detected in the 

urine samples included phenanthrene, fluoranthene, pyrene, benz[a]anthracene, chrysene, 

benzo[e]pyrene, and benzo[a]pyrene. Results of the cytogenetic analysis of peripheral lymphocytes of 

the exposed workers indicated that the frequency of sister chromatid exchange was not influenced by 

the presence of large amounts of PAHs. These findings were reported to be consistent with the 

negligible increase in lung cancer found in epidemiological studies of aluminum workers. The 

investigators, therefore, questioned the relevance of PAH air monitoring as a measure of the 

occupational hazards associated with PAH exposure. Alternatively, it is possible that there are no 

occupational hazards associated with PAH exposure at these levels. Similar results were obtained with 

iron factory workers (length of employment: 2 46 years) exposed to 0.0005-0.00 mg/m3 

benzo[a]pyrene (Perera et al. 1993), who exhibited an increased rate of mutations in peripheral 

lymphocytes that were not correlated with PAH exposure. These authors suggest that both 

biomonitoring and personal monitoring may be necessary to evaluate exposure. 

The high lung cancer rate in Xuan Wei, China, is associated with smokey coal use in unvented homes, 

but not with wood or smokeless coal use (Mumford et al. 1993). Smoky coal combustion emits higher 

PAH concentrations than wood combustion. This study evaluated PAH-DNA adducts in placentas, and 

in peripheral and cord white blood cells (WBC) from Xuan Wei women burning smoky coal or wood 

and from Beijing women using natural gas. Exposures were based on benzo[a]pyrene concentrations 

determined by personal monitors. Women in Xuan Wei burning smoky coal without a chimney were 
PAHs 21 

2. HEALTH EFFECTS 

exposed to 0.383 mg/m3 benzo[a]pyrene, those burning smoky coal with chimneys were exposed to 

0.184 mg/m3, and women burning wood or using natural gas (Beijing) had no detectable exposure to 

benzo[a]pyrene. Positive results (detection of PAH-DNA adducts) were found in 58, 47, and 5% of 

the placentas from Xuan Wei women burning smoky coal without a chimney, with a chimney, and 

Beijing women using natural gas, respectively. Positive results were found in 46, 6.5, 56, and 25% of 

placentas from Xuan Wei women who lived in houses without and with chimneys, Xuan Wei women 

burning wood, and Beijing controls, respectively. Peripheral WBC samples were positive in 7 of 9, 

8 of 9, and 3 of 9 for the Xuan Wei women who lived in houses without and with chimneys and 

Beijing women, respectively. No dose-response relationship was observed between the air 

benzo[a]pyrene concentrations and DNA adduct levels or percentage of detectable samples. However, 

using the fluorescent color assay, there was a significant association between DNA adduct detection in 

the placenta and cooking methods. Moreover, individual comparisons of the data revealed a 

significant difference between both smokey coal groups (chimney, no chimney) and natural gas 

cooking. The results of this study suggest that DNA adducts can be used as a biomarker to assess 

human exposure to combustion emissions. 

Thirty-four workers in an electrode paste plant were monitored for response to exposure (Ovrebo et al. 

1994). Exposure to benzo[a]pyrene was 0.9 μg/m3

; exposure to pyrene was 3.5μg/m3. 

1-Hydroxypyrene was measured in the urine, and PAH-DNA adducts were measured in white blood 

cells to demonstrate their relationship to the exposure. Results from these workers were compared to 

two reference control groups: research and development (R&D) workers and nickel refinery workers. 

Mean values of PAH-DNA adducts in the white blood cells from randomly selected participants in the 

three groups were only marginally different, with the exception of two smokers in the electrode plant, 

who had the highest levels. Mean PAH-DNA adduct levels were 10.9 adducts per 108 nucleotides for 

the electrode workers, 10.8 adducts per 108

 nucleotides for the R&D personnel, and 10.0 adducts per 

108

 nucleotides for nickel plant workers not occupationally exposed to PAHs. No correlation was 

found between PAH-DNA adducts and 1-hydroxypyrene in the urine. 

In an ongoing comprehensive evaluation of biological markers, workers in or near an iron foundry 

with varying exposures to PAHs were examined for response to exposure (Santella et al. 1993). 

Exposure to benzo[a]pyrene, determined by personal monitors, was 2-60 ng/m3, which are the lowest 

levels yet analyzed in foundry workers. 1-Hydroxypyrene was measured in the urine, and PAH-DNA 

adducts were measured in white blood cells to demonstrate their relationship to the exposure. 
 

PAHs 22 

2. HEALTH EFFECTS 

Cigarette smoking, but not age or charbroiled food, influenced the level of 1-hydroxypyrene but not 

PAH-DNA adducts. When workers were classified into three categories of exposure (low, 

<.0005 mg/m3

; medium, 0.0005-0.0012 mg/m3; high, >0.0012 mg/m3), PAH-DNA adducts showed an 

increasing trend, with exposure from 5.2 to 6.2-9.6 adducts per 108 nucleotides in the low-, medium, 

and high-exposure groups, respectively. However, the three exposure groups did not differ 

significantly from each other, and no independent control group was used. 

In order to evaluate the correlation between peripheral blood leukocyte DNA adducts as an indicator of 

exposure to PAHs and the airborne contamination of PAH at the workplaces, a survey of 69 coke oven 

workers was carried out (Assennato et al. 1993). In each workplace, total PAH and specific 

(benz[a]anthracene, benzo[a]pyrene, chrysene) PAH airborne concentrations were measured. Job titles 

included supervisor, door maintenance, machine operator, gas regulators, temperature operators, and 

top side workers. For the workplaces evaluated, the range of airborne concentrations (μg/m3) for 

benz[a]anthracene, benzo[a]pyrene, and chrysene, respectively, were: supervisor (0.41, 0.29, 0.32), 

door maintenance (4.26-14.79, 2.31-6.37, 2.34 6.53), machine operator (0.11-33.19, 0.08-13.17, 

0.03-12.63), gas regulators (0.21-2.10, 0.12-1.61, 0.13-1.60), temperature operators (1.77-10.07, 

1.37-5.03, 0.98 4.78), and top side workers (0.45-3.40, 0.47-4.73, 0.23-2.42). Mean values (fmol/μg 

DNA) for PAH-DNA adducts in leukocytes by job title were: supervisor (0.059), door maintenance 

(0.174) machine operator (0.065), gas regulators (0.081), and temperature operators (0.071). Levels of 

exposure were correlated with PAH-DNA adduct formation. However, the differences were not 

statistically significant. The major limitations of the study included no record of length of exposure, 

no independent control group, no reporting of expected background levels of adducts, and no 

estimation of the length of time individual workers were exposed to particular levels of the PAHs. 

Other genotoxicity studies are discussed in Section 2.4. 

2.2.1.8 Cancer 

No studies were located regarding cancer in humans following inhalation exposure to any of the 

17 PAHs discussed in this profile. However, epidemiologic studies have shown increased mortality 

due to lung cancer in humans exposed to coke oven emissions (Lloyd 1971; Mazumdar et al. 1975; 

Redmond et al. 1976), roofing-tar emissions (Hammond et al. 1976), and cigarette smoke (Maclure 

and MacMahon 1980; Wynder and Hoffmann 1967). Each of these mixtures contains benzo[a]pyrene, 

chrysene, benz[a]anthracene, benzo[b]fluoranthene, and dibenz[a,h]anthracene as well as other 
 

PAHs 23 

2. HEALTH EFFECTS 

potentially carcinogenic PAHs and other carcinogenic and potentially carcinogenic chemicals, tumor
 

promoters, initiators, and co-carcinogens such as nitrosamines, coal tar pitch, and creosote. It is thus
 

impossible to evaluate the contribution of any individual PAH to the total carcinogenicity of these
 

mixtures in humans because of the complexity of the mixtures and the presence of other carcinogens.
 

Furthermore, the levels of individual or total PAHs were not quantified in any of these reports.
 

Despite these limitations, reports of this nature provide qualitative evidence of the potential for
 

mixtures containing PAHs such as benzo[a]pyrene, chrysene, benz[a]anthracene, benzo[b]fluoranthene,
 

and dibenz[a,h]anthracene to cause cancer in humans.
 

Several inhalation studies for animals given benzo[a]pyrene were located. Shulte et al. (1993) found a
 

significant increase in all lung tumors and a dose-dependent increase in malignant lung tumors for
 

mice exposed to PAH-enriched exhausts containing 0.05 or 0.09 mg/m3 benzo[a]pyrene. The chronic
 

study of Thyssen et al. (1981) provides clear-cut evidence of a dose-response relationship between
 

inhaled benzo[a]pyrene particles (99% of the benzo[a]pyrene particles were between 0.2 and
 

0.54 microns in diameter) and respiratory tract tumorigenesis. Respiratory tract tumors were induced 

in the nasal cavity, pharynx, larynx, and trachea in a dose-related manner in hamsters exposed to 

9.5 mg/m3 or 46.5 mg/m3 for 109 weeks. No lung tumors were found, and the reason for the absence 

of lung tumors is not known. Furthermore, the particle sizes were reported to be within the respirable 

range (0.2-0.5 microns in diameter). Tumors were also observed following exposure to 46.5 mg/m3 in 

the esophagus and forestomach (presumably as a consequence of mucocilliary particle clearance) 

(Thyssen et al. 1981). These tumor types consisted of papillomas, papillary polyps, and squamous cell 

carcinomas. 

The CEL from the Thyssen et al. (1981) study is recorded in Table 2-1 and plotted in Figure 2-l. 

2.2.2 Oral Exposure 

2.2.2.1 Death 

No studies were located regarding death in humans after oral exposure to any of the 17 PAHs 

discussed in this profile. 
 

PAHs 24 

2. HEALTH EFFECTS 

Oral exposure to 120 mg/kg/day benzo[a]pyrene has resulted in decreased survival time in two strains 

of mice (DBA/2N and AKR/N) whose hepatic aryl hydrocarbon hydroxylase (AHH) activity is not 

induced by PAHs (“nonresponsive” mice) (Robinson et al. 1975). AHH is a microsomal enzyme 

believed to be responsible for the metabolism of benzo[a]pyrene. All of the mice in the treatment 

group died, with at least half the deaths occurring within 15 days of dosing. Only three mice in the 

control group died. Death appeared to be caused by bone marrow depression (aplastic anemia, 

pancytopenia), leading to hemorrhage or infection. In contrast, only 6 of 90 (7%) mice with inducible 

AHH activity (“responsive” mice) similarly exposed to benzo[a]pyrene died over the same period of 

time. The authors concluded that the decreased survival in the nonresponsive mice was associated 

with a single gene difference encoding aromatic hydrocarbon responsiveness and was dependent on 

route of exposure. Benzo[a]pyrene was not as rapidly metabolized by the liver and excreted following 

oral administration in nonresponsive mice as in responsive mice. Therefore, more benzo[a]pyrene was 

available to reach the target tissue (i.e., bone marrow) in the nonresponsive mice, resulting in bone 

marrow depression and death. 

A LOAEL for death for intermediate-duration exposure in mice is recorded in Table 2-2 and plotted in 

Figure 2-2. 

2.2.2.2 Systemic Effects 

No studies were located regarding respiratory, cardiovascular, hematological, musculoskeletal, hepatic, 

dermal, or ocular effects in humans following oral exposure to any of the 17 PAHs discussed in this 

profile. The systemic effects observed in humans or animals after oral exposure are discussed below. 

The highest NOAEL values and all LOAEL values from each reliable study for each species and 

duration category are recorded in Table 2-2 and plotted in Figure 2-2. 

Respiratory Effects. Male and female mice were exposed to 0, 175, 350, or 700 mg/kg/day 

acenaphthene by gavage for 13 weeks (EPA 1989c). No signs of respiratory distress were seen during 

life for any dose group, and no gross or microscopic damage was seen upon necropsy. Similar 

findings were reported after 13-week administration of 1,000 mg/kg/day anthracene, and 

500 mg/kg/day fluoranthene, or 500 mg/kg/day fluorene (EPA 1988e, 1989d, 1989e). 










PAHs 35 

2. HEALTH EFFECTS 

Cardiovascular Effects. Male and female mice were exposed to 0, 175, 350, or 700 mg/kg/day 

acenaphthene by gavage for 13 weeks (EPA 1989c). No signs of cardiovascular distress were seen 

during life for any dose group, and no gross or microscopic damage was seen upon necropsy. Similar 

findings were reported after 13-week administration of 1,000 mg/kg/day anthracene, and 

500 mg/kg/day fluoranthene, or 500 mg/kg/day fluorene (EPA 1988e, 1989d, 1989e). 

Gastrointestinal Effects. Minimal information is available on the gastrointestinal effects of 

human oral exposure to PAHs. In one study, humans that consumed anthracene-containing laxatives 

(the anthracene concentration was not specified) for prolonged periods of time were found to have an 

increased incidence of melanosis of the colon and rectum (i.e., unusual deposits of black pigments in 

the colon and rectum) compared to patients who did not consume anthracene laxatives. However, no 

definitive conclusions can be drawn from these results because of study limitations that include 

possible misclassification of patients with respect to the level of anthracene laxative use over 30 years 

and no accounting for other factors involved in the pathogenesis of melanosis (Badiali et al. 1985). 

Enzyme alterations in the mucosa of the gastrointestinal tract have been observed in animals acutely 

exposed to anthracene, benz[a]anthracene, benzo[a]pyrene, or phenanthrene. In rats, acute intragastric 

administration of 50 or 150 mg/kg/day benz[a]anthracene or benzo[a]pyrene, respectively, for 4 days 

resulted in suppression of carboxylesterase activity in the intestinal mucosa (reduction of activity by 

30% and 44%, respectively); rats exposed to 100 mg/kg/day of anthracene or phenanthrene exhibited 

carboxylesterase activity that was increased by 13% and 30%, respectively (Nousiainen et al. 1984). 

Enzyme alteration in the absence of other signs of gastrointestinal toxicity is not considered an adverse 

health effect, but it may precede the onset of more serious effects. Based on this very limited 

information, it would appear that acute ingestion of anthracene, benz[a]anthracene, benzo[a]pyrene, or 

phenanthrene at these doses may not adversely affect the gastrointestinal tract of animals; however, 

exposed animals exhibited biochemical changes and it is possible that more serious effects could occur 

at high doses. 

Male and female mice were exposed to 0, 175, 350, or 700 mg/kg/day acenaphthene by gavage for 

13 weeks (EPA 1989c). No adverse effects on the gastrointestinal system were seen during life for 

any dose group, and no gross or microscopic damage was seen upon necropsy. Similar findings were 

reported after 13-week administration of 1,000 mg/kg/day anthracene, 500 mg/kg/day fluoranthene, or 

mg/kg/day fluorene (EPA 1988e, 1989d, 1989e). 
PAHs 36 

2. HEALTH EFFECTS 

Hematological Effects. Male and female mice were exposed to 0, 175, 350, or 700 mg/kg/day 

acenaphthene by gavage for 13 weeks (EPA 1989 ). No hematological effects were seen during life 

for any dose group, and no gross or microscopic damage was seen upon necropsy. Similar findings 

were reported after 13-week administration of 1,000 mg/kg/day anthracene (EPA 1989d). 

Administration of 250 mg/kg/day fluoranthene by gavage for 13 weeks to mice resulted in decreased 

packed cell volume in females, but not in males, given doses up to 500 mg/kg/day (EPA 1988e). Both 

male and female mice exposed to 250 mg/kg/day fluorene exhibited hematologic effects, including 

decreased packed cell volume and hemoglobin content (EPA 1989e). 

Adverse hematopoietic effects (e.g., aplastic anemia, pancytopenia) that ultimately led to death were 

reported in the Ah-nonresponsive strains of mice, DBA/2N and AKR/N, following oral exposure to 

120 mg benzo[a]pyrene/kg/day for 180 days. Death was attributed to hemorrhage or infection that 

resulted from pancytopenia (Robinson et al. 1975). Similar results were obtained by Legraverend et 

al. (1983). The Ah gene encodes a cytosolic receptor (Ah receptor) that regulates the induction of the 

cytochrome P-450 enzymes. Differences in this gene locus determine whether the Ah receptor will be 

“high-affinity” (i.e., will allow for the induction of the cytochrome P-450 enzymes [more specifically, 

AHH] and is found in responsive mice) or “low-affinity” (i.e., does not allow for the induction of the 

AHH and is found in nonresponsive mice). Mice with a high-affinity Ah receptor (i.e., responsive 

mice) were administered 120 mg/kg/day benzo[a]pyrene in the diet for 3 weeks and exhibited no 

myelotoxicity. However, all nonresponsive mice that were treated according to the same regimen died 

from myelotoxic effects within 3 weeks (Legraverend et al. 1983). These results support the results of 

Robinson et al. (1975). 

Musculoskeletal Effects. Male and female mice were exposed to 0, 175, 350, or 700 mg/kg/day 

acenaphthene by gavage for 13 weeks (EPA 1989c). No signs of musculoskeletal effects were seen 

during life for any dose group, and no gross or microscopic damage was seen upon necropsy. Similar 

findings were reported after 13-week administration of 1,000 mg/kg/day anthracene, 500 mg/kg/day 

fluoranthene, or 500 mg/kg/day fluorene (EPA 1988e, 1989d, 1989e). 

Hepatic Effects. The induction of foci of altered hepatocytes is often seen in rats and mice that 

also develop liver tumors. These foci have altered enzyme activities and higher rates of cell 

proliferation than normal hepatocytes. A l-day intragastric administration of 200 mg/kg of 

benzo[a]pyrene or dibenz[a,h]anthracene, or of 180 mg/kg benz[a]anthracene to rats was followed by a 
PAHs 37 

2. HEALTH EFFECTS 

diet containing 2-acetylaminofluorene (2-AAF) and carbon tetrachloride induced gamma-GT foci 

(Tsuda and Farber 1980). Partially hepatectomized rats and sham hepatectomized rats were used, to 

provide proliferating and non-proliferating hepatocytes, respectively. Partially hepatectomized rats 

were more responsive to treatment than the sham-operated animals. For partially hepatectomized rats, 

benzo[a]pyrene was a more potent foci inducer than either benz[a]anthracene or dibenz[a,h]anthracene. 

Increased relative liver weight was seen in male mice and increased absolute and relative liver weight 

was seen in female mice given 175 mg/kg/day acenaphthene daily by gavage for 13 weeks; these 

effects were unaccompanied by other hepatic effects (EPA 1989c). Increased absolute and relative 

liver weight correlated with hepatocellular hypertrophy was seen in male and female mice given 

350 mg/kg/day acenaphthene daily by gavage for 13 weeks (EPA 1989c). Increased serum cholesterol 

was also seen in females receiving 350 mg/kg/day acenaphthene (EPA 1989c). Increased liver weight 

and dose-related centrilobular pigmentation accompanied by an increase in liver enzymes were 

observed in both male and female mice receiving 250 mg/kg/day fluoranthene by gavage for 13 weeks 

(EPA 1988e). Male mice exposed to 125 mg/kg/day fluoranthene exhibited a slight increase in 

centrilobular pigmentation, and an increase in relative liver weight (EPA 1988e). Increased relative 

liver weight was observed in all treated groups, whereas increased absolute and relative liver weight 

was observed in the mid- and high-dose animals receiving 0, 125, 250, and 500 mg/kg/day fluorene 

for 13 weeks (EPA 1989e). However, there were no accompanying histopathological changes. No 

statistically significant effects of treatment were reported after 13-week administration of 

1,000 mg/kg/day anthracene (EPA 1989d) 

The ability to induce aldehyde dehydrogenase (ADH) in animals has been correlated with carcinogenic 

potency. Rats that were intragastrically administered 100 mg/kg/day of benzo[a]pyrene, 

benz[a]anthracene, anthracene, chrysene, or phenanthrene for 4 days exhibited cytosolic ADH 

induction (Torronen et al. 1981). However, benzo[a]pyrene and benz[a]anthracene were much more 

effective than phenanthrene, chrysene, or anthracene. Exposure to benzo[a]pyrene and 

benz[a]anthracene also increased the relative liver weights by 27% and 19%, respectively (Torronen et 

al. 1981). The authors concluded that anthracene, phenanthrene, and chrysene, which have been 

characterized as either noncarcinogens or equivocal carcinogens (see Section 2.2.2.8), are poor ADH 

inducers (Torronen et al. 1981). 

The induction of carboxylesterase activity has also been observed in animals exposed to PAHs 

(Nousiainen et al. 1984). Benzo[a]pyrene, benz[a]anthracene, and chrysene were moderate inducers of 
PAHs 38 

2. HEALTH EFFECTS 

hepatic carboxylesterase activity in rats that were intragastrically administered 50, 100, and 

150 mg/kg/day (100 mg/kg/day for chrysene), respectively, for 4 days. However, rats administered 

100 mg/kg/day anthracene or phenanthrene did not exhibit induction of hepatic carboxylesterase 

activity. Induction of hepatic microsomal enzymes generally results in enhanced biotransformation of 

other xenobiotics (to either more or less toxic forms). 

Increases in liver weight following partial hepatectomy have also been examined following acute oral 

exposure to various PAHs. Partially hepatectomized rats were fed diets containing various PAHs for 

10 days. Administration of 51.4 mg/kg/day acenaphthene or 180 mg/kg/day fluorene resulted in 

statistically significant increases in liver weight compared to controls, which may have indicated an 

effect on regeneration, although rates of cell proliferation were not determined. Administration of 

15.4 mg/kg/day acenaphthene, 51.4 mg/kg/day benzo[a]pyrene, or 51.4 mg/kg/day pyrene, anthracene, 

or phenanthracene had no effect. Diets containing 51.4 mg/kg/day acenaphthene or 

dibenz[a,h]anthracene, 180 mg/kg/day anthracene or phenanthracene, or 437 mg/kg/day pyrene 

produced no increase in the liver-to-body-weight ratio. Rats that were fed a diet containing 

514 mg/kg/day chrysene exhibited equivocal results: in one trial, a significant increase in liver weight 

gain was noted, while in another trial, no increase in. liver-to-body-weight ratio was observed 

(Gershbein 1975). Thus, both suspected carcinogenic and noncarcinogenic PAHs can affect liver 

weights, although much higher doses are required for noncarcinogenic PAHs. The livers of rats 

administered single doses of ,fluorene by gavage in dimethyl sulfoxide (DMSO) were evaluated for the 

promotion of growth (i.e., cell proliferation as determined by organ weight and mitotic index) (Danz et 

al. 1991). The authors claimed that liver weight was increased in a dose-dependent manner to 20% 

over control values, and that the mitotic index of the hepatocytes was increased by 6-fold after 

48 hours. However, the organ weight data were not presented, and the mitotic index data presented 

graphically in the text do not indicate a 6-fold increase over controls. 

Ah-responsive strains of mice (C57BL/6, C3H/HeN, BALB/cAnN) that were orally administered 

120 mg benzo[a]pyrene/kg/day in their diet for 180 days exhibited a 13% increase in relative liver 

weights (Robinson et al. 1975). 

The hepatic effects observed in animals following oral exposure to PAHs are generally not considered 

serious. However, the enzyme alterations, gamma-GT foci induction, liver regeneration, and increased 

liver weight may precede the onset of more serious hepatic effects. 
PAHs 39 

2. HEALTH EFFECTS 

Renal Effects. The kidney microsomal carboxylesterase activity of rats was moderately induced by 

50-150 mg/kg of benzo[a]pyrene following 4 days of intragastric administration; however, rats 

administered 100 mg/kg/day of anthracene or phenanthrene and 50-150 mg/kg benz[a]anthracene did 

not exhibit increased activity. The authors conclude that anthracene, phenanthrene, and 

benz[a]anthracene are not inducers of kidney carboxylesterase activity (Nousiainen et al. 1984). 

Enzyme induction is considered an adverse effect when observed concurrently with more serious 

effects such as impaired renal function and/or histopathological changes of the kidney. 

Increasing dietary doses of pyrene ranging from 1,000 mg/kg food (127 mg/kg/day) up to 

25,000 mg/kg food (917 mg/kg/day) for a mean dose of 426.6 mg/kg/day over a 25-day study 

produced dilation of the renal tubules in an unspecified number of mice. This effect was not observed 

until the highest dose was administered (Rigdon and Giannukos 1964). The limitations of this study 

(e.g., doses changed throughout exposure period and no statistical analyses performed) render these 

results of questionable toxicological significance. 

Male and female mice were exposed to 0, 175, 350, or 700 mg/kg/day acenaphthene by gavage for 

13 weeks (EPA 1989c). No signs of renal toxicity were seen during life for any dose group, and no 

gross or microscopic damage was seen upon necropsy. Similar findings were reported after 13-week 

administration of 1,000 mg/kg/day anthracene (EPA 1989d). Increased absolute and relative kidney 

weight was observed in males, but not females receiving 500 mg/kg/day fluorene for 13 weeks (EPA 

1989e). Renal tubular regeneration, and interstitial lymphocytic infiltrates and/or fibrosis were 

observed after 13-week oral administration of fluoranthene to female mice at 250 mg/kg/day, and male 

mice at 500 mg/kg/day (EPA 1988e). 

Endocrine Effects. Male and female mice were exposed to 0, 175, 350, or 700 mg/kg/day 

acenaphthene by gavage for 13 weeks (EPA 1989c). No signs of endocrine imbalance were seen 

during life for any dose group, and no gross or microscopic damage was seen upon necropsy. Similar 

findings were reported after 13-week administration of 1,000 mg/kg/day anthracene, 500 mg/kg/day 

fluoranthene, or 500 mg/kg/day fluorene (EPA 1988e, 1989d, 1989e). 

Dermal Effects. Male and female mice were exposed to 0, 175, 350, or 700 mg/kg/day 

acenaphthene by gavage for 13 weeks (EPA 1989c). No signs of derrnal effects were seen during life 

for any dose group, and no gross or microscopic damage was seen upon necropsy. Similar findings 
PAHs 40 

2. HEALTH EFFECTS 

were reported after 13-week administration of 1,000 mg/kg/day anthracene, 500 mg/kg/day 

fluoranthene, or 500 mg/kg/day fluorene (EPA 1988e, 1989d, 1989e). 

Ocular Effects. Male and female mice were exposed to 0, 17.5, 350, or 700 mg/kg/day 

acenaphthene by gavage for 13 weeks (EPA 1989c). No signs of ocular toxicity were seen during life 

for any dose group, and no gross or microscopic damage was seen upon necropsy. Similar findings 

were reported after 13-week administration of 1,000 mg/kg/day anthracene, 500 mg/kg/day 

fluoranthene, or 500 mg/kg/day fluorene (EPA 1988e, 1989d, 1989e). 

Body Weight Effects. Male and female mice were exposed to 0, 175, 350, or 700 mg/kg/day 

acenaphthene by gavage for 13 weeks (EPA 1989c). No adverse effects on body weight were seen 

during life or upon necropsy. Similar findings were reported after 13-week administration of 

1,000 mg/kg/day anthracene, and 500 mg/kg/day fluoranthene (EPA 1988e, 1989d). After 

administration of 500 mg/kg/day fluorene for 13 weeks; however, female mice exhibited increased 

body weight, although male mice showed no effect at the same dose level (EPA 1989e). 

Other Systemic Effects. The number of thymic glucocorticoid receptors in 6-week-old rats 

treated once with 2 mg/kg benzo[a]pyrene was measured (Csaba et al. 1991). It is assumed that 

administration was by oral gavage, but this was never explicitly stated. The number of these receptors 

was decreased by 40% in females and unaffected in males relative to the vehicle control animals. The 

statistical significance of these effects was not indicated, nor was the functional consequences of a 

decrease in receptor number assessed by examination of functional parameters. 

2.2.2.3 Immunological and Lymphoreticular Effects 

No studies were located regarding immunological effects in humans following oral exposure to any of 

the 17 PAHs discussed in this profile. 

A single gavage dose of 150 mg/kg fluorene to male Sprague-Dawley rats had no effect on thymus or 

spleen weight (Danz and Brauer 1988). Little useful information can be obtained from this study as 

only one dose was tested (thereby precluding assessment of the validity of the negative response) and 

no tests of immune function were evaluated. Male and female mice exposed to 0, 175, 350, or 

700 mg/kg/day acenaphthene by gavage for 13 weeks showed no effect of treatment on splenic weight 
PAHs 41 

2. HEALTH EFFECTS 

or histopathology (EPA 1989c). Similar findings were reported after 13-week administration of 

1,000 mg/kg/day anthracene, and 500 mg/kg/day fluoranthene (EPA 1988e, 1989d). After 

administration of 2.50 mg/kg/day fluorene for 13 weeks, however, increased absolute and relative 

spleen weight was seen in both sexes (EPA 1989e). 

Lee and Strickland (1993) looked for antibodies specific to PAH-DNA adducts in the serum of 

BALB/c mice treated orally twice per week for 8 weeks with 0.5 or 5 mg/kg benz[a]anthracene, 

benzo[a]pyrene, benzo[b]fluoranthene, chrysene, dibenz[a,h]anthracene, or fluoranthene. Increased 

antibody response was noted in animals treated with the low dose of benz[a]anthracene and 

benzo[b]fluoranthene, but not any of the other PAHs. 

2.2.2.4 Neurological Effects 

No studies were located regarding neurological effects in humans following oral exposure to any of the 

17 PAHs discussed in this profile. 

Male and female mice exposed to 0, 175, 350, or 700 mg/kg/day acenaphthene by gavage for 

13 weeks showed no effect of treatment on behavior, or histopathologic effects on nerve or brain 

samples (EPA 1989c). Similar findings were reported after 13-week administration of 

1,000 mg/kg/day anthracene, and 500 mg/kg/day fluoranthene (EPA 1988e, 1989d). After 

administration of 500 mg/kg/day fluorene for 13 weeks, however, increased brain weight was observed 

in females, but not in males (EPA 1989e). No histopathologic changes were observed. 

2.2.2.5 Reproductive Effects 

No studies were located regarding reproductive effects in humans following oral exposure to the PAHs 

discussed in this profile. Three animal studies were located that evaluated the reproductive effects of 

benzo[a]pyrene in animals. The results of two oral studies in mice (Mackenzie and Angevine 1981; 

Rigdon and Neal 1965) and one in rats (Rigdon and Rennels 1964) indicate that benzo[a]pyrene 

induces reproductive toxicity in animals. The incidence and severity of these effects depends on the 

strain, method of administration, and dose levels used. In one study, benzo[a]pyrene administered by 

gavage to pregnant CD-l mice decreased the percentage of pregnant females that reached parturition 

and produced a high incidence of sterility in the progeny (Mackenzie and Angevine 1981). In 
PAHs 42 

2. HEALTH EFFECTS 

contrast, benzo[a]pyrene administered in the diet caused no adverse effects on fertility of Swiss mice 

(Rigdon and Neal 1965) but reduced the incidence of pregnancy in female rats (Rigdon and Rennels 

1964). Based on these studies, the LOAEL for benzo[a]pyrene-induced reproductive toxicity in 

parental mice was 160 mg/kg/day, and the LOAEL for these effects in the progeny of exposed animals 

was 10 mg/kg/day (Mackenzie and Angevine 1981). Because only the parental doses are quantifiable, 

these are the only data presented in Table 2-2. 

When CD-l mice were administered benzo[a]pyrene by gavage daily for 10 days during gestation, 

there was a significant reduction in the percentage of pregnant females to reach parturition at 

160 mg/kg/day, the highest dose tested (Mackenzie and Angevine 1981). When F1 progeny were bred 

with untreated animals, the fertility index decreased significantly in all treatment groups. At 

10 mg/kg/day, the lowest dose tested, the reduced fertility noted was associated with significant 

alterations in gonadal morphology and germ cell development. The treatment at higher doses resulted 

in total sterility. Contrary to these results, no adverse effects on reproduction were observed in Swiss 

mice fed benzo[a]pyrene in the diet at ≤133 mg/kg/day over varying time spans during mating, 

gestation, and parturition (Rigdon and Neal 1965). The apparent discrepancy in the results of the two 

studies may be attributable to the method of benzo[a]pyrene administration and metabolic differences 

in the two strains of mice used. 

Dietary administration of benzo[a]pyrene for 28 days revealed no treatment-related effects on the 

estrous cycle of female rats. These rats experienced no significant adverse effects on their fertility 

when bred to untreated male rats (Rigdon and Rennels 1964). In another series of experiments, when 

benzo[a]pyrene-fed male and female rats were bred, only two of seven females became pregnant (as 

compared to 3 of 6 controls); the offspring of one rat were stillborn while those of others were 

resorbed (Rigdon and Rennels 1964). Although the data suggest that benzo[a]pyrene may induce 

reproductive toxicity in rats, they are inconclusive because of the use of a single dose level, small 

number of animals, and inadequate reporting of data. 

Male mice exposed to 0, 175, 350, or 700 mg/kg/day acenaphthene by gavage for 13 weeks showed 

no effect of treatment on reproductive organ weight or histology (EPA 1989c). Female mice, 

however, exhibited decreased ovary weights correlated with an increase of inactivity of the ovary and 

uterus (EPA 1989c). No adverse effects on reproductive organs were reported after 13-week 
PAHs 43 

2. HEALTH EFFECTS 

administration of 1,000 mg/kg/day anthracene, 500 mg/kg/day fluoranthene, or 500 mg/kg/day fluorene 

to male and female mice (EPA 1988e, 1989d, 1989e). 

The available information from animal studies suggests that benzo[a]pyrene may have the potential to 

produce adverse reproductive effects in exposed humans. The highest NOAEL and all LOAEL values 

from each reliable study for reproductive effects following acute- and intermediate-duration exposures 

are reported in Table 2-2 and plotted in Figure 2-2. 

2.2.2.6 Developmental Effects 

No studies were located regarding developmental effects in humans following oral exposure to PAHs. 

Three animal studies were reviewed that assessed developmental effects of benzo[a]pyrene in inbred 

strains of rats and mice. The data from these studies indicate that prenatal exposure to benzo[a]pyrene 

produced reduced mean pup weight during postnatal development and caused a high incidence of 

sterility in the F1 progeny of mice (Mackenzie and Angevine 1981). Using Ah-responsive and 

Ah-nonresponsive strains of mice, the increased incidences of stillboms, resorptions, and 

malformations observed correlated with the maternal and/or embryonal genotype (Legraverend et al. 

1984). In another study, negative results were obtained when benzo[a]pyrene was administered to 

Swiss (responsive) mice (Rigdon and Neal 1965). 

Benzo[a]pyrene was administered by gavage to pregnant CD-l mice during gestation at doses of 10, 

40, and 160 mg/kg/day. The viability of litters at parturition was significantly reduced in the highest 

dose group (Mackenzie and Angevine 1981). The mean pup weight was significantly reduced in all 

treatment groups by 42 days of age. The F1 progeny that were exposed prenatally to benzo[a]pyrene 

(10, 40, and 160 mg/kg/day) were bred with untreated animals and further studied for postnatal 

development and reproductive function. The F1 progeny from the 10-mg/kg/day group experienced 

decreased fertility with associated alterations in gonadal morphology and germ-cell development. 

Because only the parental doses are quantifiable, these are the only data presented in Table 2-2. 

Therefore, the LOAEL of 10 mg/kg/day noted in the F1 progeny discussed above is not presented in 

Table 2-2. Higher doses produced total sterility. This study provides good evidence for the 

occurrence of developmental effects following in utero exposure to benzo[a]pyrene. 
PAHs 44 

2. HEALTH EFFECTS 

The effect of genetic differences in metabolism of orally administered benzo[a]pyrene on in utero 

toxicity and teratogenicity was evaluated in mice that either metabolize benzo[a]pyrene readily 

(Ah-responsive) or not (Ah-nonresponsive) (Legraverend et al. 1984). Pregnant mice, either B6AKFl 

(Ah-responsive) or AKR/J (Ah-nonresponsive), were fasted prior to a diet containing 120 mg/kg/day 

benzo[a]pyrene on days 2-10 of gestation. The mice were killed on day 18 of gestation. On 

day 16 of gestation, intraperitoneal injections of napthoflavone were administered to distinguish 

between fetuses with different Ah-genotypes (Ahb/Ahd and Ahd/Ahd). Oral administration of 

benzo[a]pyrene to the pregnant AKR/J mice (non-responsive) caused more stillbirths, decreased weight 

gain, resorptions, and birth defects among Ahd/Ahd (Ah-nonresponsive) than among Ahb/Ahd 

(Ah-responsive) embryos. However, no differences in in utero toxicity or teratogenicity were observed 

in Ah-genetically different embryos (Ahd/Ahd and Ahb/Ahd) of B6AKFl mothers (responsive). The 

authors concluded that differences in in utero toxicity and teratogenicity are specific to the route of 

administration and can be attributed to “first pass” liver metabolism occurring with oral dosing. They 

also concluded that in utero toxicity and teratogenicity are directly related to the maternal and/or 

embryonal genotype controlled by the Ah-locus; that is, both maternal metabolism as well as target 

organ metabolism (embryo/fetus) were important in determining susceptibility to developmental 

toxicity. Specifically, metabolism by a responsive mother reduces in utero toxic effects in the fetus. 

Similarly, responsive fetuses in the uterus of a non-responsive mother show fewer in utero toxic 

effects. Non-responsive fetuses in the uterus of a non-responsive mother show the highest incidence 

of in utero toxic effects. Although the study emphasizes the importance of administrative route in 

benzo[a]pyrene metabolism and resulting toxicity, it had the following limitations: 1) only one dose 

was evaluated; 2) no quantitative comparisons between treated groups and corresponding control 

animals were presented for any of the reported in utero toxicity or teratogenic effects; 3) small sample 

size; 4) purity of benzo[a]pyrene was not specified. 

In another study, negative results were obtained when the potential developmental effects of 

benzo[a]pyrene were studied in mice (Rigdon and Neal 1965). Dietary administration of this chemical 

to mice at concentrations equivalent to 33.3, 66.7, or 133.3 mg/kg/day at various times before and 

after mating elicited no adverse effects on the developing embryos. Maternal weight gain was reduced 

in the mice administered the higher levels of benzo[a]pyrene, but this effect was reversed when the 

animals were changed to the control ration. Limitations of this study that preclude its inclusion in 

Table 2-2 consist of the use of an inconsistent protocol, varying number of animals, and varying time 

of gestation exposure. 
PAHs 45 

2. HEALTH EFFECTS 

The highest NOAEL values and all LOAEL values from each reliable study for developmental effects 

in mice for acute-duration exposure is recorded in Table 2-2 and plotted in Figure 2-2. 

2.2.2.7 Genotoxic Effects 

No studies were located regarding genotoxic effects in humans following oral exposure to any of the 

17 PAHs discussed in this profile. 

Pregnant Erythrocebus patas monkeys were treated once on gestation day (Gd) 50, 100, or 

150 (term = 160 days) with 5-50 mg/kg benzo[a]pyrene (Lu et al. 1993). Fetuses were removed by 

Cesarean section l-50 days after treatment and fetal organs, placentae, and maternal livers were 

assayed for DNA adducts. Benzo[a]pyrene-DNA adducts were high in fetal organs, placentae, and 

maternal livers in all three trimesters of gestation. Adduct levels were higher in mid-gestation 

compared to early or late gestation. dG-N2-BPDE was the major adduct detected. The adduct levels 

in fetal tissues increased with benzo[a]pyrene dose, but at a much lower rate that placentae or maternal 

livers. Preference in binding to DNA of various fetal tissues was more apparent in early gestation 

compared to late gestation, and at lower doses compared to higher doses. During early gestation and 

at lower doses, benzo[a]pyrene produced a similar level of DNA binding in fetal lung, liver, maternal 

liver, and placenta. Individual fetal organ adduct levels correlated significantly with placental adduct 

levels, indicating placental and/or maternal contribution to adduct formation in fetuses. Evidence of 

fetal contribution to adduct formation was also found. DNA adduct levels in fetal skin were lowest of 

all fetal organs tested and less affected by gestational stage at time of treatment. In contrast, DNA 

adduct levels in fetal liver exhibited distinct gestation stage specificity with higher adduct, levels 

attained during mid-gestation compared to other stages of gestation. Adduct levels decreased at a 

much faster rate during the first 10-15 days compared to 15-50 days after treatment. However, 10% 

of the DNA adducts persisted 50 days after treatment in all organs studied. Together, the results 

suggest that placental adduction accurately indicates fetal exposure. 

Male B6C3F1 mice were fed 0, 0.325, 0.1825, 1.625, 3.25, or 6.5 mg/kg/day benzo[a]pyrene for 

21 days (Culp and Beland 1994). Animals were killed and the liver, lung, and forestomach DNA 

extracted and analyzed for benzo[a]pyrene-DNA adducts. The major adduct, dG-N2-BPDE, was 

quantified. Adduct levels in liver and lung increased in a linear manner. Adduct levels in the 

forestomach appeared to plateau at the highest dose. At doses below the highest, adduct levels were in 
PAHs 46 

2. HEALTH EFFECTS 

the order of forestomach > liver > lung, with the values of average slopes being 3.0 + 0.59, 

2.1 + 0.17, 1.3 + 0.37 fmol adduct/mg DNA/μg benzo[a]pyrene/day, respectively. At these doses, the 

lung and the forestomach were not significantly different. At the high dose, liver > forestomach > 

lung, and each tissue was significantly different from the other. 

DNA binding of coal tar components in male mice was investigated following the ingestion of coal tar 

obtained from a manufactured coal plant (Weyand et al. 1991). One of four different samples (A-D) 

of coal tar or a mixture of four equal portions of the four samples was administered in a gel diet 

which contained 0.25% coal tar. The coal tar contained phenanthrene, fluoranthene, pyrene, 

benz[a]anthracene, chrysene, benzo[b]fluoranthene, benzo[k]fluoranthene, benzo[a]pyrene, 

indeno[ 1,2,3-c,d]pyrene, and benzo[g,h,i]perylene, among other PAHs. In addition, a diet containing 

benzo[a]pyrene at the same level as the 0.25% diet prepared with Sample C was administered; animals 

consuming this diet ingested 0.01-0.02 mg benzo[a]pyrene per day. The diets were administered for 

15 days. Chemical-DNA adduct formation was evaluated in animals following 14 days of treatment. 

Chemical-DNA adduct formation was also evaluated in animals maintained on a 0.1, 0.2, 0.5, and 

1.0% coal tar diet prepared with one of the coal tar samples (C). Chemical-DNA adduct formation in 

animals dosed with 0.l-1.0% Sample C indicated a dose-related effect in lung DNA adduct formation, 

but no dose-related effect was observed for forestomach tissue. In addition, overall adduct levels in 

lung tissue were considerably higher than forestomach levels for animals on the 0.5 or 1% diet. In 

contrast, adduct levels were highest in the forestomach of animals on diets lower in coal tar content 

(0.1 or 0.2%). Chemical-DNA adducts of coal tar components evaluated for Samples A-D and the 

mixture of the four coal tar samples at 0.25% in the diet administered for 15 days indicated adducts in 

the lung, liver, and spleen of all animals. Adduct patterns were similar, but quantitative differences 

were observed between coal tar samples and tissue sites. The highest adduct levels were detected in 

lung DNA. Adduct formation in animals fed the benzo[a]pyrene diet, could not account for the 

differences in the adduct levels observed in animals given the mixtures. Also, adduct formation in 

animals fed the coal tar mixtures correlated with benzo[a]pyrene content in the coal tar, indicating the 

adducts arose from a variety of PAHs in the coal tar mixtures. The levels of 1-hydroxypyrene in the 

urine of these animals correlated with the pyrene content of these coal tars. 

The DNA binding of manufactured gas plant residue (MGP) components in male B6C3Fl mice was 

investigated following oral administration (Weyand and Wu 1994). Male mice were fed a gel diet 

containing manufactured gas plant (MGP) residue (coal tar) at 0.3% for 28 days, or the corresponding 
PAHs 47 

2. HEALTH EFFECTS 

control diet. Two mixtures of MGP residue were used: Mix of 3 combining equal amounts of 

samples from three different MGP plant sites, and Mix of 7 combining equal amounts of samples from 

seven different MGP plant sites, including those used in the Mix of 3. The mixtures contained pyrene, 

benz[a]anthracene, chrysene, benzo[b]fluorene, benzo[k]fluorene, benzo[a]pyrene, 

indeno[l,2,3-cd] pyrene, dibenz[a,h]anthracene, benzo[g,h,i]perylene. Data was presented in terms of 

pyrene consumed. Animals were sacrificed on the twenty-ninth day and lung and forestomach were 

excised and DNA isolated. Chemical-DNA adduct formation was evaluated. Ingestion of the 

adulterated diets resulted in a relatively low level of DNA adducts in the forestomach in comparison 

with the lung (one-tenth the level). PAH-DNA adduct levels in the lung of mice maintained on the 

Mix of 3 (1.4 mg/kg/day pyrene) were two times greater than the level induced by the Mix of 7 

(1.2 mg/kg/day pyrene) suggesting that the composition of the MGP residue may have influence PAH 

absorption or DNA adduct formation. 

Oral exposure to a total dose of 10 mg/kg benzo[a]pyrene produced gene mutations in the mouse coat 

color spot test (Davidson and Dawson 1976, 1977). Dose-related increases in the frequency of 

micronuclei were seen in bone marrow cells harvested from MS/Ae and CD-l male mice (four 

mice/strain/dose) 48 hours after administration of a single oral dose of benzo[a]pyrene ranging from 

62.5 to 500 mg/kg (Awogi and Sato 1989). Although the response appeared to be stronger in the 

MS/Ae strain, the reduction in polychromatic erythrocytes, indicative of target cell toxicity at all levels 

in the CD-1 strain, limited the comparative evaluation of strain specificity. 

In another study, a dose of benzo[a]pyrene (150 mg/kg) known to induce a clastogenic response was 

orally administered to groups of five adult males and females, pregnant females, and fetal ICR mice. 

An increased incidence of micronuclei in bone marrow cells harvested from the various groups of 

adult animals and also in the livers of the fetuses was observed (Harper et al. 1989). Genetic damage 

was most severe in the fetuses. The approximately 7-fold increase in micronuclei in fetal livers as 

compared to maternal bone marrow suggests that the transplacentally-induced genotoxicity was 

probably associated with the immature detoxification processes of fetal liver as compared to adult bone 

marrow. It would, nevertheless, appear that the fetus may be at an increased risk. 

Data showing that orally administered benzo[a]pyrene induces micronuclei were confirmed in 

subsequent studies (Shimada et al. 1990, 1992) using rats (Sprague-Dawley) and mice (CD-l and 

BDF1), different dosing regimes (single, double, or triple doses), and different target cells (bone 
PAHs 48 

2. HEALTH EFFECTS 

marrow and peripheral blood reticulocytes). A single oral gavage dose of 63 mg/kg benzo[a]pyrene 

significantly (p<0.0l) increased the yield of chromosomes with abnormal morphology in bone marrow 

cells collected from hybrid lC3F1 male mice (Adler and Ingwersen 1989). 

There is conflicting evidence that the genetic damage induced by benzo[a]pyrene is partially controlled 

by the expression of structural genes for benzo[a]pyrene-specific cytochromes P-450. In one study, 

two inbred strains of mice differing in AHH inducibility (AHH-inducible strain C57BL/6 and 

AHH-noninducible strain DBA/2) received two consecutive daily doses of either 10 or 100 mg/kg of 

the test material (Wielgosz et al. 1991). Animals were sacrificed 5 days postexposure, and bone 

marrow and spleen cells were examined for sister chromatid exchange and DNA adducts. Results 

showed a marked increase in sister chromatid exchange induction and the formation of DNA adducts 

in bone marrow and spleen cells recovered from the DBA/2 mice (AHH-noninducible) in both dose 

groups compared to the C57BL/6 (AHH-inducible) mice. However, no clear correlation between AHH 

inducibility and the positive clastogenic response induced by 150 mg/kg benzo[a]pyrene was found in 

adult male and female mice with genetically determined differences in AHH induction (Adler et al. 

1989). Similarly, the transplacental exposure of 11-day-old homozygous and hybrid embryos (dams 

received a single oral gavage dose of 150 mg/kg and embryos were sampled 15 hours after treatment) 

to benzo[a]pyrene showed that the clastogenic response was independent of genetic constitution. 

In contrast to the relatively uniform evidence that benzo[a]pyrene is a genotoxin in whole animals, the 

test material failed to induce unscheduled DNA synthesis (UDS) in the parenchymal liver cells of 

Brown Norway rats exposed by oral gavage to 12.5 mg/mL (Mullaart et al. 1989). There was, 

however, a clear increase in single-strand DNA breaks in cells from the two major centers of 

metabolism (the parenchymal liver and intestinal cells) of the treated animals that was not apparent in 

the nonparenchymal liver cells. 

Significant (p<0.05), but marginal, increases in the frequency of abnormal sperm were found in CD-l 

mice (8-12/group) exposed via oral gavage to benzo[a]pyrene doses ranging from 360 to 432 mg/kg 

(Salamone et al. 1988). The effect, however, was not clearly dose related, and the wide variation in 

the background frequency rendered the data inconclusive. Comparable doses produced no adverse 

effects in B6C3F1 mice. Similarly, the evaluation of pyrene (241-844 mg/kg) in this study yielded 

uniformly negative results. 
PAHs 49 

2. HEALTH EFFECTS 

Orally administered fluoranthene (400 and 750 mg/kg) did not increase the sister chromatid exchange 

frequency in mice (Palitti et al. 1986). Gene mutations were not produced in bacteria or yeast in a 

host-mediated assay in which anthracene, benzo[a]pyrene, chrysene, or fluoranthene were administered 

to mice by gavage; positive results were produced in bacteria in the same test system in which mice 

were exposed to benz[a]anthracene and injected intraperitoneally with the bacteria (Simmon et al. 

1979). Other genotoxicity studies are discussed in Section 2.4. 

2.2.2.8 Cancer 

No studies were located regarding cancer in humans following oral exposure to the 17 PAHs discussed 

in this profile. The animal studies discussed in this section are presented first by exposure duration 

(acute, intermediate, and chronic), and within each duration category the information on individual 

PAHs is discussed in alphabetical order. PAHs for which no information was available for specified 

exposure durations were omitted. 

Acute-Duration Exposure. Mice acutely administered 1.5 mg/day benz[a]anthracene by oral 

gavage two times over 3 days exhibited increased incidences of hepatomas and pulmonary adenomas 

(80% and 85%, respectively) as compared to control incidences (10% and 30% for hepatomas and 

pulmonary adenomas, respectively) after 568 days of observation (Klein] 1963). No malignant tumors 

were observed in this study. 

Mice fed benzo[a]pyrene in the diet at a concentration equivalent to 33.3 mg/kg/day exhibited 

forestomach neoplasms following 2 or more days of consumption. However, a lower concentration of 

benzo[a]pyrene (equivalent to 13.3 mg/kg/day) administered for up to 7 days did not produce 

forestomach tumors (Neal and Rigdon 1967) (see Table 2-2). Hamsters have also been observed to 

develop papillomas and carcinomas of the alimentary tract in response to gavage or dietary exposure to 

benzo[a]pyrene (Chu and Malmgren 1965). A 77% mammary tumor incidence was observed 

90 weeks after a single oral dose of 50 mg benzo[a]pyrene (100 mg/kg) was administered to rats, as 

compared to a 30% incidence in untreated animals (McCormick et al. 1981). 

A single dose of 0.05 mg/kg dibenz[a,h]anthracene in polyethylene glycol (PEG)-400 failed to induce 

tumors in male Swiss mice after 30 weeks. However, forestomach papillomas were found in 10% of 

mice administered a single dose of 0.05 mg/kg dibenz[a,h]anthracene followed by 30 weekly doses of 
PAHs 50 

2. HEALTH EFFECTS 

PEG alone, and in 21% of the mice when the dibenz[a,h]anthracene dose was followed by 30 weekly 

doses of PEG plus 3% croton oil (Berenblum and Haran 1955). Treatment with croton oil alone 

yielded a 14-16% tumor incidence. These results suggest that the carcinogenic activities of croton oil 

and dibenz[a,h]anthracene are additive in the mouse forestomach. 

Intermediate-Duration Exposure. One intermediate-duration study was located that evaluated 

the carcinogenic potential of acenaphthene. Male and female mice exposed to 0, 175, 350, or 

700 mg/kg/day acenaphthene by gavage for 13 weeks showed no evidence of tumorigenesis at 

necropsy (EPA 1989c). 

Similarly, only one intermediate-duration study was located that evaluated the carcinogenic potential of 

anthracene. Male and female mice exposed to 0, 250, 500, 1,000 mg/kg/day anthracene by gavage for 13 

weeks showed no evidence of tumorigenesis at necropsy (EPA 1989d). 

One intermediate-duration study was located that evaluated the carcinogenic potential of 

benz[a]anthracene. Mice that received intermittent gavage doses of 1.5 mg/kg/day benz[a]anthracene 

for 5 weeks (Klein 1963). Mice were sacrificed at a median age of 437 or 547 days. The treated 

mice killed at 437 days exhibited a 95% incidence of pulmonary adenomas at an average of 3 per lung 

and a 46% incidence of hepatomas, with an average of 2.1 per tumor-bearer. Forestomach papillomas 

were found in 5% of the mice. Control animals killed after 441 days exhibited a 10% incidence of 

pulmonary adenomas. Treated mice sacrificed after 547 days exhibited a 95% pulmonary adenoma 

incidence, as was observed in the group sacrificed earlier, but an increased hepatoma incidence of 

100%. Control animals sacrificed after 600 days had 30 and 10% incidences of pulmonary adenomas 

and hepatomas, respectively. This study was not adequately reported; it did not include complete 

histopathology, adequate treatment durations, large enough sample sizes, or statistical analysis. 

Although this study is inconclusive because of methodological limitations, it does provide some 

qualitative evidence for the potential carcinogenicity of benz[a]anthracene by the oral route. 

Intragastric doses of 67-100 mg/kg benzo[a]pyrene have been shown to elicit pulmonary adenomas 

and forestomach papillomas in mice (Sparnins et al. 1986; Wattenberg and Leong 1970). Intermittent 

gavage exposure of mice to 67-100 mg/kg benzo[a]pyrene resulted in increased forestomach (100%) 

and pulmonary tumor incidences relative to controls at 30 weeks of age (Sparnins et al. 1986; 
PAHs 51 

2. HEALTH EFFECTS 

Wattenberg and Leong 1970). The study by Wattenberg and Leong (1970) involved gavage 

administration of approximately 1.0 mg of benzo[a]pyrene once a week for 8 weeks. 

The incidence of forestomach tumors (papillomas and carcinomas) in mice was related to the duration 

of oral exposure to benzo[a]pyrene following intermediate-duration administration of dietary 

benzo[a]pyrene at various doses up to 250 ppm (33.3 mg/kg/day) for 30-197 days (Neal and Rigdon 

1967, see Table 2-2). The tumor incidence also increased with increasing dose. In the same study, 

mice fed 250 ppm (33 mg/kg/day) for periods of 1-7 days exhibited increased forestomach tumor 

incidences following 2 or more days of benzo[a]pyrene exposure (total dose of 2 mg), while mice fed 

10 ppm (13.3 mg/kg/day) for 110 days (total dose of 4.48 mg) did not develop tumors. The authors 

suggest that these findings provide evidence that there are no cumulative carcinogenic effects of 

benzo[a]pyrene or its metabolites in mice. These data suggest that differences in susceptibility may be 

strongly influenced by the age of the mice at the time that they were initially exposed. This study 

provides the best dose-response information available for the oral route of exposure despite the 

irregular protocol employed, although the relevance of forestomach tumors in rodents to human cancer 

is the subject of some controversy because humans lack a forestomach. 

An association between dietary benzo[a]pyrene and the development of leukemia and tumors of the 

forestomach and lung has been observed in mice. Tumor incidence was related to both dose and 

length of exposure (except in the case of leukemia). Mice administered dietary doses of up to 

1,000 ppm (up to 133 mg/kg/day) for intermediate lengths of time (23-238 days) exhibited an 

increased incidence of forestomach tumors (papillomas and carcinomas) (Rigdon and Neal 1966, 1969) 

(see Table 2-2). Mice administered 250 ppm (33.3 mg/kg/day) benzo[a]pyrene developed papillomas or 

carcinomas of the forestomach (64%) and all the mice in the 1,000-ppm (133 mg/kg/day) group 

exhibited forestomach tumors after 86 days of benzo[a]pyrene consumption. A similar relationship was 

observed for the incidence of lung tumors: mice fed 250 ppm (33.3 mg/kg/day) benzo[a]pyrene 

exhibited an increased lung adenoma incidence. The occurrence of leukemia was related to the ingestion 

of 250 ppm (33.3 mg/kg/day) benzo[a]pyrene; 37% of the treated mice developed leukemias (Rigdon 

and Neal 1969) (see Table 2-2). The lack of consistent protocol in these experiments and the short 

exposure duration and observation periods preclude the assessment of a dose-response relationship. 

Furthermore, because tumors were reported as combined papillomas and carcinomas, no distinction 

between these benign and malignant tumors can be made. 
PAHs 52 

2. HEALTH EFFECTS 

Mammary tumors have also been observed following intermediate-duration exposure to benzo[a]pyrene 

in rats. Eight weekly oral doses of 6.25 mg benzo[a]pyrene (12.5 mg/kg) administered to rats resulted in 

a 67% increase in the incidence of mammary tumors in female rats after 90 weeks of observation 

(McCormick et al. 1981). A 30% incidence of these tumors was observed in the control animals. 

Two intermediate-duration studies investigated the carcinogenicity of dibenz[a,h]anthracene in animals 

following oral exposure. Mammary carcinomas were observed in 5% of the female BALB/c mice 

dosed with 0.5% dibenz[a,h]anthracene after 15 weeks of dosing; however, no control group was 

included (Biancifiori and Caschera 1962). In the other study, male and female rats were administered 

an emulsion of aqueous olive oil and dibenz[a,h]anthracene in place of their drinking water for up to 

200 days (Snell and Stewart 1963). Pulmonary adenomatosis, alveologenic carcinoma, mammary 

carcinoma, and hemangioendotheliomas were observed in the treated rats. These tumors were not 

observed in the control animals. However, extensive dehydration and emaciation occurred because the 

animals did not tolerate the vehicle well, which lead to early death and the need to periodically 

remove the animals from the treatment vehicle. Neither of these studies was adequately reported: 

they did not perform appropriate histopathologic evaluations, treatment or study durations were 

inadequate, and the sample size was inadequate. Despite these methodological limitations, these 

studies do provide some evidence of dibenz[a,h]anthracene’s carcinogenicity by the oral route. 

One intermediate-duration study was located that evaluated the carcinogenic potential of fluoranthene. 

Male and female mice exposed to 0, 125, 250, or 500 mg/kg/day fluoranthene by gavage for 13 weeks 

showed no evidence of tumorigenesis at necropsy (EPA 1988e). 

Similarly, only one intermediate-duration study was located that evaluated the carcinogenic potential of 

fluorene. Male and female mice exposed to 0, 125, 250, or 500 mg/kg/day fluorene by gavage for 

13 weeks showed no evidence of tumorigenesis at necropsy (EPA 1989e). 

Chronic-Duration Exposure. Benzo[a]pyrene was administered in the diet of 32 Sprague-Dawley 

rats/sex/group either every 9th day or 5 times/week at a dose of 0.15 mg/kg until the animals were 

either moribund or dead (Brune et al. 1981). An untreated control group consisted of 32 animals/sex. 

There was no treatment-related effect on survival and no treatment-related increase in tumors at any 

one site. However, a statistically significant increase in the proportion of animals with tumors of the 

forestomach, esophagus, and larynx combined was noted among animals receiving treatment 
 

 

 

PAHs 53 

2. HEALTH EFFECTS 

5 times/week (the combined incidence of animals with these tumors was 3/64, 10/64, and 3/64 in the 

controls, the group fed benzo[a]pyrene 5 times/week, and the group fed benzo[a]pyrene every 9th day). 

In the same study, groups of 32 Sprague-Dawley rats/sex were administered 0.15 mg/kg 

benzo[a]pyrene by gavage in a 1.5% caffeine solution either every 9th day (Group 3), every 3rd day 

(Group 2), or 5 times/week (Group 1) until the animals were moribund or dead, resulting in average 

annual doses of 6, 18, or 39 mg/kg, respectively. Survival was adversely affected only in Group 3 

(mean survival.time = 87 weeks versus 102 weeks in the controls). Treatment with benzo[a]pyrene 

significantly increased the proportion of animals with tumors of the forestomach, esophagus, and 

larynx (the combined tumor incidence was 3/64, 6/64, 13/64, 2664, and 14/64 for the untreated 

controls, the gavage controls, and Groups 3, 2, and 1, respectively). 

Summary. These results indicate that benz[a]anthracene, benzo[a]pyrene, dibenz[a,h]anthracene, 

and possibly other PAHs are carcinogenic to rodents following oral exposure at high doses. 

All reliable CELs in mice for acute- and intermediate-duration exposure are recorded in Table 2-2 and 

plotted in Figure 2-2. 

2.2.3 Dermal Exposure 

2.2.3.1 Death 

No studies were located regarding death in humans or animals after dermal exposure to the 17 PAHs 

discussed in this profile. 

2.2.3.2 Systemic Effects 

No studies were located regarding respiratory, cardiovascular, gastrointestinal, hematological, 

musculoskeletal, hepatic, renal, or ocular effects in humans or animals after dermal exposure to any of 

the 17 PAHs discussed in this profile. Other systemic effects observed after dermal exposure are 

discussed below. 

The highest NOAEL values and all LOAEL values from each reliable study for each species and 

duration category are recorded in Table 2-3. 





PAHs 59 

2. HEALTH EFFECTS 

Dermal Effects. Mixtures of carcinogenic PAHs cause skin disorders in human and animals; 

however, specific effects in humans of individual PAHs, except for benzo[a]pyrene, have not been 

reported. Mixtures of PAHs are also used to treat some skin disorders in humans. From these patients 

comes much of the data describing dermal effects of PAH exposure. 

Regressive verrucae (i.e., warts) was reported following up to 120 dermal applications of 1% 

benzo[a]pyrene in benzene to human skin over 4 months (Cottini and Mazzone 1939). Although 

reversible and apparently benign, the changes were thought to represent neoplastic proliferation. 

Adverse dermal effects have been noted in humans following intermediate-duration dermal exposure to 

benzo[a]pyrene in patients with the preexisting dermal conditions of pemphigus vulgaris (acute or 

chronic disease characterized by occurrence of successive crops of blisters) and xeroderma 

pigmentosum (a rare disease of the skin marked by disseminated pigment discolorations, ulcers, and 

cutaneous and muscular atrophy) (Cottini and Mazzone 1939). A 1% benzo[a]pyrene solution 

topically applied to patients with pemphigus resulted in local bullous eruptions characteristic of the 

disease. Patients with xeroderma pigmentosum exposed to 1% benzo[a]pyrene slightly longer than the 

pemphigus patients exhibited only pigmentary and slight verrucous effects. Similarly treated patients 

with preexisting active skin lesions due to squamous cell cancer showed a general improvement and/or 

retardation of the lesion. The severity of abnormal skin lesions appeared to be related to age; those in 

the lowest age range exhibited fewer and less-severe effects than those in the mid-range groups. No 

such age relationship of effects involving those patients with normal or preexisting skin lesions was 

noted. 

Adverse dermal effects have also been ohserved in animals following both acute- and 

intermediate duration dermal exposure to various PAHs. For example, acute topical application of 

benzo[a]pyrene, benz[a]anthracene, or dibenz[a,h]anthracene applied to the shaved backs of Swiss mice 

were all reported to suppress sebaceous glands (Bock and Mund 1958). However, controls were not 

employed; therefore, it is not possible to determine if the effects seen were due to the solvent and/or the 

application procedures. 

Benzo[a]pyrene was applied once weekly to the skin of female ICR/Harlan mice (43-50/group) at 

doses of 16, 32, or 64 μg per application for 29 weeks (Albert et al. 1991b). Cell cycle kinetics and 
PAHs 60 

2. HEALTH EFFECTS 

morphometrics were evaluated. Evidence of epidermal cytotoxicity and death followed by 

regeneration was seen in animals administered 64 μg benzo[a]pyrene beginning the first weeks of 

exposure and later in the lower dose groups. This evidence included dose-related epidermal thickening 

and vertical nuclei stacking, increased mitotic labeling (2-4-fold with increasing dose), increased 

incidence of pyknotic and dark cells, and a pronounced inflammatory response in the dermis. The 

increase in cell proliferation was accompanied by only a minor increase in the size of the epidermal 

cell population, indicating that the proliferation was a regenerative response. 

An acute (96-hour) dermal application of anthracene to the backs of hairless mice followed by 

ultraviolet radiation exposure for 40 minutes resulted in enhanced dermal inflammation compared to 

mice exposed exclusively to ultraviolet radiation. However, this effect was reversed within 48 hours 

(Forbes et al. 1976). Anthracene thus potentiates the skin damage elicited by sunlight exposure and 

may be considered a photosensitizer in hairless mice. 

In animals, dermal application of 1% benzo[a]pyrene to the skin of hairless mice resulted in epidermal 

cell growth alterations (Elgjo 1968). Increases were observed in mitotic rates, mitotic counts, and 

mitotic duration and the author suggested that these were indicative of a regenerative reaction. 

However, concurrent controls were not utilized. The authors concluded that the alterations in the 

kinetics of epidermal cell growth produced by benzo[a]pyrene were more sustained than after 

application of croton oil. The study is limited for drawing conclusions concerning the dermal toxicity 

of benzo[a]pyrene because experimental data were compared with historical controls only, no acetone 

control was evaluated, and the statistical significance of the increased values was not determined. 

2.2.3.3 Immunological and Lymphoreticular Effects 

No studies were located regarding immunological effects in humans following dermal exposure to the 

17 PAHs discussed in this profile. 

Benzo[a]pyrene can elicit an immune response when applied dermally to the skin of animals. In mice, 

acute application of 120 μg benzo[a]pyrene elicited an allergic contact hypersensitivity in C3H mice 

that was antigen specific (Klemme et al. 1987). Slight contact hypersensitivity was also observed in 

guinea pigs following two dermal applications of 0.001% benzo[a]pyrene given over a period of 

2-3 weeks. This response was more severe at a dose of 1.0% benzo[a]pyrene (Old et al. 1963). 
PAHs 61 

2. HEALTH EFFECTS 

In addition to eliciting a contact hypersensitivity response, benzo[a]pyrene has been shown to suppress 

this response to other sensitizers. The effects of dermally applied benzo[a]pyrene (alone or following 

dermal pretreatment with the prostaglandin synthetase inhibitor, indomethacin) on contact 

hypersensitivity (cell-mediated immunity), and production of antibodies to dinitrophenol (DNP) 

(humoral immunity) were studied in male BALB/c mice treated for 6 weeks to 6 months (Andrews et 

al. 1991a). A group of mice treated with acetone served as controls. Benzo[a]pyrene alone caused a 

significant reduction (p<0.01) in the contact hypersensitivity response to dinitrofluorobenzene (DNFB) 

as measured by increases in ear thickness when compared to the vehicle controls. However, 

indomethacin pretreatment prevented the benzo[a]pyrene-induced contact hypersensitivity response. 

Benzo[a]pyrene also reduced antibody titres to DNP in treated mice. This suppressive effect on 

humoral immune function was not restored by pretreatment with indomethacin. These findings led the 

authors to conclude that the mechanism of benzo[a]pyrene-induced suppression of cell-mediated 

immunity involved prostaglandins, whereas benzo[a]pyrene-induced suppression of humoral immunity 

operated via a mechanism independent of prostaglandins. In a subsequent experiment, the effects of 

dermally applied benzo[a]pyrene (alone or following subcutaneous implantation of the prostaglandin 

synthetase inhibitor, indomethacin) on Langerhans cells and on skin prostaglandin (PGE2) levels were 

studied in male BALB/c mice treated for 3 weeks (Andrews et al. 1991b). Langerhans cells are 

antigen-presenting cells involved in cell-mediated immunity in skin. A group of mice treated with 

indomethacin served as controls. Benzo[a]pyrene alone caused a significant increase in the number of 

skin Langerhans cells, but reduced the percentage of Langerhans cells with dendritic morphology. 

Skin PGE2 levels were also significantly increased by benzo[a]pyrene. Indomethacin attenuated the 

increase in Langerhans cell number and the changes in their morphology, and increased PGE2 levels, 

such that all of these parameters were similar to those measured in the control animals. Based on 

these results, the authors suggested that benzo[a]pyrene induces increases in skin PGE2 that in turn 

alter Langerhans cell number and morphology such that the cell-mediated immune response to skin 

antigens is suppressed. 

An earlier study also demonstrated that benzo[a]pyrene affects epidermal Langerhans cells and dermal 

immunological responses. Female BALB/c mice were administered dermal applications on the dorsal 

skin of 0.5% benzo[a]pyrene in acetone twice weekly for up to 6 months (Ruby et al. 1989). Animals 

treated with acetone served as controls. The density, area, perimeter, and morphology of epidermal 

Langerhans cells were evaluated, along with the contact hypersensitivity response to DNFB. 

Benzo[a]pyrene treatment caused an increase in the number of epidermal Langerhans cells (as 
 

 

 

 

PAHs 62 

2. HEALTH EFFECTS 

determined by Ia antigens and -glucuronidase) from week 2 to week 5 of treatment and after weeks 

10 and 18. The area and perimeter of these cells were unaffected by benzo[a]pyrene treatment, but the 

morphology was altered in that the dendrites appeared shortened. The contact hypersensitivity 

response to DNFB was significantly reduced in the benzo[a]pyrene-treated mice from 4 to 24 weeks of 

treatment. The authors propose that benzo[a]pyrene alters Langerhans cell number and morphology 

such that the cell-mediated immune response to skin antigens is suppressed. Skin tumors appeared in 

20% of the benzo[a]pyrene-treated mice after 18 weeks of treatment, and 35% of the mice had 

l-3 tumors after 24 weeks of treatment. The tumors were squamous papillomas (58%) and squamous 

cell carcinomas (42%). The changes in Langerhans cell number, distribution, and morphology 

coincided with the onset of tumors and other nonneoplastic skin lesions that were observed (epidermal 

hyperplasia and cellular atypia). 

All reliable LOAELs from each reliable study for immunological effects for each species and duration 

category are recorded in Table 2-3. 

No studies were located regarding the following health effects in humans or animals following dermal 

exposure to the 17 PAHs discussed in this profile: 

2.2.3.4 Neurological Effects 

2.2.3.5 Reproductive Effects 

2.2.3.6 Developmental Effects 

2.2.3.7 Genotoxic Effects 

No studies were located regarding genotoxic effects in humans following dermal exposure to the 

17 PAHs discussed in this profile. A single topical application of benzo[a]pyrene (0.5-500 μg/mouse) 

or chrysene (50-1,000 μg/mouse) to groups of HRA/Skh hairless mice (four mice/dose/group) resulted 

in significantly increased frequencies of micronucleated keratinocytes (He and Baker 1991). In the 

same study, micronuclei were not induced in the mouse skin cells following application of 

2.5-2,500 μg/pyrene per mouse. Male SENCAR mice receiving two topical applications of 20 μg 

benzo[a]pyrene at 72-hour intervals exhibited increased DNA adduct formation in both epidermal and 
PAHs 63 

2. HEALTH EFFECTS 

lung tissue (Mukhtar et al. 1986). Following a single topical application, 100 μg benzo[b]fluoranthene, 

benzo[j]fluoranthene, benzo[k]fluoranthene, and indeno[ 1,2,3-c,d]pyrene were reported to bind to DNA 

in CD-l mouse skin (Weyand et al. 1987). The relative extent of binding was benzo[b]fluoranthene > 

benzo[j]fluoranthene > benzo[k]fluoranthene > indeno[ 1,2,3-c,d]pyrene. Covalent binding of chemicals 

to DNA can result in strand breaks and DNA damage, ultimately leading to mutations. 

Benzo[a]pyrene (62.5 or 500 μg) was applied once to the shaved backs of male C57BL/6 mice 

(Bjelogrlic et al. 1994). Mice were killed at different time intervals after the treatment. DNA was 

isolated from the skin, purified, and analyzed for benzo[a]pyrene-7,8-diol-9,10-epoxide-DNA adducts. 

Skin was also evaluated for monoclonal antibody binding to mouse p53 protein, which has been shown 

to increase in response to DNA damage. Alterations in p53 are the most frequently observed 

mutations in human cancer. Benzo[a]pyrene-7,8-diol-9,10-epoxide-DNA adducts reached their 

maximum concentration 24 hours after the treatment, and decreased sharply within 1 week, regardless 

of the dose. An increase in p53 protein was seen only after treatment with 500 μg benzo[a]pyrene. 

Benzo[j]fluoranthene, benzo[j]fluoranthene-4,5-diol, and benzo[j]fluoranthene-9,10-diol were applied to 

the shaved backs of CD-l mice and the DNA adducts were isolated and separated using 

multidimensional thin-layer chromatography (TLC) and reverse-phase high performance liquid 

chromatography (HPLC) (Weyand et al. 1993a). The highest level of adducts was observed with 

benzohlfluoranthene-4,5-diol, which resulted in the formation of 383 mol of DNA adducts/mg DNA. 

This level of DNA modification was more than 2 orders of magnitude greater than that observed with 

benzo[j]fluoranthene. In contrast, the major DNA adducts detected with benzo[j]fluoranthene-9,10-diol 

had chromatographic properties distinctly different than the adducts formed from either 

benzo[j]fluoranthene or B[j]F-4,5-diol. The adducts of the diols corresponded to DNA adducts 

produced in vitro from the respective diolepoxides. In a companion study, benzo[b]fluoranthene, 

benzo[b]fluoranthene-9,10 diol, 6-hydroxy-benzo[b]fluoranthene-9,10-diol, or 5-hydroxybenzo[ 

b]fluoranthene-9,10-diol were applied to the shaved backs of CD-1 mice and the DNA adducts 

were isolated and separated using multidimensional TLC and reverse-phase HPLC (Weyand et al. 

1993b). Benzo[b]fluoranthene formed one major adduct and 4 minor adducts. The DNA adducts 

formed from 5-hydroxybenzo[b]fluoranthene-9, l0 diol had identical retention to the major and one of 

the minor adducts of benzo[b]fluoranthene. These two adducts accounted for 58% of the modified 

nucleotides produced by benzo[b]fluoranthene application to mouse skin. 
 

PAHs 64 

2. HEALTH EFFECTS 

The DNA binding of manufactured gas plant residue (MGP) components in male B6C3Fl mice was 

investigated following topical administration (Weyand and Wu 1994). For topical exposure, male 

mice were treated with 10 mg MGP residue in 200 μL acetone, and sacrificed 24 hours later. Two 

mixtures of MGP residue were used: Mix of 3 combining equal amounts of samples from three 

different MGP plant sites, and Mix of 7 combining equal amounts of samples from seven different 

MGP plant sites, including those used in the Mix of 3. The mixtures contained pyrene, 

benz[a]anthracene, chrysene, benzo[b]fluorene, benzo[k]fluorene, benzo[a]pyrene, 

indeno[l,2,3-cd]pyrene, dibenz[a,h]anthracene, benzo[g,h,i]perylene. Data were presented in terms of 

pyrene. Animals were sacrificed 24 hours after treatment, and skin and lung were excised and DNA 

isolated. Chemical-DNA adduct formation was evaluated. Topical application MGP residue in 

acetone resulted in similar levels of DNA adduct in the skin for both the Mix of 3 and the Mix of 7. 

The total level of adducts detected in the lung after topical administration was identical to the response 

after dietary exposure, i.e., the Mix of 3 (1.4 mg/kg/day pyrene) produced adduct levels that were two 

times greater than the levels induced by the Mix of 7 (1.2 mg/kg/day pyrene). Other genotoxicity 

studies are discussed in Section 2.4. 

2.2.3.8 Cancer 

No studies were located that gave evidence of a direct association between human dermal exposure to 

individual PAHs and cancer induction. However, reports of skin tumors among individuals exposed to 

mixtures containing PAHs lend some qualitative support to their potential for carcinogenicity in 

humans. The earliest of these is the report by Pott (1775) of scrotal cancer among chimney sweeps. 

More recently, skin cancer among those dermally exposed to shale oils has been reported (Purde and 

Etlin 1980). However, these reports provide only qualitative suggestions pertaining to the human 

carcinogenic potential of all of the 17 PAHs discussed in this profile, or at least the compounds found 

in chimneys and shale oils, such as benzo[a]pyrene, chrysene, dibenz[a,h]anthracene, 

benz[a]anthracene, and benzo[b]fluoranthene. Limitations in these reports include no quantification of 

exposure to individual PAHs and concurrent exposure to other putative carcinogens in the mixtures. 

It has been suggested that an increase in the number of skin melanocytes correlates with the sebaceous 

gland suppression index, and that the short-term melanocyte-activation test is useful for the detection 

of skin carcinogens and promoters. Some chemical carcinogens have been shown to induce 

melanogenesis in melanoblasts in the skin. Anthracene, benzo[a]pyrene, chrysene, fluoranthene, 
PAHs 65 

2. HEALTH EFFECTS 

fluorene, and pyrene were examined for their ability to induce melanocyte activation by topical 

application to the backs of mice for 1 or 2 consecutive days. Benzo[a]pyrene, an animal skin 

carcinogen, was a potent melanocyte inducer at doses of 20-100 μg/mouse (0.005-0.025 mg/cm2

) as 

demonstrated by an increase of up to 19 times over controls in the number of dopa-positive cells, 

whereas no effects were seen at 4 μg/mouse (0.001 mg/cm2

). Chrysene, a weak skin carcinogen (in 

animals), increased the number of dopa-positive cells to four times that of controls following an 

application of 50 μg/mouse (0.0125 mg/cm2

), while larger doses did not cause further increases in the 

numbers of these cells. Other PAHs such as anthracene, fluoranthene, fluorene, and pyrene (PAHs 

that are considered to be noncarcinogenic) produced no increases in the number of active melanocytes 

when applied at a dose of 200 μg/mouse (0.05 mg/cm2) (Iwata et al. 1981). 

Complete Carcinogenesis Studies. Studies in laboratory animals have demonstrated the ability of 

benz[a]anthracene, benzo[b]fluoranthene, benzo[j]fluoranthene, benzo[a]pyrene, chrysene, 

dibenz[a,h]anthracene, and indeno[l,2,3-c,d]pyrene to induce skin tumors (i.e., they are complete 

carcinogens) following intermediate dermal exposure. Anthracene, fluoranthene, fluorene, 

phenanthrene, and pyrene do not act as complete carcinogens. The data supporting these conclusions 

are discussed below by chemical. Only those studies considered adequate and reliable with respect to 

study design and adequacy of reporting are presented in Table 2-3. 

Anthracene. Skin painting experiments were conducted on groups of 20 male C3H/HeJ mice 

(Warshawsky et al. 1993). Anthracene dissolved in toluene was applied to shaved skin twice weekly 

for six months at a dose of 0.05 mg. Tumor incidence was determined at the end of the study. For 

anthracene, administration alone produced tumors in 0 of 14 animals. With coadministration of 

0.05 mg benzo[a]pyrene, 1 of 13 (8%) had a papilloma, with a mean latency period of 85 weeks. 

Anthracene was negative as a complete carcinogen following chronic dermal exposure (Habs et al. 

1980). Swiss mice receiving 10% anthracene in acetone topically applied to their backs three times a 

week throughout their lifetime did not develop any skin tumors after 20 months (Wynder and 

Hoffmann 1959a). 

Benz[a]anthracene. Benz[a]anthracene has been shown to cause skin tumors in mice following 

intermediate-duration dermal application. Graded concentrations of benz[a]anthracene in toluene or 

n-dodecane applied to the backs of mice for 50 weeks resulted in dose-related increases in tumor 

incidence (Bingham and Falk 1969). This response was enhanced when n-dodecane was the solvent 
PAHs 66 

2. HEALTH EFFECTS 

compared with toluene. Malignant tumors were observed at dose levels of about 0.02% 

benz[a]anthracene (0.15 mg/kg/day) and above when toluene was the solvent; however, when 

n-dodecane was the solvent, tumors were observed.at much lower concentrations of 0.0002% 

benz[a]anthracene (0.0015 mg/kg/day). The dose-response relationship reported in this study is 

extremely shallow (sublinear) over two orders of magnitude. A possible explanation for this is that the 

tumorigenic potency of certain PAHs is tempered by their cytotoxicity. Thus, the cytotoxic effects to 

epithelial cells may self-limit their potency as tumorigens. No solvent controls were included for 

comparison (Bingham and Falk 1969). 

Intermediate-duration topical application of benz[a]anthracene to the backs of mice for 30 weeks 

resulted in a slightly elevated (2.6%) (but not statistically significant) skin tumor incidence. No 

definitive conclusions can be drawn from this study since only one dose was employed and no 

statistical analysis was performed. 

Benzo[b]fluoranthene. A dose-response relationship for the dermal carcinogenicity of 

benzo[b]fluoranthene has been demonstrated over a one order-of-magnitude dose range in Swiss mice 

receiving (0.0l-0.5%) benzo[b]fluoranthene throughout their lifetime. Survival was also dose related. 

Although this study was designed as a long-term (chronic) bioassay, malignant tumors (90% 

carcinomas) appeared as early as 4 months in the high-dose group. Papillomas and carcinomas (65% 

and 85%, respectively) also appeared after 5 months in the mid-dose group. As a result, this study 

provides evidence that benzo[b]fluoranthene is carcinogenic following intermediate-duration exposure. 

The lowest dose at which benzo[b]fluoranthene elicited malignant tumors was 0.l%, which is 

approximately equal to a dose of 2.9 mg/kg received three times weekly, or an average daily dose of 

1.2 mg/kg (Wynder and Hoffmann 1959b, see Table 2-3). In another chronic dermal study, 

benzo[b]fluoranthene produced a significant carcinogenic response of approximately one-third the 

potency of benzo[a]pyrene. The lowest dose at which tumors appeared was 3.4 μg 

benzo[b]fluoranthene; however, no distinction was made between papillomas and carcinomas (Habs et 

al. 1980). 

Benzo[j]fluoranthene. Benzofilfluoranthene (0.1% or 0.5%) applied to the skin of female Swiss mice 

thrice weekly for life induced skin papillomas in 70% and 95% of the animals, respectively, and skin 

carcinomas in 105% and 95% of the animals, respectively, after 9 months of treatment (Wynder and 

Hoffmann 1959b). 
PAHs 67 

2. HEALTH EFFECTS 

No statistically significant increase in the incidence of skin tumors was noted in female NMRI mice 

dermally administered 3.4, 5.6, or 9.2 μg benzo[j]fluoranthene in acetone twice weekly for life (Habs 

et al. 1980). A total of 4 benzo[j]fluoranthene-treated mice developed skin tumors (application site 

sarcoma, papillomas, and a carcinoma). 

Benzo[k]fluoranthene. Chronic dermal application of benzo[k]fluoranthene to Swiss mice resulted in 

no tumors, but skin papillomas were observed in 10% of animals when the concentration of 

benzo[k]fluoranthene was increased. Statistical analyses were not performed (Wynder and Hoffmann 

1959b). In another study, no significant increase in tumor incidence was observed in NMRI mice 

painted with up to 9.2 μg of benzo[k]fluoranthene twice a week for a lifetime; no effect on mortality 

was noted (Habs et al. 1980). 

Benzo[a]pyrene. Benzo[a]pyrene is a potent experimental skin carcinogen, and it is often used as a 

positive control in bioassays of other agents. Mixtures of PAHs that include benzo[a]pyrene such as 

coal tar were shown to be dermal carcinogens in animals as early as 1918 (Yamagiwa and Ichikawa 

1918). In its role as a positive control, benzo[a]pyrene is usually administered at a single dose level, 

and thus quantitative evaluation of dose-response relationships is not possible. 

Intermediate (19-20 weeks) topical application of a benzo[a]pyrene solution to the backs of mice 

resulted in a dose-related development of skin papillomas and squamous cell carcinomas (Cavalieri et 

al. 1988b, see Table 2-3; Shubik and Porta 1957). Benzo[a]pyrene was applied once weekly to the 

skin of female ICR/Harlan mice (43-50/group) at doses of 16, 32, or 64 μg for 29 weeks (Albert et al. 

1991ba). Cell kinetics, morphometrics, and tumor formation were evaluated. Skin tumors were first 

apparent 12-14 weeks after the start of exposure in the 32- and 64-μg groups and after 18 weeks in 

the 16-μg group. The overwhelming majority of these tumors were benign. The average time of 

progression from benign papillomas to malignant carcinomas was 8.1 ± 4.5 weeks. Because there was 

good correspondence between the dose-response patterns for epidermal damage and the occurrence of 

skin tumors, and because tumors that initially appear as benign can be the result of tumor promoting 

agents that increase cell proliferation rates, the authors proposed that the tumors seen after 

benzo[a]pyrene treatment were the result of promotion related to benzo[a]pyrene-induced tissue 

damage. However, benign tumors can be formed as a result of genetic damage as well. Because 

benzo[a]pyrene causes genetic damage in addition to increased rates of cell proliferation, it is likely 

that genetic damage also played a role. Following a similar protocol in mice (once weekly dermal 
PAHs 68 

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applications of benzo[a]pyrene in acetone at doses of 16, 32, and 64 μg/mouse for 34 weeks), tumor 

development was reported to be best described by a dose-squared response. Quantitative data for 

tumor incidence were not presented (Albert et al. 1991b). 

Carcinogenicity experiments on mouse skin were conducted with groups of 20 female CD-l mice. 

Benzo[a]pyrene (0.05 mg) dissolved in 50 mL toluene was applied to shaved skin of mice twice 

weekly for 6 months (Warshawsky et al. 1993). Tumor incidence was determined at the end of the 

study. Benzo[a]pyrene produced no tumors. 

In mice, the tumorigenic dose of benzo[a]pyrene is influenced by the solvent used for delivery. 

Graded concentrations of benzo[a]pyrene dissolved in decalin or a solution of n-dodecane and decalin 

were topically administered to the backs of mice for 50 weeks (Bingham and Falk 1969). Use of the 

n-dodecane and decalin solvent mixture significantly enhanced the potency of benzo[a]pyrene at lower 

doses in comparison with decalin alone. Malignant tumors appeared in 21% of the animals at 

0.00002% (0.0054 mg/kg/day) benzo[a]pyrene in dodecane and decalin solvent. In contrast, a 42% 

skin tumor incidence was not observed until 0.02% (4.8 mg/kg/day) benzo[a]pyrene in decalin alone 

was applied. The method of application was not specified, sample sizes were small and no decalin 

solvent controls were included; however, decalin is not considered to be carcinogenic. In this same 

study, intermediate (50 weeks) dermal application of benzo[a]pyrene dissolved in the co-carcinogens 

1-dodecanol or 1-phenyldodecane produced skin tumors in animals exposed to 0.05% benzo[a]pyrene 

in either solvent. The tumor incidence varied depending on the solvent concentration; however, the 

latency period was reduced only when 1-dodecanol was the solvent (Bingham and Falk 1969). 

Mice receiving 0.00l-0.01% of benzo[a]pyrene dermally applied to their backs throughout their 

lifetimes exhibited a dose-response relationship for skin tumors (Wynder and Hoffmann 1959a). A 

dose of 0.001% benzo[a]pyrene produced skin carcinoma and papilloma incidences of 4% and 45%, 

respectively. In another study conducted by Wynder and Hoffmann (1959b), higher concentrations of 

benzo[a]pyrene produced an 85% incidence of combined papillomas and carcinomas. These studies 

had a number of weaknesses, including no statistical treatment and no solvent control group. Dose 

quantification is difficult because of the method of application (Wynder and Hoffmann 1959a, 1959b). 

NMRI mice topically administered 2 μg benzo[a]pyrene throughout their lifetime also developed skin 

papillomas and carcinomas (45%) (Habs et al. 1984, see Table 2-3). CH3 mice administered a higher 

dose of 12.5 μg benzo[a]pyrene for 99 weeks exhibited malignant skin tumors (94%) (Warshawsky 
PAHs 69 

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and Barkley 1987, see Table 2-3). Increasing malignant carcinoma incidences in these dermal 

application studies can be correlated to increasing benzo[a]pyrene concentrations. 

In mice, the tumorigenic dose of benzo[a]pyrene is dependent on the strain. For example, Habs et al. 

(1980) tested 1.7-4.6 μg benzo[a]pyrene (0.016-0.04 mg/kg/day) in order to determine its 

dose-response relationship as a carcinogen when topically applied to the backs of NMRI mice 

throughout their lifetimes. A clear-cut dose-response relationship was seen for benzo[a]pyrene and the 

induction of tumors. The lowest dose at which skin tumors appeared was 1.7 μg (0.016 mg/kg/day). 

This strain of NMRI mice also has a high (70%) background incidence rate of systemic tumors, so an 

evaluation of the effects of benzo[a]pyrene on any organ other than the site of administration was not 

possible: 

Chrysene. Skin painting experiments with intermediate (6 months) dermal exposure with chrysene 

were conducted on groups of 20 male C3H/HeJ mice (Warshawsky et al. 1993). Chrysene (0.05 mg) 

dissolved in toluene was applied to shaved skin twice weekly for 6 months. Tumor incidence was 

determined at the end of the study. For chrysene, administration alone produced papillomas in 1 of 

15 animals (7%), with a mean latency period of 81 weeks. With coadministration of 0.0005 mg 

benzo[a]pyrene, 3 of 13 (23%) had tumors (papillomas and malignant), with a mean latency period of 

70 weeks. 

Chrysene has elicited skin tumors in mice following chronic (68-82 weeks) dermal exposure. Topical 

application of a chrysene solution in n-dodecane/decalin to the skin of mice produced a significant 

increase in the carcinogenic potency of chrysene compared with the use of decalin alone; 26% and 

63% of mice exhibited papillomas and carcinomas, respectively, at 49 weeks (Horton and Christian 

1974). Because only one dose level was employed, no dose-response relationship can be inferred, and 

no solvent control was included. However, in other experiments decalin and n-dodecane have been 

shown to be noncarcinogenic in mice (Bingham and Falk 1969). An average dose of 1.2 mg/kg/day is 

the lowest dose of chrysene that has been found to elicit malignant tumors in laboratory animals. 

In another chronic study, a higher concentration of chrysene applied dermally to the backs of Swiss 

mice for a lifetime also resulted in increased papilloma and carcinoma incidences (48% and 42%, 

respectively) compared to controls (Wynder and Hoffmann 1959a). Since only one dose was 

employed, no dose-response information can be inferred from this study. 
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Dibenz[a,h]anthracene: Dibenz[a,h]anthracene has also demonstrated a dose-response relationship for 

skin tumors over a two-orders-of-magnitude dose range following chronic exposure. Swiss mice 

received concentrations of 0.00l-0.1% dibenz[a,h]anthracene applied to their backs throughout their 

lifetimes and exhibited dose-related papilloma and carcinoma incidences at the site of application at 

the two lowest doses. A decreased tumor rate at the highest dose tested probably reflects 

dibenz[a,h]anthracene’s toxicity and the resulting decreased survival observed. The lowest 

concentration at which dibenz[a,h]anthracene elicited tumors was 0.001% (40% incidence of 

papillomas and 40% incidence of carcinomas), which is approximately equal to a dose of 0.029 mg/kg 

(0.012 mg/kg/day) (Wynder and Hoffmann 1959a). In another chronic derrnal study of 

dibenz[a,h]anthracene, a dose-related increase in skin carcinoma formation was observed, as well as 

decreased survival time and tumor latency period (Van Duuren et al. 1967). 

Groups of 50 female NMRI mice received dermal applications of dibenz[a,h]anthracene in acetone 

(total doses = 0, 136, 448, or 1,358 nmol) three times a week for a total of 112 weeks (Platt et al. 

1990). Papillomas were observed in 6%, 8%, and 32% of the treated animals, respectively. 

Fluoranthene. Skin painting experiments with intermediate-duration (6 months) dermal exposure were 

conducted on groups of 20 male C3H/HeJ mice (Warshawsky et al. 1993). Fluoranthene was 

dissolved in toluene and applied to shaved skin twice weekly for 6 months. Tumor incidence was 

determined at the end of the study. For fluoranthene, administration alone produced tumors in 0 of 

15 animals. With coadministration of 0.0005 mg benzo[a]pyrene, 1 of 12 (8%) had papillomas, with a 

mean latency period of 95 weeks. 

Chronic dermal application of up to 1% fluoranthene to the backs of mice did not induce skin tumors 

following a lifetime of application (Hoffmann et al. 1972; Horton and Christian 1974; Wynder and 

Hoffmann 1959a). 

Fluorene. Fluorene has been reported to be negative as a complete carcinogen (dose not specified) 

(Kennaway 1924). This information was obtained from an old, secondary source and therefore, its 

reliability is not known. 

Indeno[l,2,3-c,d]pyrene. Indeno[ 1,2,3-c,d]pyrene was applied to the skin of female Swiss mice three 

times weekly for 12 months in concentrations of 0.5% (500 μg/application), 0.l%, 0.05%, and 0.01% 
PAHs 71 

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(20 mice per group) using acetone as the solvent. A tumor dose-response with 7 papilloma-bearing 

mice and 5 carcinoma-bearing mice for 0.5%, 6 papilloma-bearing and 3 carcinoma-bearing mice for 

0.l%, and no skin tumors for 0.05% and 0.01% solutions was observed (Hoffmann and Wynder 1966). 

Chronic dermal application of indeno[ 1,2,3-c,d]pyrene in dioxane to mice did not produce an increased 

incidence of skin tumors. Similarly, chronic topical application of up to 9.2 μg of 

indeno[l,2,3-c,d]pyrene in acetone to the backs of mice for a lifetime resulted in no tumor induction 

(Habs et al. 1980). 

Phenanthrene. Phenanthrene tested negative as a complete carcinogen in a mouse study inadequately 

reported in an old secondary source (Kennaway 1924). Skin painting experiments with intermediateduration (6 months) dermal exposure were conducted on groups of 20 male C3H/HeJ mice 

(Warshawsky et al. 1993). Phenanthrene dissolved in toluene was applied to shaved skin twice weekly 

for 6 months. Tumor incidence was determined at the end of the study. For phenanthrene, 

administration alone produced papillomas in 1 of 12 animals (8%), with a mean latency period of 

100 weeks. With coadministration of 0.0005 mg benzo[a]pyrene, 1 of 17 (6%) had malignant tumors, 

with a mean latency period of 53 weeks. 

Pyrene. Skin painting experiments with intermediate-duration (6 months) exposure were conducted on 

groups of 20 male C3H/HeJ mice. Pyrene dissolved in toluene was applied to shaved skin twice 

weekly for 6 months. Tumor incidence was determined at the end of the study. For pyrene, 

administration alone produced papillomas in 1 of 13 animals (8%), with a mean latency period of 

96 weeks. With coadministration of 0.0005 mg benzo[a]pyrene, 0 of 13 animals had tumors. 

Mice chronically administered a 10% pyrene solution throughout their lifetimes did not develop skin 

tumors (Wynder and Hoffmann 1959a). However, prolonged dermal exposure of mice to 0.5% pyrene 

in decalin/n-dodecane solvent produced a slightly elevated (15%) skin carcinoma incidence; the level 

of statistical significance was not provided (Horton and Christian 1974). 

Mixtures. Chronic dermal exposure of NMRI mice to a tar condensate that contained several PAHs 

(pyrene, fluoranthene, chrysene, benz[a]anthracene, benzo[a]pyrene, indeno[ 1,2,3-c,d]pyrene, 

benzo[g,h,i]perylene) in addition to other compounds produced a carcinogenic effect as evidenced by 

an increase in the incidence of skin papillomas and carcinomas (Habs et al. 1984). Because of the 
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2. HEALTH EFFECTS 

presence of other compounds in the tar condensate, the carcinogenic effect cannot be definitely 

attributed to the PAHs present in the mixture. 

Skin painting experiments of a mixture of anthracene, chrysene, fluoranthene, phenanthrene, and 

pyrene were conducted on groups of 20 male C3H/HeJ mice (Warshawsky et al. 1993). Compounds 

dissolved in toluene were applied to shaved skin twice weekly for 6 months. Tumor incidence was 

determined at the end of the study. Treatment included solutions of a mixture of the five 

noncarcinogenic PAHs at 0.05 mg, or the same compounds in solution with 0.0005 mg 

benzo[a]pyrene, a dose known to be noncarcinogenic in a similar study design. For the mixture of the 

5 PAHs at 0.05 mg each, administration alone produced papillomas and malignant tumors in 3 of 

13 animals (23%), with a mean latency period of 73 weeks. With coadministration of benzo[a]pyrene, 

8 of 17 (47%) had tumors (papillomas and malignant), with a mean latency period of 66 weeks. 

Initiation-Promotion Studies. Carcinogenesis has been demonstrated to be a multistage process 

in the cells of certain animal tissues, including skin, lung, liver, and bladder. This process is believed 

to occur in human tumorigenesis as well. The PAHs have been studied extensively for their ability to 

act as tumor initiators and/or promoters. Following is a brief discussion, by chemical, of the results of 

the initiation-promotion studies performed with 13 of the 17 PAHs discussed in this profile. Only 

those studies considered adequate and reliable are presented in Table 2-3. 

The difficulty inherent in extrapolating initiation-promotion experiments to human exposure precludes 

their being used as the basis for human cancer effect levels. Since PAHs occur in complex mixtures 

of chemicals that may include tumor promoters, their activity as initiating agents is noteworthy. Thus, 

it is possible that humans dermally exposed to PAHs that are initiating agents, concomitantly with 

other chemicals that may be active as tumor promoters (including other PAHs) found at nearby 

hazardous waste sites, may have an increased risk of skin cancer. 

Anthracene. Anthracene has been found to be inactive as an initiating agent under a dermal 

initiation/promotion protocol using tetradecanoyl phorbol acetate (TPA) as the promoter (LaVoie et al. 

1983a). 

Benz[a]anthracene. Benz[a]anthracene has been observed to be a tumor initiator in an intermediateduration dermal study. CD-l mice topically administered 2.5 μmol (0.57 mg) benz[a]anthracene 
PAHs 73 

2. HEALTH EFFECTS 

followed by promotion with TPA (for 25 weeks) exhibited an increased skin tumor incidence (36%) as 

compared to controls (Levin et al. 1984, see Table 2-3). 

Benzo[a]pyrene. Benzo[a]pyrene is active as a tumor initiator using initiation/promotion protocols. 

Topical application of a single initiating dose of benzo[a]pyrene to the backs of mice followed by 

promotion with TPA or croton oil resulted in an 80-92% incidence of skin papillomas (Cavalieri et al. 

1988b, see Table 2-3). Ten doses of benzo[a]pyrene (0.1 mg/dose) topically applied to the backs of 

Swiss mice followed by promotion with croton oil (for 20 weeks) also resulted in the development of 

skin tumors (Hoffmann et al. 1972). 

In a dermal initiation/promotion assay, groups of 24 female SENCAR mice were administered a single 

dermal application of benzo[a]pyrene at doses ranging from 4 to 300 nmol (initiating dose), followed 

7 days later by twice weekly applications of the promoter TPA, for a total of 24 weeks. 

Benzo[a]pyrene was active as a skin tumor initiator; the number of tumors per tumor bearing mouse, 

the percentage of tumor bearing mice, and the number of tumors per mouse were all significantly 

greater than in acetone controls and increased in a dose-related manner at doses ≥20 nmol (Cavalieri et 

al. 1991). 

In a similar experiment, 24 8-week-old female SENCAR mice were treated dermally with 0.0002 mg 

(1 nmol) of benzo[a]pyrene in acetone on a shaved portion of dorsal skin (Higginbotham et al. 1993). 

One week later, tumor promotion was begun with TPA twice weekly for 27 weeks. The number of 

skin tumors was charted weekly and the mice were killed after experimental weeks. Complete 

necropsies were performed and tissues were fixed in formalin. Benzo[a]pyrene-treated mice had no 

skin tumors. 

Benzo[e]pyrene. Benzo[e]pyrene is inactive as a skin tumor initiator in mouse skin (Slaga et al. 

1980a). 

Benzo[b]fluoranthene. The ability of benzo[b]fluoranthene to initiate skin tumor formation has been 

demonstrated using a standard initiation/promotion protocol with either croton oil or phorbol myristate 

acetate as a tumor promoter (Amin et al. 1985a; LaVoie et al. 1982). In another study, dermal 

applications of initiation doses of benzo[b]fluoranthene (10-100 μg) followed by TPA (for 20 weeks) 
PAHs 74 

2. HEALTH EFFECTS 

to the backs of CD-l mice elicited a dose-related skin tumor incidence, predominantly consisting of 

squamous cell papillomas (LaVoie et al. 1982). 

In a dermal initiation/promotion assay, groups of 20 female CD-l mice were administered 10 dermal 

applications of benzo[b]fluoranthene at total doses of 0, 1, and 4 umol (initiating dose), followed 

10 days later by thrice weekly applications of the promoter TPA for a total of 20 weeks (Weyand et 

al. 1991). Benzo[b]fluoranthene was active as a skin tumor initiator; the number of tumors per 

tumor-bearing mouse (8.5) and the percentage of tumor-bearing mice (100%) were significantly greater 

than in acetone controls and were increased in a dose-related manner. 

Tumor initiation experiments were conducted with groups of 20 female CD-1 mice. 

Benzo[b]fluoranthene was applied to the shaved backs of the mice every other day using a total of 

10 subdoses (Weyand et al. 1993b). Total doses were 10.1, 30.3, and 100.9 μg in acetone. Negative 

control mice were treated with acetone. Ten days after the last dose, promotion was begun by 

applying 2.5 μg TPA thrice weekly for 20 weeks. Tumors were counted weekly. 

Benzo[b]fluoranthene produced a 35%, 90%, and 95% incidence of tumor-bearing mice with 0.45, 

3.70, and 8.65 tumors per mouse for the low, mid and high doses, respectively. No distinction was 

made between papillomas and carcinomas (Weyand et al. 1993b). Groups of 20 female Crl:CD-1 mice 

were dermally exposed to a total of 0, 0.01, 0.03, or 0.1 mg of benzo[b]fluoranthene in acetone 

applied every other day for 20 days in 10 subdoses (LaVoie et al. 1993a). Negative control mice were 

treated with acetone only. Ten days after the last application of acetone or hydrocarbon, tumor 

promotion was begun by applying 2.5 μg TPA in acetone three times weekly for 20 weeks. Tumor 

incidence was recorded after 20 weeks of promotion. The study was repeated for the 0.03 and 0.1 mg 

doses. Of the mice receiving 0.01 mg benzo[b]fluoranthene, 35% had developed tumors, with an 

average of 0.45 tumors/mouse. For the 0.03 mg dose, 70-90% of the mice developed tumors with an 

average of 1.4-3.7 tumors per mouse, whereas the 0.1 mg dose caused 95% of the mice to develop 

and average of 7.1-8.6 tumors. Of the animals treated with acetone only, 5-15% developed skin 

tumors. In this study, tumors were not identified as papillomas or carcinomas. 

Benzo[j]fluoranthene. Benzo[j]fluoranthene has also been demonstrated to be a tumor initiator in 

mice, although it is not as potent as benzo[b]fluoranthene. Benzo[j]fluoranthene, however, is more 

potent than benzo[k]fluoranthene. Mice receiving initiating doses of benzo[j]fluoranthene 

(30-1,000 μg) followed by TPA promotion exhibited a dose-related increase in tumor incidence 
PAHs 75 

2. HEALTH EFFECTS 

(LaVoie et al. 1982, 1993a; Weyand et al. 1992). Groups of 20 female Crl:CD-1 mice were dermally 

exposed to a total of 0, 0.006, 0.012, 0.025, or 0.25 mg of benzojjlfluoranthene in acetone applied 

every other day for 20 days in 10 subdoses (LaVoie et al. 1993b). Negative control mice were treated 

with acetone only. Ten days after the last application of acetone or hydrocarbon, tumor promotion 

was begun by applying 2.5 μg TPA in acetone three times weekly for 20 weeks. Tumor incidence was 

recorded after 20 weeks of promotion. Five percent of the mice receiving 0.006 mg 

benzo[j]fluoranthene had developed tumors, with an average of 0.4 tumors per mouse. For the 

0.012 mg dose, 10% of the mice developed tumors with an average of 0.4 tumors per mouse; 

0.025 mg PAH caused 45% of the mice to develop tumors at an average of 0.65 tumors per mouse, 

whereas the 0.25 mg dose caused 95% of the mice to develop an average of 8.7 tumors. The control 

animals developed no tumors. 

Benzo[k]fluoranthene. Benzo[k]fluoranthene has also been demonstrated to be a tumor initiator in 

mice, although it too, is not as potent as benzo[b]fluoranthene. Mice receiving initiating doses of 

benzo[k]fluoranthene (30-1,000 μg) followed by TPA promotion exhibited a dose-related increase in 

‘tumor incidence (LaVoie et al. 1982). 

Benzo[g,h,i]perylene. Benzo[g,h,i]perylene has been shown to be inactive as an initiating agent when 

applied at a total dose of 0.25 mg/animal and negative as a complete carcinogen when a 1% solution 

was applied thrice weekly for 12 months (IARC 1983). 

Chrysene. Chrysene is a tumor initiator in classic initiation/promotion bioassays on mouse skin using 

croton oil or phorbol myristate acetate as promoting agents (Slaga et al. 1980a; Wood et al. 1979a). 

Initiating doses of chrysene followed by promotion with TPA or croton resin induced a dose-related 

papilloma incidence in mice (Slaga et al. 1980a; Wood et al. 1979a). 

Dibenz[a,h]anthracene. Dibenz[a,h]anthracene has also demonstrated tumor-initiating activity using a 

standard initiation/promotion protocol (Slaga et al. 1980a). Dibenz[a,h]anthracene has been reported to 

initiate skin development in a dose-response relationship at doses as low as 0.028 μg followed by 

promotion with TPA (for 25 weeks) (Buening et al. 1979a). 

In a dermal initiation/promotion assay, groups of 50 female NMRI mice were administered a single 

dermal application of dibenz[a,h]anthracene at doses of 0, 300, or 600 nmol (initiating dose) followed 
PAHs 76 

2. HEALTH EFFECTS 

7 days later by twice weekly applications of the promoter, TPA, for a total of 24 weeks (Platt et al. 

1990). Dibenz[a,h]anthracene was active as a skin tumor initiator only at the highest dose tested; 93% 

of the animals administered 600 nmol dibenz[a,h]anthracene developed skin tumors by 24 weeks. 

Fluoranthene. Fluoranthene did not exhibit initiating activity in Swiss mice topically administered 

10 doses followed by promotion with croton oil (for 20 weeks) (Hoffmann et al. 1972). 

Indeno[l,2,3-c,d]pyrene. A pronounced dose-response relationship has been exhibited by 

indeno[l,2,3-c,d]pyrene in an initiation-promotion bioassay when TPA was employed as the promoting 

agent, although it was not as potent an initiator as benzo[b]fluoranthene (Rice et al. 1985a). In 

another study, 2.83 tumors/mouse were noted after a total initiating dose of 1.0 mg 

indeno[ 1,2,3-c,d]pyrene and promotion with TPA for 20 weeks (Rice et al. 1986). 

The skin tumor initiating activity of indeno[ 1,2,3-c,d]pyrene and several of its metabolites generated in 

viva in mouse skin was tested in female Crl:CD/l mice (Rice et al. 1986). Initiating doses of 

indeno[ 1,2,3-c,d]pyrene or the metabolites were applied every other day to the shaved skin of groups 

of 25 mice for a total of 10 doses, which was followed 10 days later by thrice weekly applications of 

the tumor promotor, TPA, for 20 weeks. None of the metabolites were as active in inducing skin 

tumors as the parent compound (2.83 tumors/mouse as compared to 0.48-1.68 tumors/mouse at 

20 weeks). These findings led the authors to conclude that the principal ultimate mutagenic 

metabolite, indeno[l,2,3-c,d]pyrene-1,2-oxide, is not the ultimate carcinogenic metabolite of 

indeno[l,2,3-c,d]pyrene. 

Phenanthrene. Phenanthrene was ineffective as an initiator in various mouse strains (LaVoie et al. 

1981b; Salaman and Roe 1956; Wood et al. 1979a). CD-1 mice topically administered a single dose 

of 10 μmol phenanthrene followed by a promoter were observed to have a papilloma incidence 

2-4 times that of background; however, the incidences were not statistically significant in comparison 

to controls because of the small number of animals tested and the high spontaneous tumor incidence 

(Wood et al. 1979a). 

Pyrene. Pyrene has been shown to be inactive as an initiating agent (Salaman and Roe 1956; Van 

Duuren and Goldschmidt 1976). 
PAHs 77 

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Mixtures. Co-administration of pyrene and benzo[a]pyrene to the backs of ICR/Ha mice has produced 

an enhancement of benzo[a]pyrene tumorigenicity (Van Duuren and Goldschmidt 1976). There is 

evidence that benzo[g,h,i]perylene is a co-carcinogen with benzo[a]pyrene when both are applied 

simultaneously to the skin of Swiss mice (Van Duuren et al. 1973). Dermal pretreatment with 100 μg 

pyrene substantially enhanced benzo[a]pyrene tumor initiation in CD-1 mice, while 100 μg 

fluoranthene produced a marginal enhancement (Slaga et al. 1979). 

2.3 TOXICOKINETICS 

Occupational studies provide evidence that inhaled PAHs are absorbed by humans. Animal studies 

also show that pulmonary absorption of benzo[a]pyrene occurs and may be influenced by carrier 

particles and solubility of the vehicle; however, the extent of absorption is not known. Absorption of 

benzo[a]pyrene following ingestion is low in humans, while oral absorption in animals varies among 

the PAH compounds depending on the lipophilicity. Oral absorption increases with more lipophilic 

compounds or in the presence of oils in the gastrointestinal tract. Percutaneous absorption of PAHs 

appears to be rapid for both humans and animals, but the extent of absorption is variable among these 

compounds and may be affected by the vehicle used for administration. Therefore, absorption of 

PAHs following inhalation, oral, or dermal exposure may be affected by vehicle of administration. 

There was no information available on the distribution of PAHs in humans. PAHs appear to be 

widely distributed in tissues of animals following oral and inhalation exposure; peak tissue 

concentrations occurred earlier with higher exposure levels. Placental transfer of PAHs appears to be 

limited, and therefore, fetal levels are not as high as maternal levels. 

Metabolism of PAHs occurs in all tissues and involves several possible pathways. Metabolism of 

PAHs has been studied extensively in vitro and in vivo. The metabolism products include epoxide 

intermediates, dihydrodiols, phenols, quinones, and their various combinations. The phenols, quinones, 

and dihydrodiols can all be conjugated to glucuronides and sulfate esters; the quinones also form 

glutathione conjugates. 

Quantitative data on the excretion of PAHs in humans are lacking. In general, feces is the major 

elimination route of PAHs in animals following inhalation exposure. Excretion of benzo[a]pyrene 

appears to be high following low-level exposure in rats but low in dogs and monkeys. PAHs are 
PAHs 78 

2. HEALTH EFFECTS 

eliminated to a large extent within 2 days following low- and high-level oral exposure in rats. 

Following dermal exposure, elimination of PAHs occurs rapidly in the urine and feces of guinea pigs 

and rats. 

Absorption of inhaled PAHs appears to occur through the mucous lining of bronchi, while ingested 

PAHs are taken up by the gastrointestinal tract in fat-soluble compounds. Percutaneous absorption is 

through passive diffusion. The mechanism of action of most PAHs involves covalent binding to DNA 

by PAH metabolites. The bay region diol epoxide intermediates of PAHs are currently considered to 

be the ultimate carcinogen for alternant PAHs. Once the reactive bay region epoxide is formed, it may 

covalently bind to DNA and other cellular macromolecules and presumably initiate mutagenesis and 

carcinogenesis. 

2.3.1 Absorption 

2.3.1.l Inhalation Exposure 

Absorption of PAHs in humans following inhalation exposure can be inferred from the presence of 

urinary metabolites of PAHs in workers exposed to these compounds in an aluminum plant (Becher 

and Bjorseth 1983). The high concentration of PAHs in the occupational setting did not correspond to 

the amount of PAHs deposited, metabolized, and excreted in the urine in this study. The authors 

suggested that PAHs adsorbed to airborne particulate matter may not be bioavailable and that the 

dose-uptake relationship may not be linear over the entire PAH concentration range. 

Twelve workers from a coke plant participated in an intensive skin monitoring program combined with 

personal air sampling and biological monitoring during 5 consecutive 8-hour shifts (Van Rooij et al. 

1993b). The mean concentration of total pyrene in the breathing zone air of the 12 workers ranged 

from 0.1 to 5.4 μg/m3

. The mean respiratory uptake of pyrene varied between 0.5 and 32.2 μg/day. 

Based on the estimates of the dermal and respiratory pyrene uptake, it is concluded that an average of 

75% of the total absorbed amount of pyrene enters the body through the skin. The total excreted 

amount of urinary 1-hydroxypyrene as a result of exposure to PAHs during the five consecutive work 

shifts varied between 36 and 239 nmol. Analysis indicated that dermal absorption was most important 

in contributing to 1-hydroxypyrene excretion. Of the total dose absorbed by both routes combined, 
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2. HEALTH EFFECTS 

13-49% is excreted as 1-hydroxypyrene. Variation in excretion is influenced by smoking habits, and 

consumption of alcohol (see Section 2.3.1.3). 

Eleven healthy male smokers and 11 male smokers with lung cancer between the ages of 30-60 years, 

with a smoking history of 15-25 cigarettes per day for over 10 years were involved in a study 

(Likhachev et al. 1993). Urinary excretion of benzo[a]pyrene-7,8-diol and 3-hydroxybenzo[a]pyrene 

was determined. Both benzo[a]pyrene metabolites were detected in the urine, but quantities of 

3-hydroxybenzo[a]pyrene were very low. The level of benzo[a]pyrene-7,8-diol in the urine varied 

considerably both in healthy smokers and smokers with lung cancer. However, the average value of 

this metabolite in the urine of healthy smokers was significantly higher than in the urine of lung 

cancer patients who smoked (1.06 mg/kg/day versus 0.56 mg/kg/day). 

Animal studies on inhalation absorption of PAHs are limited to benzo[a]pyrene exposure. Rapid 

absorption was evident following inhalation exposure of low and high levels of benzo[a]pyrene to rats. 

Acute and intermediate-duration exposure to 4.8 mg/m3 [

14C]-benzo[a]pyrene by nose-only inhalation 

in rats resulted in elevated levels of radioactivity in tissues and excreta within 3 hours of exposure 

(Wolff et al. 1989c). High levels of radioactivity were detected in the gastrointestinal tract, which 

may be due to biliary excretion or mucocilliary clearance of benzo[a]pyrene from the upper respiratory 

tract. Intratracheal administration of 0.001 mg/kg [3H]-benzo[a]pyrene to rats also resulted in rapid 

absorption through the lungs. Radioactivity in the liver reached a maximum of 21% of the 

administered dose within 10 minutes of instillation (Weyand and Bevan 1986, 1988). Presence of 

radioactivity in other tissues and the bile was also indicative of its absorption in rats. Similar results 

were also seen in guinea pigs and hamsters following intratracheal exposure to benzo[a]pyrene 

(Weyand and Bevan 1986, 1987b, 1988). 

Pregnant Wistar rats were exposed head-only to 200, 350, 500, 650, or 800 mg/m3

 of 

[

14C]-benzo[a]pyrene aerosol for 95 minutes on gestational day 17 (Withey et al. 1993a). Animals 

were killed immediately or 6 hours postexposure. Concentration of benzo[a]pyrene and metabolites in 

maternal blood sampled immediately after exposure were elevated 10-fold over the 4-fold increase in 

dose. At 6 hours postdosing, the increase was still approximately 10-fold, although the actual 

concentrations were 2-7-fold less than at 0 hours. Concentrations of benzo[a]pyrene and metabolites 

in fetal blood sampled immediately after exposure were elevated 5-fold over the 4-fold increase in 

exposure concentrations. Fetal tissues sampled 6 hours post-dosing had a 9-fold increase in 
PAHs 80 

2. HEALTH EFFECTS 

benzo[a]pyrene and metabolite concentration over the dose range, due to lower concentration in the 

200-650 mg/m3 dose groups at 6 hours compared to 0 hours. Fetal concentrations were 2-10-fold less 

than maternal concentrations. Benzo[a]pyrene concentrations in blood, lung, liver, and fetal tissues 

were significantly decreased from 0 to 6 hours postexposure while levels in fat tissue increased. For 

benzo[a]pyrene sampled immediately postdosing, lung > blood > liver > kidney > fat > fetus. For 

total metabolites sampled immediately postdosing, lung > blood > liver > kidney > fetus > fat. For 

benzo[a]pyrene sampled 6 hours postdosing, fat > lung > kidney > liver > blood > fetus. For total 

metabolites sampled 6 hours postdosing, lung = fat > kidney > liver = blood > fetus. 

Twenty male Fisher 344 DuCrj rats were divided into two groups and exposed to diesel exhaust 

containing 0.151 mg/m3

 pyrene or HEPA-filtered air for 8 weeks, 5 days/week, 7 hours/day (Kanoh et 

al. 1993). At 2, 4, and 8 weeks during the exposure, the rats from each group were put into a 

metabolic cage and their urine was collected for 24 hours. Urinary levels of 1-hydroxypyrene in the 

rats of the exposure group increased remarkably over those of the control group, reaching 2.4 times as 

much by the end of the 2nd week, and 5.6 times by the 4th and 8th weeks. 

Inhalation absorption of benzo[a]pyrene may be affected by the size of particles on which 

benzo[a]pyrene is adsorbed. The elimination of benzo[a]pyrene from the lungs was studied following 

intratracheal administration of pure benzo[a]pyrene crystals or benzo[a]pyrene coated on carbon 

particles in two size ranges (0.5-1.0 μm and 15-30 μm) (Creasia et al. 1976). Fifty percent of the 

pure benzo[a]pyrene crystals was cleared from the lungs within 1.5 hours and >95% cleared within 

24 hours, while only 50% of the benzo[a]pyrene adsorbed to the small carbon particles cleared within 

36 hours. Elution of benzo[a]pyrene was even slower with the larger particle size (approximately 

4-5 days). These results indicate that the bioavailability of benzo[a]pyrene is altered by the particle 

size of the carrier. The initial lung deposition of [3

H]-benzo[a]pyrene adsorbed onto gallium oxide 

(Ga2O3) particles was 4.9 μg of which 3.1% remained after 30 minutes (Sun et al. 1982). A control 

study, conducted without the Ga2O3 particles at a concentration of 1 mg/m3

, found that 8.2 μg was 

inhaled, of which 0.9% remained in lungs after 30 minutes. The excretion of hydrocarbon was 

monitored for over 2 weeks at which time nearly all the initial lung burden was recovered in the 

excreta, indicating complete absorption of the instilled hydrocarbon. Significant differences in the 

clearance of benzo[a]pyrene coated with Ga2O3 and pure benzo[a]pyrene suggested that a substantial 

amount of benzo[a]pyrene/Ga2O3 particles was removed from the lungs by mucocilliary clearance and 

subsequent ingestion. 
 

PAHs 81 

2. HEALTH EFFECTS 

The above results corroborate findings in an in vitro experiment by Gerde and Scholander (1989) who 

developed a model of the bronchial lining layer. These investigators concluded that the release rate of 

PAHs from carrier particles is the rate-determining step in the transport of PAHs from these particles 

to the bronchial epithelium. 

The absorption of benzo[a]pyrene may also be affected by the solubility of the vehicle used in 

administration. Approximately 70% of benzo[a]pyrene administered with triethylene glycol was 

excreted 6 hours following intratracheal instillation (Bevan and Ulman 1991). Excretion rates of 

benzo[a]pyrene were only 58.4% and 56.2% with ethyl laurate and tricaprylin, respectively, within a 

6-hour period. The small volume of benzo[a]pyrene instilled is probably deposited in the bronchial 

region which allows more water-soluble materials (triethylene glycol) to pass the mucous layer lining 

than water-insoluble compounds (ethyl laurate and tricaprylin) (Bevan and Ulman 1991). 

Nasal instillation of [3

H]-benzo[a]pyrene (0.13 mg/kg) to hamsters resulted in the metabolism of 

[

3

H]-benzo[a]pyrene in the nasal cavity (Dahl et al. 1985). A large fraction of the metabolites was 

recovered from the epithelial surface, indicating that benzo[a]pyrene was first absorbed in the mucosa, 

metabolized, and returned to the mucus. Monkeys and dogs received nasal instillation of 

[

14C]-benzo[a]pyrene at doses of 0.16-0.21 mg/kg (Petridou-Fischer et al. 1988). Radiolabeled 

metabolites were detected in the nasal cavity, but little or no activity was detected in the blood and 

excreta of either species during the 48 hours after exposure. These results indicate that absorption of 

benzo[a]pyrene and/or its metabolites was poor or very slow following nasal instillation in monkeys 

and dogs. 

2.3.1.2 Oral Exposure 

There is evidence suggesting that benzopyrene is orally absorbed in humans (Buckley and Lioy 1992; 

Hecht et al. 1979). Following ingestion of diets containing very low levels of benzo[a]pyrene, the 

metabolite, 1-hydroxypyrene, was detected in the urine (Buckley and Lioy 1992). No quantitative data 

on the excretion of the benzo[a]pyrene were provided. The concentration of benzo[a]pyrene in human 

feces was examined in eight volunteers who ingested broiled meat that contained approximately 9 μg 

of benzo[a]pyrene (Hecht et al. 1979). The feces of these individuals did not contain detectable levels 

of benzo[a]pyrene (<0.1 μg/person), which is similar to what was seen following consumption of 
PAHs 82 

2. HEALTH EFFECTS 

control meat that contained undetectable amounts of benzo[a]pyrene by these same volunteers, 

suggesting that most of the ingested benzo[a]pyrene was absorbed. 

Oral absorption of benzo[a]pyrene in rats is incomplete and may be influenced by the presence of oils 

and fat in the gastrointestinal tract. Oral absorption of benzo[a]pyrene was estimated to be 40%, with 

a bioavailability of 7.8-11.5%, in Sprague-Dawley rats infused intraduodenally to a total dose of 

approximately 0.0005 mg/kg for 90 minutes (Foth et al. 1988a). Nearly 80% of a gavage dose of 

0.0527 mg/kg [14C]benzo[a]pyrene in peanut oil was detected in the excreta of rats 48 hours after 

exposure; however, some of the recovered radioactivity may never have been absorbed by the 

alimentary tract of the rats, but may have passed into the excreta in the peanut oil (Hecht et al. 1979). 

Radioactivity found in the liver, lungs, kidneys, and testis following a low dose of 

[

3

H]-benzo[a]pyrene to Sprague-Dawley rats provides supporting evidence of oral absorption 

(Yamazaki and Kakiuchi 1989; Yamazaki et al. 1987). The extent of oral absorption in rats is 

enhanced when benzo[a]pyrene is solubilized in a vehicle (triolein, soybean oils, high-fat diet) that is 

readily absorbed following low- and high-dose levels (Kawamura et al. 1988; O’Neill et al. 1991). 

Oral absorption of benzo[a]pyrene was estimated to be 38-58% following dietary or gavage exposure 

to high levels in rats (Chang 1943). Anthracene was absorbed to a slightly higher extent (53-74%) 

than benzo[a]pyrene in rats while phenanthracene was poorly absorbed (4-7%) (Chang 1943). 

However, the data were limited because an inadequate number of rats was used and study details were 

lacking. 

In general, the oral absorption of chrysene, dibenzanthracene, and pyrene was high following exposure 

to high doses in rats (Chang 1943; Grimmer et al. 1988; Withey et al. 1991). Following dietary or 

gavage administration of chrysene in rats, 64-87% of the dose was excreted in the feces (Chang 1943). 

Recovery of chrysene in excreta of Wistar rats was 74% four days after a single gavage dose of 

22 mg/kg chrysene in corn oil (Grimmer et al. 1988). Administration of dibenz[a,h]anthracene in the 

diet (250 mg) or by stomach tube (200 mg) resulted in more than 90% of the dose being excreted in 

the feces of white rats (Chang 1943). As with chrysene, absorption of dibenz[a,h]anthracene could not 

be quantified. Male Wistar rats administered 2-15 mg/kg of [14C]-pyrene recovered 68-92% of the 

dose in the excreta by 6 days postexposure (Withey et al. 1991). Bioavailability of pyrene and its 

metabolites was 65-84% over a period of 8 hours following administration. 
PAHs 83 

2. HEALTH EFFECTS 

Blood levels of fluoranthene, pyrene, and benz[a]anthracene after oral administration were examined in 

rats (Lipniak and Brandys 1993). Fluoranthene, pyrene, or benz[a]anthracene in Tween 80/isotonic 

saline, was administered orally to rats at a dose of 20 mg/kg. Blood levels after administration 

indicated that peak concentrations of the three compounds were reached at 1-2 hours after 

administration. The peak concentration of fluoranthene (≈30 mg/cm3) was twice as high as that of 

pyrene, and 5 times higher than benz[a]anthracene. 

The effect of diet matrix (gel or powder) on urinary excretion of 1-hydroxypyrene and hydrocarbon 

binding to DNA was investigated in mice (Wu et al. 1994). Female mice were fed a gel or powder 

diet containing manufactured gas plant (MGP) residue (coal tar) at 0.1% or 0.3% for 15 days, or the 

corresponding control diet. Two mixtures of MGP residue were used: Mix of 3 combining equal 

amounts of samples from three different MGP plant sites, and Mix of 7 combining equal amounts of 

samples from seven different MGP plant sites, including those used in the Mix of 3. The mixtures 

contained pyrene, benz[a]anthracene, chrysene, benzo[b]fluorene, benzo[k]fluorene, benzo[a]pyrene, 

indeno[ 1,2,3-cd]pyrene, dibenz[a,h]anthracene, benzo[g,h,i]perylene. Data were presented in terms of 

pyrene consumed and 1-hydroxypyrene excreted. Urine was collected on the first, seventh, and 

fourteenth day of diet administration. 1-Hydroxypyrene levels in the urine were determined using 

HPLC and fluorescence. Diet matrix had little effect on the bioavailability of the PAHs. 

The intestinal absorption of PAHs is highly dependent on the presence of bile (Rahman et al. 1986). 

To study the role of bile in the intestinal absorption of PAHs, conscious rats with bile duct and 

duodenal catheters were given [3

H]-benzo[a]pyrene, phenanthrene, anthracene, 2,6-dimethylnaphthalene 

(DMN), and 7,12-dimethylbenz[a]anthracene (DMBA) with or without exogenous bile. The efficiency 

of PAH absorption was estimated from the cumulative recovery of radioactivity in the bile and urine 

over 24 hours. The efficiencies of absorption without bile (as a percentage of absorption with bile) 

were benzo[a]pyrene, 22.9%; phenanthrene, 96.7%; anthracene, 70.8%; DMN, 91.6%; DMBA, 43.4%. 

Absorption of the four- or five-membered rings (DMBA and benzo[a]pyrene) was strongly dependent 

on the presence of bile in the intestinal lumen. The absorption of the tricyclic PAHs (phenanthrene 

and anthracene) differed with respect to their dependency on bile for efficient absorption. This 

difference correlated with a difference in water solubility, with anthracene being 18 times less 

water-soluble than phenanthrene. Those products with low water solubility are dependent on the 

creation of an intermediate phase of the products of lipolysis and bile salts (Rahman et al. 1986). 

These reactions occur during the normal process of lipid digestion and absorption in the intestine. 
 

PAHs 84 

2. HEALTH EFFECTS 

2.3.1.3 Dermal Exposure 

Application of 2% crude coal tar to the skin of humans for 8-hour periods on 2 consecutive days 

yielded evidence of PAH absorption (Storer et al. 1984). Phenanthrene, anthracene, pyrene, and 

fluoranthene were detected in the blood, but benzo[a]pyrene (which is present in coal tar) was not 

detected. This difference was attributed to differences in percutaneous absorption, rapid tissue 

deposition after absorption, or metabolic conjugation with rapid urine excretion. In another study, coal 

tar ointment was applied to skin of volunteers at various sites (Van Rooij et al. 1993a). The surface 

disappearance of PAH and the excretion of urinary 1-hydroxy-pyrene were used as parameters for 

dermal absorption. Surface disappearance measurements showed low but significant differences in 

dermal PAH absorption between anatomical sites: shoulder > forehead, forearm, groin > ankle, hand 

(palmar site). An in vitro study using human skin revealed that the extent of permeation across viable 

human skin after 24 hours was estimated to be 3% of the total applied radioactivity from 

[

14C]-benzo[a]pyrene (10 μg/cm2

) (Kao et al. 1985). Using human cadaver skin, it was shown that 

23.7±9.7% of the applied benzo[a]pyrene penetrated into the skin (Wester et al. 1990). These results 

suggest that substantial metabolism and/or binding of benzo[a]pyrene takes place in viable human skin 

which limits the amount of PAH available to penetrate the skin into the systemic circulation. 

Twelve workers from a coke plant participated in an intensive skin monitoring program combined with 

personal air sampling and biological monitoring during 5 consecutive 8-hour shifts (Van Rooij et al. 

1993b). Measurements on exposure pads at six skin sites (jaw/neck, shoulder, upper arm, wrist, groin, 

ankle) showed that mean total skin contamination of the 12 workers ranged between 21 and 166 μg 

pyrene per day. The dermal uptake of pyrene ranged between 4 and 34 μg/day, which was about 20% 

of the pyrene contamination on the skin. The mean concentration of total pyrene in the breathing zone 

air of the 12 workers ranged from 0.1 to 5.4 μg/m3

. The mean respiratory uptake of pyrene varied 

between 0.5 and 32.2 μg/day. Based on the estimates of the dermal and respiratory pyrene uptake, it 

is concluded that an average of 75% of the total absorbed amount of pyrene enters the body through 

the skin. The total excreted amount of urinary 1-hydroxy-pyrene as a result of exposure to PAHs 

during the 5 consecutive work shifts varied between 36 and 239 nmol. Analysis indicated that dermal 

absorption was most important in contributing to 1-hydroxy-pyrene excretion. Of the total dose 

absorbed by both routes combined, 13-49% is excreted as 1-hydroxy-pyrene. Variation in excretion is 

influenced by smoking habits, and consumption of alcohol. 
PAHs 85 

2. HEALTH EFFECTS 

Percutaneous absorption of [14C]-benzo[a]pyrene in mice, rats, monkeys, and guinea pigs is rapid and 

high (Ng et al. 1992; Sanders et al. 1986; Wester et al. 1990; Yang et al. 1989). A single dose of 

7 μg/cm2

 [14C]-benzo[a]pyrene in acetone was applied to a 4-cm2 area of the dorsal skin of female 

hairless guinea pigs for 24 hours (Ng et al. 1992). Approximately 73% of the administered dose was 

absorbed dermally by 7 days postexposure; most of the dose was absorbed by day 3. The skin wash 

at 24 hours of exposure contained about 10.6% of dose (Ng et al. 1992). In vitro absorption of 

benzo[a]pyrene through guinea pig skin demonstrated similar results; 67% absorption in a 24-hour 

exposure (Ng et al. 1992). Seven days after exposure to 125 μg/cm2

 benzo[a]pyrene, 80% of the total 

recovered radioactivity was eliminated in the feces of mice (Sanders et al. 1986). The site of 

application still retained 7% of the recovered radioactivity after 7 days. However, the area of 

application was not covered to prevent animals from licking the test material which may have lead to 

ingestion of benzo[a]pyrene. 

Groups of 12 male Wistar rats were dosed with 2, 6, or 15 mg/kg of [14C]-pyrene applied to 4 cm2 of 

a shaved area of the mid-back (Withey et al. 1993b). Three animals in each dose group were killed at 

1,2,4, and 6 days postdosing, and the brain, lungs, heart, liver, spleen, kidneys, testes, muscle, and 

perirenal fat were removed and analyzed for pyrene and [ 14C]-pyrene equivalents. Blood, urine, and 

feces, as well as the skin from the application site were also analyzed. The rate of uptake from the 

skin was rapid (t1/2 = 0.5-0.8 d) relative to rate processes for the other organs, and about 50% of the 

applied dose was excreted over the 6 days of the study. Levels of pyrene were highest in the liver, 

kidneys, and fat. Levels of metabolites were also high in the lung. 

Dermal absorption of benzo[a]pyrene in rats and monkeys may be affected by the vehicle of 

administration (Wester et al. 1990; Yang et al. 1989). Following application of 10 ppm 

benzo[a]pyrene on the skin of rhesus monkeys, an average absorption of 51 ± 22% was reported with 

acetone vehicle and 13.2 ± .4% with soil; however, absorption data were based on radioactivity 

recovered in urine only, and not in feces (Wester et al. 1990). The great variation in the absorption 

with the acetone vehicle limits these results. This may be related, in part, to the dependence on 

monitoring radioactivity recovered in urine only as opposed to monitoring radioactivity recovered in 

urine and feces. Disappearance of the applied dose from the application site was 40% at 24 hours 

following administration (Wester et al. 1990). Sprague-Dawley rats absorbed 4-5 times more of a 

1 ppm dose of benzo[a]pyrene when it was applied dermally alone, compared to a soil-sorbed crude 

oil mixture (Yang et al. 1989). The greater lipophilicity of the crude oil alone probably increased the 
PAHs 86 

2. HEALTH EFFECTS 

rate of dermal uptake of the test material, since the authors determined that absorption was dependent 

on the monolayer of soil in contact with the skin, which is comparable to the contact between the skin 

and oil. Therefore, the soil binding of the PAHs may have slowed absorption. However, no 

quantitative data were available. 

Female Sprague-Dawley rats were exposed dermally to [3H]-benzo[a]pyrene (1 ppm) containing 

petroleum crude oil alone or in fortified soil matrix for 4 days (Yang et al. 1989). Recovery of 

radioactivity was 35.3% of the dose applied in oil, as follows: urine (5.3% of dose), feces (27.5%), 

and tissues (2.5%) 96 hours after beginning of exposure. Recovery was 9.2% of applied dose with 

benzo[a]pyrene from petroleum crude-fortified soil; recoveries in urine, feces, and tissues were 1.9%, 

5.8%, and 1.5%, respectively, at 96 hours. Benzo[a]pyrene (10 ppm) with acetone vehicle or in soil 

was applied to a 12 cm2

 area of abdominal skin of female rhesus monkeys for 24 hours (Wester et al. 

1990). Urine contained 5l ± 22% of the dose with acetone vehicle and 13.2 ± 3.4% with soil. 

The percutaneous absorption of [14C]-anthracene (9.3 μg/cm2) was 52.3% in rats, estimated from 

radioactivity in urine, feces, and tissues over a 6-day period (Yang et al. 1986). Over time, the 

permeation of anthracene significantly decreased suggesting that anthracene was dermally absorbed in 

a dose-dependent manner. Diffusion of anthracene through the skin (stratum corneum) depended on 

the amount of anthracene on the skin’s surface. 

When 6.25 μg/cm2

 [14C]-phenanthrene and pyrene was applied to guinea pigs, dermal absorption was 

80% and 94%, respectively (Ng et al. 1991, 1992). In vitro absorption of phenanthrene and pyrene in 

guinea pig skin was about 79-89% and 70%, respectively (Ng et al. 1991, 1992). 

Monitoring the removal of compounds from the epidermis is indicative of measuring the compound’s 

dermal absorption. The disappearance of radiolabeled benzo[a]pyrene and its metabolites from the 

epidermis was monophasic, following first order kinetics with a half-life of approximately 2 hours 

(Melikian et al. 1987). Recovery of the radiolabel was 99-100% throughout the period of the 

experiment (8 hours), indicating that volatilization of benzo[a]pyrene from the skin was not a 

confounding factor (Melikan et al. 1987). In contrast, removal of one of its metabolites, 

7,8-dihydroxy-9,10-epoxy-7,8,9,l0-tetrahydrobenzo[a]pyrene (anti-BPDE), from the epidermis was 

biphasic. The second, slower phase of removal suggested that the stratum comeum, the outermost 

layer of skin which consists of several layers of inactive, keratinized cells surrounded by extracellular 
PAHs 87 

2. HEALTH EFFECTS 

lipids, may act as a reservoir that can retain and slowly release topically applied lipophilic substances 

such as benzo[a]pyrene but is penetrated rapidly by more polar compounds such as anti-BPDE. 

2.3.2 Distribution 

2.3.2.1 Inhalation Exposure 

No studies were located regarding the distribution of PAHs in humans following inhalation exposure. 

In general, tissue distribution of benzo[a]pyrene following inhalation exposure is qualitatively similar 

for different species (Bevan and Weyand 1988; Weyand and Bevan 1986, 1987a, 1988; Wolff et al. 

1989c). Highest radioactivity was distributed to the cecum, small intestine, trachea, kidneys, and 

stomach in rats following a 3-hour or 4-week inhalation exposure to 4.8 mg/m3

 [14C]-benzo[a]pyrene 

(Wolff et al. 1989c). The lungs and liver of rats contained 2.7% and 4.6% of the recovered dose 

6 hours after intratracheal administration of 0.001 mg/kg [3H]-benzo[a]pyrene (Bevan and Weyand 

1988). [3H]-Benzo[a]pyrene intratracheally administered to rats demonstrated that the highest fractions 

were distributed to the lung, liver, kidney, gastrointestinal tract, and carcass (Weyand and Bevan 1986, 

1987a, 1988). The concentration of benzo[a]pyrene and its metabolites in the intestine increased with 

time, suggesting the occurrence of biliary excretion and enterohepatic recirculation. Tissue distribution 

of radioactivity was qualitatively similar in guinea pigs and hamsters (Weyand and Bevan 1987b). 

Mice that received 2.5 mg/kg benzo[a]pyrene intratracheally also experienced a similar tissue 

distribution, but Schnizlein et al. (1987) noted that as the lung burden of benzo[a]pyrene diminished, 

radioactivity continued to increase in the lung-associated lymph nodes for 6 days. This accumulation 

may eventually affect humoral immunity. 

The distribution pattern of benzo[a]pyrene was not significantly affected following aerosol exposure 

with or without Ga2O3 particles (Sun et al. 1982). However, significant differences in the levels of 

benzo[a]pyrene delivered to the different tissues did exist. Maximum levels were achieved in the liver, 

esophagus, small intestine, and blood at 30 minutes following exposure. At 12 hours, maximum levels 

were seen in the lower gastrointestinal tract. Higher tissue levels of hydrocarbon resulted from 

absorption of benzo[a]pyrene-Ga2O3 particles. Inhaled benzo[a]pyrene adsorbed on insoluble Ga2O3 

particles was cleared predominantly by mucocilliary transport and ingestion. This latter mechanism of 
 

PAHs 88 

2. HEALTH EFFECTS 

absorption led to the increased levels and longer retention times of benzo[a]pyrene in the stomach, 

liver, and kidneys (Sun et al. 1982). 

Pregnant Wistar rats inhaled head-only high levels of [14C]-pyrene aerosol on gestational day 

17 (Withey et al. 1992). Concentrations of pyrene and metabolites in maternal and fetal blood were 

elevated 8-fold with a fourfold increase in exposure concentrations. However, pyrene levels in fetal 

blood were about 10 times lower than maternal blood immediately after exposure. In general, 

radioactivity increased in fat but decreased in blood, lungs, and liver tissues from 0 to 6 hours 

postexposure. There was only a small increase in the concentration of radioactivity in fetuses over the 

whole exposure range compared to maternal levels, suggesting placental transfer of pyrene and its 

metabolites are limited or that metabolism in fetal tissues is limited. 

In a similar study, pregnant Wistar rats were exposed head-only to 200, 350, 500, 650, or 

800 mg/m3 of [14C]-benzo[a]pyrene aerosol for 95 minutes on gestational day 17 (Withey et al. 

1993a). Animals were killed immediately or 6 hours postexposure. Concentration of benzo[a]pyrene 

and metabolites in maternal blood sampled immediately after exposure were elevated 10-fold over the 

4-fold increase in dose. At 6 hours postdosing, the increase was still approximately 10-fold, although 

the actual concentrations were 2-7-fold less than at 0 hours. Concentrations of benzo[a]pyrene and 

metabolites in fetal blood sampled immediately after exposure were elevated 5-fold over the 4-fold 

increase in exposure concentrations. Fetal tissues sampled 6 hours post-dosing had a 9-fold increase in 

benzo[a]pyrene and metabolite concentration over the dose range, due to lower concentration in the 

200-650 mg/m3 dose groups at 6 hours compared to 0 hours. Fetal concentrations were 2-10-fold less 

than maternal concentrations. Benzo[a]pyrene concentrations in blood, lung, liver, and fetal tissues 

were significantly decreased from 0 to 6 hours postexposure while levels in fat tissue increased. For 

benzo[a]pyrene sampled immediately postdosing, lung > blood > liver > kidney > fat > fetus. For 

total metabolites sampled immediately postdosing, lung > blood > liver > kidney > fetus > fat. For 

benzo[a]pyrene sampled 6 hours postdosing, fat > lung > kidney > liver > blood > fetus. For total 

metabolites sampled 6 hours postdosing, lung = fat > kidney > liver = blood > fetus. 

2.3.2.2 Oral Exposure 

No studies were located regarding the distribution of PAHs in humans following oral exposure. 
PAHs 89 

2. HEALTH EFFECTS 

Tissue levels of benzo[a]pyrene in Sprague-Dawley rats were highest 2-8 days after initial exposure to 

multiple doses of 0.0005 mg [3H]-benzo[a]pyrene (Yamazaki and Kakiuchi 1989). The highest 

radioactivities were found in the kidney and testis. [3

H]-Benzo[a]pyrene distributed to the protein 

fractions of the liver, lung, and kidney (Yamazaki et al. 1987). The radioactivity in the protein 

fractions of these tissues increased gradually over time. In contrast, the radioactivity in the lipid 

fractions of these tissues accounted for 70% of the administered dose at 3 hours but decreased rapidly 

with time. The nucleic acid fraction maintained approximately 10% of the total radioactivity 

throughout the experiment. The increase in protein binding of radioactivity associated with 

benzo[a]pyrene and its metabolites, and the persistence of the radioactivity associated with the protein 

fractions, suggests that protein binding may allow benzo[a]pyrene and its metabolites to accumulate in 

certain tissues, thus increasing the likelihood of cytotoxicity, mutagenicity, or carcinogenicity of 

benzo[a]pyrene and its metabolites in these organs. In addition, these organs have low metabolic 

activity while the liver has a high detoxification potential and can facilitate the excretion of these toxic 

products (Yamazaki et al. 1987). 

Single oral doses of 12 mg/kg [14C]-benzo[a]pyrene were administered by gavage to pregnant 

NMRI:Han mice on gestational days 11, 12, 13, or 18 (Neubert and Tapken 1988). Distribution of 

radioactivity was measured at 6, 24, and 48 hours after exposure. Maternal and embryo levels were 

highest with exposure on gestational day 11. The difference in radioactivity between maternal and 

embryo liver tissues increased when exposure occurred at later gestation. In another experiment, mice 

were exposed to 24 mg/kg benzo[a]pyrene for 3 consecutive days during early (gestational days 9-l1) 

or late gestation (days 15-17) (Neubert and Tapken 1988). Maternal tissue levels were not much 

different from those observed following the administration of single doses. After multiple dose 

administration, elimination appeared to be faster in maternal tissues, but slower in embryonic tissues. 

Placental levels were always higher than those in embryonic tissue. Results suggest that 

benzo[a]pyrene does not cross the placental barrier readily and, therefore, that levels in embryonic 

tissues of mice never reach levels found in maternal tissues. 

In general, orally absorbed benz[a]anthracene, chrysene, and pyrene were rapidly and widely 

distributed in the rat (Bartosek et al. 1984; Withey et al. 1991). Maximum concentrations of 

benz[a]anthracene and chrysene in perfused tissues, like the liver, blood, and brain, were achieved 

within l-2 hours after administration of high doses (76 and 152 mg/kg) (Bartosek et al. 1984). 

Maximum levels in lesser perfused tissues, like adipose and mammary tissue, were reached in 
 

PAHs 90 

2. HEALTH EFFECTS 

3-4 hours. In male Wistar rats receiving a gavage dose of 2-15 mg/kg of [14C]-pyrene, the fat had 

the highest tissue levels of radioactivity, followed by kidney, liver, and lungs (Withey et al. 1991). 

Orally absorbed dibenz[a,h]anthracene in rats was also widely distributed to several tissues (Daniel et 

al. 1967). However, maximum tissue concentrations were not reached until 10 hours after 

administration. Highest tissue concentrations were in the liver and kidneys, followed by adrenal 

glands, ovaries, blood, and fat. Soon after administration, large quantities of dibenz[a,h]anthracene 

were found in the liver and kidneys. The elimination rate from these organs was rapid. At 3-4 days 

after administration, dibenz[a,h]anthracene was distributed only in the adrenal glands, ovaries, and fat. 

The permeability of the placenta to dimethylbenz[a]anthracene, benzo[a]pyrene, and 

3-methylcholanthrene (MC) was compared by Shendrikova and Aleksandrov (1974). Pregnant rats 

received the PAH orally in sunflower oil at a dose of 200 mg/kg on the 21st day of pregnancy. 

Within 30 minutes after administration of dimethylbenz[a]anthracene, trace amount of the compound 

could be detected in the fetus. Maximum levels (1.53-1.6 μg/g) were reached 2-3 hours after 

administration. Only trace amounts were detected in the fetus at 5 hours after administration. 

Concentration profiles in the liver and placenta were similar to those seen in the fetus. 

Benzo[a]pyrene was detected in the fetus at 2.77 μg/g. MC was only present in trace amounts. 

Concentration differences in the fetus among the various PAHs appeared to be highly dependent on the 

gastrointestinal absorption of the compound. The difference in fetal concentration of the PAHs did not 

reflect their ability to permeate the placenta. 

2.3.2.3 Dermal Exposure 

No studies were located regarding the distribution of PAHs in humans following dermal exposure. 

Evidence regarding the distribution of PAHs in animals following dermal exposure is limited. 

Although PAHs can readily penetrate the skin, there are few data on distribution to tissues. In one 

published study on this subject, only 1.3% of an applied dose of [14C]-anthracene (9.3 μg/cm2

) was 

detected in tissues, primarily liver and kidneys, of rats 6 days after administration (Yang et al. 1986). 

Groups of 12 male Wistar rats were dosed with 2, 6, or 15 mg/kg of [14C]-pyrene applied to 4 cm2

 of 

a shaved area of the mid back (Withey et al. 1993b). Three animals in each dose group were killed at 
PAHs 91 

2. HEALTH EFFECTS 

1, 2, 4, and 6 days postdosing, and the brain, lungs, heart, liver, spleen, kidneys, testes, muscle, and 

perirenal fat were removed and analyzed for pyrene and [14C]-pyrene equivalents. Blood, urine, feces, 

as well as the skin from the application site were also analyzed. The rate of uptake from the skin was 

rapid (t1/2 = 0.5-0.8 d) relative to rate processes for the other organs, and about 50% of the applied 

dose was excreted over the 6 days of the study. Levels of pyrene were highest in the liver, kidneys, 

and fat. Levels of metabolites were also high in the lung. 

2.3.3 Metabolism 

The lipophilicity of PAHs enables them to readily penetrate cellular membranes and remain in the 

body indefinitely. However, the metabolism of PAHs renders them more water-soluble and more 

excretable. Metabolism of PAHs occurs in all tissues. The metabolic process involves several 

possible pathways with varying degrees of enzyme activities. The activities and affinities of the 

enzymes in a given tissue determine which metabolic route will prevail. 

The metabolism of PAHs has been studied extensively in vitro and in vivo. The most commonly used 

system is the rat liver microsomal fraction, although other species are also used. Cells and cultured 

tissues from human and other animals have also significantly contributed to the elucidation of the PAH 

metabolic scheme. 

The structural similarity of PAHs contributes to the similarities that exist in their biotransformation. 

Benzo[a]pyrene metabolism has been extensively reviewed and will be used as a model for PAH 

metabolism. In the many microsomal, cell, and cultured tissue preparations that have been examined, 

the metabolic profiles are qualitatively similar. However, there are differences in the relative levels 

and rates of formation of specific metabolites among tissues and cell preparations used from various 

animal species and strains. These differences are susceptible to change as a result of pretreatment of 

the animals with either inducers or inhibitors of particular enzymes. Furthermore, it is known that the 

metabolism of alternant PAHs (such as benzo[a]pyrene, benz[a]anthracene, chrysene, and 

dibenz[a,h]anthracene) differs from nonalternant PAHs (such as benzo[b]fluoranthene, 

benzo[k]fluoranthene, benzo[j]fluoranthene, and indeno[ 1,2,3-c,d]pyrene) (see Section 2.2). Therefore, 

the metabolism of benzo[b]fluoranthene will also be discussed as a model for nonalternant PAH 

metabolism. 
PAHs 92 

2. HEALTH EFFECTS 

The metabolism of benzo[a]pyrene is summarized in Figure 2-3. Benzo[a]pyrene is metabolized 

initially by the microsomal cytochrome P-450 systems to several arene oxides. Once formed, these 

arene oxides may rearrange spontaneously to phenols, undergo hydration to the corresponding 

trans-dihydrodiols in a reaction catalyzed by microsomal epoxide hydrolase, or react covalently with 

glutathione, either spontaneously or in a reaction catalyzed by cytosolic glutathione-S-transferases 

(IARC 1983). Phenols may also be formed by the P-450 system by direct oxygen insertion, although 

unequivocal proof for this mechanism is lacking. 6-Hydroxybenzo[a]pyrene is further oxidized either 

spontaneously or metabolically to the 1,6-, 3,6-, or 6,12-quinones. This phenol is also a presumed 

intermediate in the oxidation of benzo[a]pyrene to the three quinones catalyzed by prostaglandin 

endoperoxide synthetase (Panthanickal and Marnett 1981). Evidence exists for the further oxidative 

metabolism to two additional phenols. 3-Hydroxybenzopyrene is metabolized to the 3,6-quinone and 

9-hydroxy-benzo[a]pyrene is further oxidized to the K-region 4,5-oxide, which is hydrated to the 

corresponding 4,5-dihydrodiol (4,5,9-triol). The phenols, quinones, and dihydrodiols can all be 

conjugated to glucuronides and sulfate esters; the quinones also form glutathione conjugates (Agarwal 

et al. 1991; IARC 1983). 

In addition to being conjugated, the dihydrodiols undergo further oxidative metabolism. The 

cytochrome P-450 system metabolizes benzo[a]pyrene-4,5-dihydrodiol to a number of uncharacterized 

metabolites, while the 9,10-dihydrodiol is metabolized predominantly to its l- and/or 3-phenol 

derivative with only minor quantities of a 9,10-diol-7,8-epoxide being formed. In contrast to the 

9,10-diol, benzopyrene-7,8-diol is metabolized to a 7,8-dihydrodiol-9,10-epoxide, and phenol-diol 

formation is a relatively minor pathway. The diol epoxides can be conjugated with glutathione either 

spontaneously or by a glutathione-S-transferase catalyzed reaction. They may also hydrolyze 

spontaneously to tetrols (Hall and Grover 1988). 

The route by which PAHs and other xenobiotics enter the body may determine their fate and organ 

specificity. For example, an inhaled compound may bypass the first-pass effect of the liver and reach 

peripheral tissues in concentrations higher than one would see after oral exposures. Enzyme activities 

among tissues are variable. 

Benzo[a]pyrene was metabolized in vitro by human bronchial epithelial and lung tissue to the 

9,10-dihydrodiol, 7,8-dihydrodiol, and small quantities of the 4,5-dihydrodiol and 

3-hydroxybenzo[a]pyrene, all of which are extractable into ethyl acetate (Autrup et al. 1978; Cohen et 

PAHs 94 

2. HEALTH EFFECTS 

al. 1976; Kiefer et al. 1988). These metabolites also conjugated with glutathione and sulfates, but 

none conjugated with glucuronide. The rate of formation of the dihydrodiols was greater in the 

bronchial epithelium than in the lung (Autrup et al. 1978; Cohen et al. 1976). This may render some 

areas of the respiratory tract more sensitive to the effects of carcinogens. One principal difference 

seen in human lung was the generation of a major ethyl acetate-soluble metabolite that was identified 

as the sulfate conjugate of 3-hydroxybenzo[a]pyrene, benzo[a]pyrene-3-yl-hydrogen sulfate. This 

sulfate is very lipid soluble and, thus, would not be readily excreted in the urine (Cohen et al. 1976). 

Activation of benzo[a]pyrene has also been detected in human fetal esophageal cell culture (Chakradeo 

et al. 1993). 

Intratracheal instillation of benzo[a]pyrene to rats resulted in quinones constituting the highest 

concentration of metabolites in the lung and the liver within 5 minutes after instillation (Weyand and 

Bevan 1986, 1988). An in vitro study with rat lung demonstrated that the lung tissue has a high 

capacity to form quinones originating from oxidation at the six position of benzo[a]pyrene to form 

quinones and subsequently to water-soluble products. Ozone exposure resulted in an increase in the 

metabolism of benzo[a]pyrene metabolites with the greatest increase observed in the formation of 

metabolites generated by oxidation at the six position. The proposed retention of quinones following 

ozone exposure might lead to cytotoxicity associated with superoxide-anion generation by 

quinone-quinol redox-cycling. However, the high levels of benzo[a]pyrene used in this in vitro study 

may not relate to what occurs in vivo. Metabolism of benzo[a]pyrene at carbon six was higher at a 

lower dose than at the higher dose. Therefore, quinone production and detoxification may represent a 

major pathway of lung PAH detoxification in vivo (Basett et al. 1988). 

Approximately 50% of the benzo[a]pyrene that was intratracheally instilled in hamsters was 

metabolized in the nose (Dahl et al. 1985). The metabolite produced in the hamster nose included 

tetrols, the 4,5-, 7,8-, and 9,10-dihydrodiol, quinones, and 3-and 9-hydroxybenzo[a]pyrene. Similar 

metabolites were detected in nasal and lung tissues of rats inhaling benzo[a]pyrene (Wolff et al. 

1989b). The prevalence of quinone production was not seen in hamsters as it was in rats (Dahl et al. 

1985; Weyand and Bevan 1987a, 1988). In monkeys and dogs, dihydrodiols, phenols, quinones, and 

tetrols were identified in the nasal mucus following nasal instillation of benzo[a]pyrene 

(Petridou-Fischer et al. 1988). In vitro metabolism of benzo[a]pyrene in the ethmoid turbinates of 

dogs resulted in a prevalence of phenols (Bond et al. 1988). However, small quantities of .quinones 

and dihydrodiols were also identified. 
PAHs 95 

2. HEALTH EFFECTS 

Rat lung microsomes facilitated the dissociation of small amounts of benzo[a]pyrene from diesel 

particles, but only a small fraction of the amount dissociated was metabolized (Leung et al. 1988). 

The ability to dissociate benzo[a]pyrene was related to the lipid content of the microsomal fraction. 

Microsomes are able to enhance the slow dissociation of a small amount of benzo[a]pyrene from diesel 

particles in a form that can be metabolized. Free benzo[a]pyrene was principally and extensively 

metabolized to the 9,10-dihydrodiol. 

A human hepatoma cell line (HepG2) has high benzo[a]pyrene-metabolizing activity and converts 

benzo[a]pyrene to metabolites (Diamond et al. 1980). When [3H]-benzo[a]pyrene was added to the 

incubate, a large fraction of the radioactivity was not extractable into chloroform. The extractable 

fraction contained 9,10-dihydrodiols, 7,8-dihydrodiols, quinones, 3-hydroxybenzo[a]pyrene, and the 

unchanged parent compound. The cell lysate also consisted of the same metabolites, but the 

proportions of 3-hydroxybenzo[a]pyrene and the parent compound were much higher than in the 

medium. Conversely, the proportion of water-soluble metabolites in the cell lysate was lower than in 

the medium. Treatment of the medium and cell lysate with β-glucuronidase converted only 5-7% of 

the water-soluble metabolites to chloroform-extractable material. Aryl sulfatase had no effect on 

radioactivity. These results suggested that this human liver tumor cell line does not extensively utilize 

the phenol detoxification pathway (Diamond et al. 1980). 

Metabolism of benzo[a]pyrene in the primary culture of human hepatocytes primarily resulted in the 

formation of 3-hydroxybenzo[a]pyrene, 4,5-dihydrodiol, 9,10-dihydrodiol, and 7,8-dihydrodiol 

(Monteith et al. 1987). As the dose of benzo[a]pyrene increased, the amount of metabolites increased 

linearly. Binding to DNA was associated with the amount of unconjugated 7,8-dihydrodiol. DNA 

binding was linear up to a benzo[a]pyrene concentration of 100 μmol. At this concentration, binding 

increased 64-844 times over the extent of binding at 10 umol. As the concentration of benzo[a]pyrene 

increased, the ratio of dihydrodiol/phenolic metabolites also increased. Although the capacity to form 

dihydrodiols was not saturated at 100 umol benzo[a]pyrene, there was a change in the relative 

proportion of the dihydrodiol metabolites formed as the dose of benzo[a]pyrene increased. As 

benzo[a]pyrene concentration increased, the 9,10-dihydrodiol was the more prevalent metabolite, but 

levels of 7,8-dihydrodiol also increased (Monteith et al. 1987). 

Epoxide hydrolase is a microsomal enzyme that converts alkene and arene oxides to dihydrodiols. 

Appreciable enzyme activity was observed in human livers. Comparison of epoxide hydrolase 
PAHs 96 

2. HEALTH EFFECTS 

activities with various substrates revealed that the human liver has a single epoxide hydrolase with 

broad substrate specificity (Kapitulnik et al. 1977). Epoxide hydrolase activity is also present in other 

tissues and increases the likelihood for carcinogenic effects in these organs. Ethyl acetate extracts of 

human and rat bladder cultures contained 9,10-dihydrodiol, 7,8-dihydrodiol, and 

3-hydroxybenzo[a]pyrene. Covalent binding of [3H]-benzo[a]pyrene with DNA occurring in both 

human and rat bladder cultures suggested that benzo[a]pyrene-7,8-diol-9, 10-epoxide is generated. The 

urothelium of the bladder clearly has the ability to generate the ultimate carcinogen (Moore et al. 

1982). 

Hepatic microsomes from rats induced with 3-methylcholanthrene convert benzo[a]pyrene to 

benzo[a]pyrene-7,8-diol-9,10-epoxide (BPDE) 10 times faster than untreated microsomes. The 

rate-limiting step in BPDE formation is the competition for P-450 between benzo[a]pyrene and the 

7,8-dihydrodiol. Formation of BPDE is directly correlated with the 3-methylcholanthrene inducible 

form(s) of P-450 (Keller et al. 1987). Formation of the proximate carcinogen, 7,8-dihydrodiol, is 

stereoselective. Rabbit hepatic microsomes generated more of the 7R,8R enantiomer with an optical 

purity of >90% (Hall and Grover 1987). The major stereoisomer formed by rat liver microsomes is 

(+)-diol-epoxide-2 (R,S,S,R absolute conformation) (Jerina et al. 1976, 1980). This metabolite is 

highly tumorigenic (Levin et al. 1982) and gives rise to the major adduct formed upon reaction with 

DNA. The adduct is a diol epoxide-deoxyguanosine formed by alkylation at the exocyclic nitrogen 

(N-2) of deoxyguanosine. This diol epoxide-deoxyguanosine has been isolated from several animal 

species (Autrup and Seremet 1986; Horton et al. 1985) and human tissue preparations (Harris et al. 

1979). 

Studies using rat liver microsomes have shown that hydroxy metabolites of benzo[a]pyrene undergo 

glucuronidation (Mackenzie et al. 1993). Assays with three different DNA-expressed glucuronidases 

from human liver indicate preferential glucuronidation for the 2- and 5-hydroxy, 4- and 11-hydroxy, or 

l-, 2-, and 8-hydroxy derivatives of benzo[a]pyrene. There are differences in preferential activities for 

the glucuronidation of various benzo[a]pyrene metaboiites among the various DNA-expressed 

glucuronidases from human liver, with some glucuronidases being relatively or totally inactive toward 

this class of compounds (Jin et al. 1993). The results of this study suggest that the relative content of 

particular types of glucuronidases in a cell or tissue may be important for determining the extent to 

which a particular carcinogen is deactivated. 
PAHs 97 

2. HEALTH EFFECTS 

Several xenobiotics can induce enzymes to influence the rat liver microsomal metabolite profiles of 

various PAHs. For example, AHH, the cytochrome P-450 isoenzyme believed to be primarily 

responsible for the metabolism of benzo[a]pyrene and other PAHs, is subject to induction by PAHs. 

Treatment of pregnant and lactating rats with a single intraperitoneal dose of Aroclor 1245 increased 

the metabolism of benzo[a]pyrene by liver microsomes from pregnant and fetal rats 9-fold and 2-fold, 

respectively, and 2-fold in lactating rats (Borlakoglu et al. 1993). The pretreatment enhanced the 

formation of all metabolites, but the ratio of the 7,8-diol (the proximate carcinogen) was increased 

3-fold in lactating rats and 5-fold in pregnant rats. Similar results were observed in rabbit lung 

microsomes (Ueng and Alvares 1993). Cigarette smoke exposure has been shown to increase PAH 

metabolism in human placental tissue (Sanyal et al. 1993), and in rat liver microsomes (Kawamoto et 

al. 1993). In studying benz[a]anthracene metabolism, some xenobiotics were found to be weak or 

moderate inducers, but even less efficient ones altered the benz[a]anthracene profile significantly. 

Thiophenes equally enhanced oxidation at the 5,6- and the 8,9-positions. Benzacridines favored 

K-region oxidation (5,6-oxidation) (Jacob et al. 1983b). Indeno[ 1,2,3-c,d]pyrene stimulated the bay 

region oxidation (3,4-oxidation) of benz[a]anthracene (Jacob et al. 1985). Similar xenobiotic effects 

were observed with chrysene as a substrate (Jacob et al. 1987). While some enzyme activities are 

being enhanced, alternate enzymatic pathways may be suppressed (Jacob et al. 1983a). 

Rat liver microsomes also catalyzed benzo[a]pyrene metabolism in cumene hydroperoxide 

(CHP)-dependent reactions which ultimately produced 3-hydroxybenzo[a]pyrene and 

benzo[a]pyrene-quinones (Cavalieri et al. 1987). At low CHP concentrations, 

3-hydroxybenzo[a]pyrene was the major metabolite. As CHP concentrations increased, levels of 

quinones increased and levels of 3-hydroxybenzo[a]pyrene decreased. This effect of varying CHP 

levels was reversed by preincubating with pyrene. Pyrene inhibited quinone production and increased 

3-hydroxybenzo[a]pyrene production. Pretreatment with other PAHs like naphthalene, phenanthrene, 

and benz[a]anthracene nonspecifically inhibited the overall metabolism. The binding of 

benzo[a]pyrene to microsomal proteins correlated with quinone formation. This suggested that a 

reactive intermediate was a common precursor. The effects of pyrene on benzo[a]pyrene metabolism 

indicated that two distinct microsomal binding sites were responsible for the formation of 

3-hydroxybenzo[a]pyrene and benzo[a]pyrene-quinone (Cavalieri et al. 1987). 

Rat mammary epithelial cells (RMEC) have been shown to activate PAHs (Christou et al. 1987). 

Cytochrome-P-450 in RMEC is responsible for the monooxygenation of DMBA. Prior exposure’of 
PAHs 98 

2. HEALTH EFFECTS 

cultured cells to benz[a]anthracene induced DMBA metabolism. The metabolite profile following 

benz[a]anthracene-induction was significantly different from the profile obtained with purified P-450c, 

the predominant PAH-inducible enzyme in rat liver. The bay region 3,4-dihydrodiol, which was not 

formed with P-450c, was clearly detectable in RMEC. Low epoxide hydrolase activity in the 

benz[a]anthracene-induced RMEC limited the formation of all other DMBA dihydrodiols. The DMBA 

monooxygenase activity of benz[a]anfhracene-induced RMEC was inhibited by a-naphthaflavone. The 

study concluded that DMBA metabolism by RMEC depended on the induction of P-450c and at least 

one additional form of P-450 that is sensitive to α-naphthaflavone (Christou et al. 1987). 

As expected from results of other studies, the perfused rat lung can release high quantities of 

benzo[a]pyrene metabolites and conjugates into the perfusate (Molliere et al. 1987). Addition of a 

liver to this perfusion system up gradient from the lungs reduces the concentration of parent compound 

and free metabolites to less than 20% of that seen in the liver’s absence. The liver provides a 

protective effect on the lung to inhibit covalent binding of benzo[a]pyrene metabolites to pulmonary 

macromolecules. 

The effects of various factors that can modify the hepatic clearance of PAHs, specifically 

benz[a]anthracene and chrysene, were studied by Fiume et al. (1983). The hepatic clearance and rate 

constants of these PAHs were significantly reduced in the perfused livers of fasted rats relative to 

those of fed rats. This reduction was attributed to a decrease in aryl hydrocarbon hydroxylase activity. 

Fasting also accelerated the depletion of cytochrome P-450 and other microsomal enzymes. In 

contrast, pretreatment of the rats with these PAHs resulted in increased clearance of both hydrocarbons 

from the perfusion medium when compared to control rats. 

It was also noted by Fiume et al. (1983) that the livers of male rats demonstrated a significantly higher 

hepatic clearance of benz[a]anthracene than female rats, perhaps suggesting a sexual difference with 

aryl hydrocarbon hydroxylase activity. Similar findings regarding sexual differences in the metabolism 

of chrysene by rat livers were also reported by Jacob et al. (1985, 1987). Furthermore, Fiume et al. 

(1983) demonstrated that age can play a role in PAH metabolism. The hepatic clearance of PAHs in 

older rats (2 years) was significantly less than the hepatic clearance in younger rats (8 weeks). 

However, activation of benzo[a]pyrene to mutagenic derivatives, as measured by the Salmonella 

typhimurium test, with hepatic microsomes from male rats from 3 weeks to 18 months of age showed 

no age-dependent changes (Hilali et al. 1993). 
PAHs 99 

2. HEALTH EFFECTS 

A proposed metabolic scheme for the metabolism of the nonaltemant PAH, benzo[b]fluoranthene is 

presented in Figure 2-4. Nonalternant PAHs, in contrast to several alternant PAHs, do not appear to 

exert their genotoxic effect primarily through the metabolic formation of simple dihydrodiol epoxides. 

In the case of benzo[b]fluoranthene, there is evidence to suggest that metabolism to the dihydrodiol 

precursor to its bay-region dihydrodiol does occur. Rather than this metabolite being converted to its 

dihydrodiol epoxide; however, it appears to be extensively converted to its 5-hydroxy derivative. It is 

the further metabolism of this phenolic dihydrodiol to 5, 9, 10-trihydroxy- 11,12-epoxy-9,10,11,12-tetrahydrobenzo[b]fluoranthene that has been linked to the genotoxic activity of benzo[b]fluoranthene in 

mouse skin (Weyand et al. 1993b). In the case of benzo[j]fluoranthene, two potentially genotoxic 

metabolites have been identified. These are the trans-4,5- and 9,10-dihydrodiols of 

benzo[j]fluoranthene. It is the conversion of trans-4,5-dihydro-4,5-dihydroxybenzo[j]fluoranthene to 

anti-4,5-dihydroxy-5,6a-epoxy-4,5,6,6a-tetrahydrobenzo[j]fluoranthene that is principally associated 

with DNA adduct formation in mouse skin (LaVoie et al. 1993b; Weyand et al. 1993a). 

Benzo[k]fluoranthene in rat microsomes was shown to result in the formation of 8,9-dihydrodiol. This 

dihydrodiol can form a dihydrodiol epoxide that is not within a bay region. This may represent an 

activation pathway of benzo[k]fluoranthene that may be associated, in part, with its genotoxic activity. 

In the case of nonalternant PAHs, reactive metabolites, that deviate from classical bay region 

dihydrodiol epoxides, have been linked to their tumorigenic activity. 

2.3.4 Excretion 

2.3.4.1 Inhalation Exposure 

Urinary metabolites of PAHs were detected in workers exposed to these compounds in an aluminum 

plant (Becher and Bjorseth 1983). The ambient levels of PAHs in the workplace (indicated by the 

authors to exceed 95 mg/m3

, although these data were not presented) did not correspond to the amount 

of PAHs deposited, metabolized, and excreted in the urine in this study. No quantitative inhalation 

data were available in humans regarding the excretion of PAHs. 

Thirty-four workers in an electrode paste plant were monitored for response to exposure (Ovrebo et al. 

1994). Exposure to benzo[a]pyrene was 0.9 μg/m3; exposure to pyrene was 3.5 μg/m3. 

1-Hydroxypyrene was measured in the urine. Results from these workers were compared to two 

reference control groups, research and development workers and nickel refinery workers. 

PAHs 101 

2. HEALTH EFFECTS 

Measurements of PAH levels were collected by personal sampling and at two stationary sites. PAH 

aerosols were collected on filters during the mixing of hot coal tar with carbon. The value of PAHs 

on the filters varied from 4.3 to 84.6 μg/m3

, with a mean of 14.4 μg/m3. The PAH particulates were 

assayed for 8 of 11 PAHs classified as carcinogenic, including benz[a]anthracene, benz[g,h,i]perylene, 

benzo[a]pyrene, benzo[e]pyrene, benzo[b]fluoranthene, benzo[k]fluoranthene, benzo[j]fluoranthene, 

chrysene, fluoranthene, and indeno[ 1,2,3-c,d]pyrene, and presented relative to the marker compound, 

pyrene. Content of these compounds ranged from 30 to 150% of the pyrene content on the filters, 

which was 1.6 μg/m3. For example, the benzo[a]pyrene level was 0.8 μg/m3. Urine analysis indicated 

a mean urinary value of 1-hydroxypyrene among electrode paste plant workers of 6.98 umol/mol 

creatinine, compared to 0.08 μmol/mol creatinine for the R&D workers, and 0.14 μmol/mol creatinine 

for the industrial worker group. Smokers had higher levels of 1-hydroxypyrene compared to 

non-smokers in all groups. The urinary 1-hydroxypyrene level in the electrode plant workers 

correlated inversely with age. No correlation was found between frequency of use of a protective 

mask and the urinary 1-hydroxypyrene concentration. 

In an ongoing comprehensive evaluation of biological markers, workers in or near an iron foundry 

with varying exposures to PAHs were analyzed for response to exposure (Santella et al. 1993). 

Exposure to benzo[a]pyrene, determined by personal monitors, was 2-60 ng/m3

. 1-Hydroxypyrene 

was measured in the urine. Cigarette smoking, but not age or charbroiled food, influenced the level of 

1-hydroxypyrene. When workers were classified into three categories (low, <.0005 mg/m3; medium, 

0.0005-0.0012 mg/m3

; high, >0.0012 mg/m3), mean 1-hydroxypyrene levels were 2.7, 1.8, and 

3.6 μmol/mol creatinine, respectively. There was a significant difference between the groups after 

controlling for smoking exposure, but there was no consistent trend. The authors indicate that this 

study evaluates biological markers of exposure at PAH levels that are very low, compared to other 

studies. 

Workers employed in a graphite electrode producing plant (n=16) and a coke oven (n=33) were 

compared to a control population of maintenance workers in a blast furnace (n=54) (Van Hummelen et 

al. 1993). The concentration of PAHs in the environment was measured by personal air samplers, the 

concentration of hydroxypyrene in urine was measured, and smoking habits were evaluated. The mean 

age of the workers was 40, and did not differ significantly between the three plants. The proportion of 

smokers was not different among the three groups. The mean exposure for workers in the graphite 

electrode producing plant was 11.33 μg/m3

 (0.011 mg/m3) and was correlated with a urinary 
 

PAHs 102 

2. HEALTH EFFECTS 

hydroxypyrene concentration of 3.18 μg/g creatinine prior to the shift and 6.25 μg/g creatinine after the 

shift. For the coke oven workers, airborne PAHs were measured at 23.7 μg/m3 but this was not 

illustrative of true exposure, since there were a few very high exposures in the sample: 90.10 μg/m3 

for 7 workers compared to 5.57 μg/m3 for 26 workers. In accordance with the predominantly low 

exposure, the urinary hydroxypyrene levels were 0.51 μg/g creatinine before the shift and 0.75 μg/g 

creatinine after the shift. 

The excretion of benzo[a]pyrene following low-level inhalation exposure is rapid and high in rats 

(Bevan and Weyand 1988; Weyand and Bevan 1986; Wolff et al. 1989c); however, elimination is low 

in dogs and monkeys (Petridou-Fischer et al. 1988). After nose-only inhalation of 4.8 mg/m3 

[

14C]-benzo[a]pyrene for a single exposure or daily for 4 weeks, excretion of radioactivity in the feces 

of Fischer-344 rats was approximately 96% of the administered concentration (Wolff et al. 1989c). 

The excretion half-lives in feces and urine were 22 and 28 hours, respectively. 

2.3.4.2 Oral Exposure 

Five volunteers (21-41 years of age) ingested specially prepared diets high in PAHs, specifically 

benzo[a]pyrene (from grilled beef), for 2-3 days for an intake of approximately 0.007-0.02 mg/day 

(Buckley and Lioy 1992). A 100-250-fold increase in ingested benzo[a]pyrene in the high-PAH diet 

corresponded to a 6-12-fold increase in the elimination of 1-hydroxypyrene. However, benzo[a]pyrene 

and its other metabolites (not specified) were not measured in excreta, which prevented determination 

of the total excretion (and an estimate of oral absorption). 

Male Sprague-Dawley rats received a low dose of [3

H]-benzo[a]pyrene (suspended in 10% 

ethanolsoybean oil) daily for 3, 5, or 7 days (Yamazaki and Kakiuchi 1989). Highest radioactivity was 

excreted 2-8 days after first exposure. Polar and phenol metabolites represented approximately 60% 

and 20% of the radioactivity detected in urine, respectively. The conjugated form accounted for 80% 

of these urinary metabolites. Only small amounts of unmetabolized benzo[a]pyrene were detected in 

the urine. Excretion into the feces was not studied. 

Male Wistar rats eliminated a large amount of a single gavage dose of 0.22 mg/kg chrysene by 2 days 

postexposure (Grimmer et al. 1988). The unchanged parent compound represented 0.17% and 13.09% 

in the urine and feces, respectively. The recovery of the dose in excreta was 74% of dose after 4 days 
 

PAHs 103 

2. HEALTH EFFECTS 

of which about 61% was represented by hydroxychrysene compounds. The metabolite pattern was 

similar for urine and feces. The major metabolites identified were 1- and 3-hydroxychrysene, with 

about 100 times higher amounts in the feces (33.1% and 17.87%, respectively, of administered dose) 

than urine (0.79% and 0.35%, respectively). Approximately 79% of a large chrysene dose 

(1,217-2,717 mg/kg in the diet) was eliminated in the feces of rats; however, levels in the urine were 

not measured (Chang 1943). As the dose of chrysene in the diet increased, the percentage of excreted 

hydrocarbon also increased. 

Male Wistar rats were given single oral doses of 2, 4, 6, 9, or 1.5 mg/kg of [14C]-pyrene (Withey et al. 

1991). Recovery of the dose in excreta was 82% for the two low-dose groups and 50-63% for the 

other groups 2 days postexposure. The urine and feces contained 34-45% and 21-50% of the dose, 

respectively, at 4 days postexposure. Recovery of radioactivity in the bile was approximately 10% of 

the dose after 6 hours. 

2.3.4.3 Dermal Exposure 

No studies were located regarding the excretion of PAHs in humans following dermal exposure to 

single PAHs. Excretion patterns of 1-hydroxypyrene in urine were studied in two psoriatic patients, 

each treated daily with coal tar pitch covering over 50% of their skin for 3 weeks (Hansen et al. 

1993). The coal tar contained 1.43 mg/g pyrene. After 1 week of treatment, the urinary concentration 

of 1-hydroxypyrene increased approximately 100 times. However, the concentration after 3 weeks of 

treatment was decreased to that observed before treatment was initiated. The authors speculate that the 

healing of the psoriatic lesions may have rendered the skin less permeable to the PAHs after 3 weeks 

of treatment and healing. 

The urinary excretion of 1-hydroxypyrene was evaluated in 65 automotive repair workers (automobile 

or diesel truck repair) whose skin was exposed to used mineral oils, and compared to values from 

non-professionally exposed control males on control diets (Granella and Clonfero 1993). Smoking 

exposure was equivalent in both groups. Pyrene concentrations were determined in the oily material 

taken from cloths used to clean various types of engines. Values of urinary 1-hydroxypyrene were 

determined on Friday at the end of the working week, and again on Monday morning prior to work. 

Smoking and PAH-rich diets were considered confounding factors. At both the beginning and the end 

of the working week, the values of urinary 1-hydroxypyrene were slightly higher in exposed subjects 
PAHs 104 

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(0.178 + 0.150 and 0.194 + 0.135 μmol creatinine on Monday and Friday, respectively) than in 

controls ((0.124 + 0.090 μmol/mol creatinine on Monday and Friday, respectively). The urinary 

1-hydroxypyrene values were higher in both smoking and non-smoking subjects than controls. The 

highest values were found in urinary samples of smokers exposed to used mineral oils 

(0.259 + 0.090 μmol/mol creatinine). In non-smoking workers, post-shift 1-hydroxypyrene values 

were 0.154 + 0.105 μmol/mol creatinine, compared to 0.083 + 0.042 μmol/mol creatinine for the 

non-smoking controls. In automobile repair workers, there was no significant difference in the levels 

of 1-hydroxypyrene at the beginning and end of the work week. Tobacco smoking had more influence 

on the levels of 1-hydroxypyrene than did occupational exposure in this group. The influence of 

PAHs in the diet was only detectable when the subjects returned to work after the weekend. No 

explanation was given for this finding. This data suggests that exposure to PAHs through dermal 

contact with used engine oil is low. 

The elimination of benzo[a]pyrene was rapid and high in mice and guinea pigs following low- and 

high-level dermal exposure (Ng et al. 1992; Sanders et al. 1986). The percentages of recovered 

radioactivity in urine and feces were 24.5%, 46.9%, and 25% for Swiss Webster mice dermally 

exposed to benzo[a]pyrene at 1.25, 12.5, and 125 μg/cm2

, respectively, for 7 days (Sanders et al. 

1986). The feces in the high-dose animals had 35%, 58%, and 80% of the total recovered 

radioactivity after 24 hours, 48 hours, and 7 days, respectively. The amount of radioactivity excreted 

in urine was about 10% of amount excreted in feces. A elimination half-life of about 30 hours was 

estimated for benzo[a]pyrene. The data are limited because the exposed area of skin was not 

reportedly covered or collars were not employed to prevent ingestion of test compound by the animal. 

In guinea pigs, 73% of the dose was excreted 7 days after low-level (0.28 mg) exposure to 

benzo[a]pyrene (Ng et al. 1992). 

The excretion of dermally absorbed phenanthrene and pyrene was rapid in guinea pigs (Ng et al. 1991, 

1992). The presence of [14C]-activity in the urine and feces of rats that received [14C]-anthracene 

applied to the skin provides evidence of its absorption (Yang et al. 1986). Six days after 

administration of 9.3 μg/cm2

, the amounts detected in the urine and feces were 29.1% and 21.9% of 

the dose, respectively. 
 

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2.3.4.4 Other Routes of Exposure 

No studies were located regarding excretion of PAHs in humans following other routes of exposure. 

Three female Beagle dogs were given a bolus of aerosolized crystals of 7.7 mg/kg benzo[a]pyrene or 

2.8 mg/kg phenanthrene in a single breath by intratracheal instillation (Gerde et al. 1993a). The blood 

borne clearance of the PAHs was monitored by repeatedly sampling blood through catheters in the 

ascending aorta and the right atrium of the dog. Half of the benzo[a]pyrene cleared within 

2.4 minutes. Half of the phenanthrene cleared in 1.0 minute. Compared to clearance of phenanthrene, 

a less lipophilic PAH, the data indicate that the clearance of benzo[a]pyrene was limited by diffusion 

through the alveolar septa, while clearance of the moderately lipophilic phenanthrene was limited 

mostly by the rate of perfusion of the blood. The results indicate that inhaled PAHs of sufficient 

lipophilicity to limit diffusion through cells have a greater potential for toxicity to the lung than less 

lipophilic PAHs. Because of the thicker epithelia, bronchi should be at greater risk than the alveoli for 

PAH-induced toxicity exerted at the port of entry. Clearance of PAHs from the respiratory tract 

follows a biphasic pattern, with a rapid clearance of most of the PAH followed by a slow clearance of 

a small fraction. Previously published models predict that the rapid phase represents clearance through 

the thin epithelial barriers in the alveoli, the slow clearance is through the thicker epithelium of the 

airways, and the rate of clearance from either region will be slowed if the PAH has a high degree of 

lipophilicity. This study sought to validate model predictions for rates of alveolar clearance of PAHs 

of different lipophilicities. In a companion study, 3 female Beagle dogs were given doses of 

0.00091 g/kg benzo[a]pyrene instilled on the surface of the conducting airways (Gerde et al. 1993b). 

Sequential lavage of the mucous-retained materials followed the instillations. Benzo[a]pyrene was 

retained within the mucous lining layer sufficiently to be transported with the mucocilliary escalator. 

Fractions of benzo[a]pyrene penetrating to the bronchial epithelium had a clearance half-time in the 

range of 1.4 hours. This long retention indicates a diffusion-limited uptake of benzo[a]pyrene by the 

airways. Physiological models have predicted that the lipophilicity of solutes such as PAHs will delay 

clearance from the respiratory tract. This clearance consists of a delayed penetration of the mucous 

lining layer, allowing mucocilliary clearance, followed by a slow penetration of PAHs through walls of 

the conducting airways. 

Excretion of radioactivity in the urine of rats following intratracheal instillation only accounted for 

2.2% of the administered benzo[a]pyrene at 6 hours (Bevan and Weyand 1988; Weyand and Bevan 
PAHs 106 

2. HEALTH EFFECTS 

1986). Amounts of radiolabel in the blood were also very small. However, levels could still be 

determined in order to derive toxicokinetic parameters. The data obtained by Weyand and Bevan 

(1986) fit best to a three-compartment model whereby the half-lives for the three phases were 4.3, 

31.5, and 277 minutes. 

Results obtained by Weyand and Bevan (1986, 1988) revealed that a large fraction of the administered 

dose was excreted in the bile of rats, therefore suggesting that fecal elimination is the major excretion 

route. After 6 hours, 53% of the 0.001 mg/kg dose was excreted into the intestine and intestinal 

contents of rats that were without a bile duct cannula, while cannulated rats excreted 74% in the bile 

over the same period (Weyand and Bevan 1986). These results indicate the occurrence of 

enterohepatic recirculation. Metabolites excreted in the bile included thioether conjugates (62.5%), 

glucuronide conjugates (22.8%), sulfate conjugates (7.4%), and free benzo[a]pyrene (9.8%) (Weyand 

and Bevan 1988). Radioactivity detected in the gastrointestinal tract of rats following pulmonary 

exposure to benzo[a]pyrene also suggests that biliary excretion occurs (Weyand and Bevan 1987a; 

Wolff et al. 1989c). The percentage of benzo[a]pyrene excretion into bile declined as intratracheal 

doses increased from 0.00016 to 0.35 mg in rats and guinea pigs (Weyand and Bevan 1987b). 

However, the biliary excretion of benzo[a]pyrene in hamsters remained the same after administration 

of either dose. 

Female Wistar rats received low doses of chrysene (0.002 and 0.004 mg/kg) intratracheally (Grimmer 

et al. 1988). The major metabolite in the excreta was 1-hydroxychrysene. Hydroxychrysene 

compounds represented 31.26-48.9% of dose in the feces and about 3% in the urine. The 

unmetabolized parent compound was 17-19% of the administered dose in feces and only l-2% in 

urine. 

Less than 10% of instilled radioactivity was excreted in the urine and feces of dogs and monkeys 

48 hours after intratracheal administration of [3

H]-benzo[a]pyrene into the nostril (Petridou-Fischer et 

al. 1988). 

Nineteen outbred male rats were dosed intraperitoneally once with 200 mg/kg benzo[a]pyrene in 

sunflower oil (Likhachev et al. 1993). Concentrations of benzo[a]pyrene-7,8-diol, a marker metabolite 

of bioactivation of benzo[a]pyrene, and 3-hydroxy-benzo[a]pyrene, a marker metabolite of 

deactivation, were measured daily in the urine and feces for 15 days. Levels of these metabolites were 
PAHs 107 

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correlated with tumor latency. Another group of- 10 rats was dosed intraperitoneally once with 

15 mg/kg benzo[a]pyrene in sunflower oil and urine was collected for 3 days. Five rats were killed on 

day 3 and the other 5 were killed on day 8. Liver DNA concentrations of benzo[a]pyrene7,8,9,10-tetrols were determined in animals killed on day 8. Considerable individual variation was 

observed in the levels of daily and total excretion of benzo[a]pyrene-7,8-diol and 

3-hydroxy-benzo[a]pyrene in rats receiving 200 mg/kg benzo[a]pyrene. Both metabolites were 

excreted primarily in the feces. More than half of the total metabolites excreted were detected during 

the first five days, and peak concentrations were observed on the second day after benzo[a]pyrene 

administration. Peritoneal malignant fibrous histiocytomas developed in 10 of the 16 survivors at 

15 days. Levels of urinary benzo[a]pyrene-7,8-diol correlated positively with tumor latency. In the 

animals exposed to 15 mg/kg benzo[a]pyrene, a high correlation was found between excretion of 

benzo[a]pyrene-7,8-diol and benzo[a]pyrene-DNA adducts in the liver. 

[

14C]-Benzo[a]pyrene was administered to male germfree rats (Yang et al. 1994). Urine was collected 

24 hours before and every 24 hours for 7 days after administration. Urinary metabolites, consisting of 

9% of the administered radioactivity, were fractionated by lipophilic ion exchange chromatography, 

and characterized by reversed-phase HPLC, ultraviolet spectrometry, and gas chromatography/mass 

spectrometry. About 90% of the administered dose was excreted within 7 days; 80% in the feces and 

9% in the urine. About 90% of the radioactivity in the urine was recovered in the methanol eluate. In 

this eluate, more than 80% of the urinary metabolites were conjugated, while neutral metabolites 

constituted 13-18%. The neutral metabolites consisted of 7,8,9,10-tetrols (trace), 

trans-11,12-dihydrodiol (major), trans-7,8-dihydrodiol (trace), three isomer 

trihydroxy-benzo[a]pyrenes (major), carboxylic methyl ester derivatives of benzo[a]pyrene quinones, 

and trioxo-benzo[a]pyrenes (major). Most of the urinary radioactivity was excreted within 72 hours of 

dosing, with a peak excretion of 24-48 hours. A similar time course was observed for excretion in 

feces (data not shown). 

Six hours after intravenous administration of 0.08 mg/kg [14C]-benzo[a]pyrene to rabbits, 30% and 

12% of the dose was excreted in the bile and urine, respectively (Chipman et al. 1982). Excretion of 

activity into the bile was biphasic over a period of 30 hours with apparent half-lives of 0.27 and 

4.623 hours for the rapid and slow phases, respectively. Treatment of the bile and urine with 

β-glucuronidase and aryl sulfatase increased the amount of activity in the bile and urine that was 

extractable into ethyl acetate indicating the presence of glucuronide and sulfate conjugates. 
PAHs 108 

2. HEALTH EFFECTS 

Intraduodenal administration of the bile resulted in 21% and 14% of the intraduodenal dose being 

excreted into the bile and urine, respectively. Since biliary metabolites undergo enterohepatic 

recirculation, the half-life for 14C-activity is expected to be longer in animals without biliary fistulae 

(Chipman et al. 1982). 

The overall elimination of [3

H]-benzo[a]pyrene following intravenous administration (0.001 mg/kg) 

best fits a triexponential model, as after inhalation exposure (Weyand and Bevan 1986). The half-lives 

for the three phases were 1.5, 22.4, and 178 minutes. These parameters were very similar to those 

derived from intratracheal instillation. 

2.3.5 Mechanisms of Action 

PAHs are absorbed through the lungs by transport across the mucus layer lining the bronchi (Bevan 

and Ulman 1991). In general, PAHs are lipophilic compounds that can cross the lungs through 

passive diffusion and partitioning into lipids and water of cells (Gerde et al. 1991, 1993a, 1993b). The 

rapid, blood-bound redistribution of hydrocarbons at low blood concentrations from lungs to other 

organs indicates that diffusion is the rate-determining step (Gerde et al. 1991). The absorption rates 

vary among the PAHs, probably depending on the octanol/water partition coefficient. Essentially all 

of gastrically instilled benzo[a]pyrene is absorbed via uptake of fat-soluble compounds (Busbee et al. 

1990). Oral absorption of benzo[a]pyrene is enhanced by some oils (such as corn oil) in the 

gastrointestinal tract (Kawamura et al. 1988). The mechanism of dermal absorption of PAHs is most 

likely passive diffusion through the stratum comeum (Yang et al. 1986). 

PAHs and their metabolites are distributed to tissues by transport through the blood. Therefore, PAHs 

reach more-perfused tissues rapidly following exposure and are eliminated more slowly from lessperfused tissues (Bartosek et al. 1984). A large fraction of orally absorbed benzo[a]pyrene is believedto 

be transported by lipoproteins from the gastrointestinal tract to the blood via the thoracic duct 

lymph flow (Busbee et al. 1990). 

The carcinogenic mechanism of action of altemant PAHs is fairly well elucidated, but it is not as well 

described for nonaltemant PAHs. Furthermore, it is not known exactly how PAHs affect rapidly 

proliferating tissues. PAHs express their carcinogenic activity through biotransformation to chemically 

reactive intermediates that then covalently bind to cellular macromolecules (i.e., DNA) leading to 
PAHs 109 

2. HEALTH EFFECTS 

mutation and tumor initiation. The products of PAH metabolism include epoxide intermediates, 

dihydrodiols, phenols, quinones, and their various combinations. The bay region (e.g., the sterically 

hindered, cup-shaped area between carbons 10 and 11 of benzo[a]pyrene or 1 and 12 of 

benz[a]anthracene) diol epoxide intermediates of PAHs are considered to be the ultimate carcinogen 

for alternant PAHs (Jerina et al. 1980). These diol epoxides are easily converted into carbonium ions 

(carbocations) which are alkylating agents and thus mutagens and initiators of carcinogenesis. 

Therefore, the carcinogenic and toxic potential of PAHs relies on their metabolites. However, several 

of the tumorigenic PAHs (i.e., the nonaltemant PAHs) discussed in this profile do not have a bay 

region, or have been shown not to be similarly activated via a simple bay region epoxide (e.g., Amin 

et al. 1985a, 1985b). This observation has important implications regarding the expression of 

carcinogenicity for the nonalternant PAHs. If these chemicals are activated to carcinogens via a 

mechanism that differs from alternant PAHs, then they may also differ with respect to tumor site and 

species specificity. 

A prerequisite for conversion of PAHs into these active bay region diol epoxides is the presence of 

cytochrome P-450 and associated enzymes.responsible for this conversion. These enzymes can be 

found primarily in the liver, but they are also present in the lung, intestinal mucosa, and other tissues. 

Thus, factors such as distribution to the target tissue(s), solubility, and intracellular localization 

proximate to these enzymes figure prominently in the expression of a PAH’s carcinogenicity. In fact, 

in order to assess whether there was any correlation between carcinogenic potency and the ability to 

induce P-450 isoenzymes, several indices of P-450 isoenzyme activity (0-demethylation of 

ethoxyresorufin, metabolic activation of 2-amino-6-methyldipyrido [ 1,2-α:3’,2’d]imadazol [Glu-P-I] to 

mutagens, and immunological detection of polyclonal antibodies against purified rat P-450 I) were 

measured in microsomal preparations incubated with benzo[a]pyrene and benzo[e]pyrene (Ayrton et al. 

1990). While both PAHs increased several parameters of P-450-I activity, benzo[a]pyrene was 

markedly more potent than benzo[e]pyrene. Based on these results, the authors concluded that the 

carcinogenic potency of the PAHs tested could be predicted by the degree to which they induced these 

enzymes. 

Changes in the cytochrome P-450 system can affect the carcinogenicity of the PAHs. This system is 

susceptible to induction by the PAHs themselves as well as other chemicals commonly found in the 

environment. The degree and specificity (i.e., which enzymes are affected) of induction depend on the 

tissue and species and strain. The induction of one enzyme particularly important to the metabolism 
PAHs 110 

2. HEALTH EFFECTS 

of PAHs, AHH, is also known to be under genetic control (see discussions in Section 2.2.2 on 

responsive versus nonresponsive mouse strains). Given the heterogeneity of human genotypes, it is 

likely that certain human subpopulations exist that are more susceptible to AHH induction and thus 

more susceptible to the induction of cancer (see Section 2.7). 

Once the reactive bay region epoxide is formed, it may covalently bind to DNA and other cellular 

macromolecules and presumably initiate mutagenesis and carcinogenesis. Indeed, the level of 

DNA-adduct formation has been found to correlate with tumor induction activity for a number of 

PAHs in newborn rat liver and lung (Weyand and LaVoie 1988) and in mouse skin (Albert et al. 

1991b; Alexandrov and Rojas-Moreno 1990). Furthermore, no benzo[a]pyrene-DNA-adducts were 

found in rat skin, which is known to be resistant to PAH-induced skin tumor formation (Alexandrov 

and Rojas-Moreno 1990). The types of adducts formed in various tissues may dictate target organ 

susceptibility to PAH-induced carcinogenicity. Various metabolites of benzo[a]pyrene were 

administered to rats intraperitoneally and DNA adducts from lung, liver, and lymphocytes were 

measured (Ross et al. 1991). The only metabolites that led to DNA binding were 2-, 9-, and 

12-hydroxybenzo[a]pyrene and the truns-7,8-dihydrodiol of benzo[a]pyrene. The authors suggested 

that different DNA adducts resulting from the in vivo metabolism of benzo[a]pyrene in different tissues 

may be related to tissue specificity of benzo[a]pyrene-induced carcinogenicity. 

Although the bulk of this work on PAH-induced carcinogenicity has been done in animal models and 

animal in vitro systems, work in human in vitro systems indicates that these same mechanisms of 

activation may be involved in humans. For example, induction of AHH and formation of the reactive 

intermediate, benzo[a]pyrene 7,8-dihydrodiol, has been observed in the epithelial tissue from human 

hair follicles (Merk et al. 1987). All the steps necessary for cellular transformation and cancer 

induction were demonstrated in cultured human skin fibroblasts: inducible AHH activity, altered 

cellular proliferation kinetics, and DNA damage (Milo et al. 1978). Thus, humans are likely to be 

susceptible to tumor induction by PAHs by these mechanisms. 

Carcinogenic PAHs have been suggested to have an effect on immune function (Luster and Rosenthal 

1993; Saboori and Newcombe 1992), thereby allowing the induction of carcinogenesis, while 

noncarcinogenic PAHs do not affect immune function (see Section 2.4). The effects of dermally 

applied benzo[a]pyrene alone or following dermal pretreatment with the prostaglandin synthetase 

inhibitor, indomethacin, on contact hypersensitivity (cell-mediated immunity), production of antibodies 
PAHs 111 

2. HEALTH EFFECTS 

to DNP (humoral immunity), and the induction of skin tumors was studied in male BALBc mice 

treated for 6 weeks to 6 months (Andrews et al. 1991b). A group of mice treated with acetone served 

as controls. Skin tumors were observed in the mice treated with benzo[a]pyrene beginning at week 

18 of treatment. Pretreatment with indomethacin significantly increased (by 21%) the latency of tumor 

induction by benzo[a]pyrene and significantly reduced (by 46%) the weight of benzo[a]pyrene-induced 

skin tumors. Based on these findings, the authors suggested that benzo[a]pyrene-induced skin 

carcinogenesis may be mediated by a mechanism that involves prostaglandin suppression of cellular 

immunity. Undoubtedly, several other factors yet to be determined are involved in the ultimate 

expression of PAH-induced toxicity and carcinogenicity. 

2.4 RELEVANCE TO PUBLIC HEALTH 

PAHs occur ubiquitously in the environment from both synthetic and natural sources. PAHs occur in 

the atmosphere most commonly in the products of incomplete combustion. These products include 

fossil fuels; cigarette smoke; industrial processes (such as coke production and refinement of crude 

oil); and exhaust emissions from gasoline engines, oil-fired heating, and burnt coals. PAHs are present 

in groundwater, surface water, drinking water, waste water, and sludge. They are found in foods, 

particularly charbroiled, broiled, or pickled food items, and refined fats and oils. Individuals living in 

the vicinity of hazardous waste sites where PAHs have been detected at levels above background may 

experience exposure to these chemicals via inhalation of contaminated air or ingestion of contaminated 

food, soil, or water. 

Evidence exists to indicate that certain PAHs are carcinogenic in humans and animals. The evidence 

in humans comes primarily from occupational studies of workers who were exposed to mixtures 

containing PAHs as a result of their involvement in such processes as coke production, roofing, oil 

refining, or coal gasification (e.g., coal tar, coke oven emissions, soot, shale, and crude oil). Cancer 

associated with exposure to PAH-containing mixtures in humans occurs predominantly in the lungs 

and skin following inhalation and dermal exposure, respectively. Some ingestion of PAHs is probably 

due to swallowing of particulates containing PAHs subsequent to mucocilliary clearance from the lung. 

Certain PAHs have also been shown to induce cancer in animals. The site of tumor induction is 

influenced by route of administration: stomach tumors are observed following ingestion, lung tumors 

following inhalation, and skin tumors following dermal exposure, although tumors can form at other 

locations (e.g., lung tumors after dermal exposure). 
PAHs 112 

2. HEALTH EFFECTS 

Noncancer adverse health effects associated with PAH exposure have been observed in animals but 

generally not in humans (with the exception of adverse hematological and dermal effects). Animal 

studies demonstrate that PAHs tend to affect proliferating tissues such as bone marrow, lymphoid 

organs, gonads, and intestinal epithelium. 

Minimal Risk Levels (MRLs) for Polycyclic Aromatic Hydrocarbons 

Inhalation MRLS 

No inhalation MRLs have been derived for PAHs because no adequate dose-response data that identify 

threshold levels for noncancer health effects are available in humans or animals for any duration of 

exposure. 

Oral MRL.s 

No acute, or chronic oral MRLs were derived for PAHs because there are no adequate human or 

animal dose-response data available that identify threshold levels for appropriate noncancer health 

effects. Serious reproductive and developmental effects in animals associated with acute oral exposure 

to PAHs have been reported. These are not appropriate end points for the derivation of an MRL. 

Noncancer effects noted in longer term oral toxicity studies in animals include increased liver weight 

(generally not considered to be adverse) and aplastic anemia (a serious effect), neither of which is an 

appropriate end point for the derivation of an MRL. Intermediate-duration oral MRLs were derived 

for acenaphthene, anthracene, fluoranthene, and fluorene. 

Acenaphthene 

•	 An MRL of 0.6 mg/kg/day has been derived for intermediate-duration oral exposure 

(15-364 days) to acenaphthene. 

The MRL was based on a minimal LOAEL of 175 mg/kg/day for liver weight (EPA 1989c). Four 

groups of CD-l mice (20/sex/group) were gavaged daily with 0, 175, 350, or 700 mg/kg/day 

acenaphthene for 90 days. The toxicological evaluations of this study included body weight changes, 

food consumption, mortality, clinical pathological evaluations (including hematology and clinical 
PAHs 113 

2. HEALTH EFFECTS 

chemistry), organ weights and histopathological evaluations of target organs. The results of this study 

indicated no treatment-related effects on survival, clinical signs, body weight changes, total food 

intake, and ophthalmological alterations. Liver weight changes accompanied by microscopic 

alterations (cellular hypertrophy) were noted in both the mid- and high-dose groups, and seemed to be 

dose-dependent. Additionally, high-dose males and mid- and high-dose females showed significant 

increases in cholesterol levels. Increased relative liver weights in males, and increased absolute and 

relative liver weight in females, without accompanying microscopic alterations or increased cholesterol 

levels were also observed at the low dose; in light of the effects seen at higher doses, this change was 

considered to be a minimum LOAEL. There was no NOAEL. The MRL was obtained by dividing 

the LOAEL value by 300 (3 for a minimum LOAEL, 10 for extrapolation from animals to humans, 

and 10 for human variability) and rounding to one significant figure. 

MRLs for acute-duration and chronic-duration have not been derived because suitable NOAEL and 

LOAEL values have not been identified in the available literature. 

Fluoranthene. 

•	 An MRL of 0.4 mg/kg/day has been derived for intermediate-duration oral exposure 

(15-364 days) to fluoranthene. 

The MRL was based on a minimal LOAEL of 125 mg/kg/day for increased relative liver weight in 

male mice (EPA 1988e). Four groups of CD-1 mice (20/sex/group) were gavaged daily with 0, 125, 

250, or 500 mg/kg/day fluoranthene for 90 days. The toxicological evaluations of this study included 

body weight changes, food consumption, mortality, clinical pathological evaluations (including 

hematology and clinical chemistry), organ weights, and histopathological evaluations of target organs. 

The results of this study indicated no treatment-related effects on survival, clinical signs, body weight 

changes, total food intake, or ophthalmological alterations. All treated mice exhibited nephropathy, 

increased salivation, and increased liver enzyme levels in a dose-dependent manner. However, these 

effects were either not significant, not dose-related, or not considered adverse at 125 mg/kg/day. Mice 

exposed to 500 mg/kg/day had increased food consumption throughout the study. Mice exposed to 

250 mg/kg/day had statistically increased SGPT values and increased liver weight. Compound-related 

microscopic liver lesions (indicated by pigmentation) were observed in 65 and 87% of the mid- and 

high-dose mice, respectively. Male mice exposed to 125 mg/kg/day had increased relative liver 
PAHs 114 

2. HEALTH EFFECTS 

weight. The LOAEL is 125 mg/kg/day, based on-relative liver weight in males. There was no 

NOAEL. The MRL was obtained by dividing the LOAEL value by 300 (3 for a minimal LOAEL, 

10 for extrapolation from animals to humans, and 10 for human variability) and rounding to one 

significant figure. 

Fluorene. 

•	 An MRL of 0.4 mg/kg/day has been derived for intermediate-duration oral exposure 

(15-364 days) to fluorene. 

The MRL was based on a minimal LOAEL of 125 mg/kg/day for relative liver weight (EPA 1989e). 

Four groups of CD-l mice (20/sex/group) were gavaged daily with 0, 125, 250, or 500 mg/kg/day 

fluorene for 90 days. The toxicological evaluations of this study included body weight changes, food 

consumption, mortality, clinical pathological evaluations (including hematology and clinical chemistry), 

organ weights and histopathological evaluations of target organs. The results of this study indicated 

no treatment-related effects on survival, body weight changes, total food intake, or ophthalmological 

alterations. All treated male mice exhibited increased salivation, hypoactivity, and urine-wet 

abdomens. A significant decrease in red blood cell count and packed cell volume was observed in 

females treated with 250 mg/kg/day and in males and females at 500 me/kg/day. Decreased 

hemoglobin concentration was also observed in the high-dose group. A dose-related increase in 

relative liver weight was observed in all treated mice, and in absolute liver weight at >250 mg/kg/day. 

A significant increase in absolute and relative spleen and kidney weight was observed at 

250 mg/kg/day.. Increases in absolute and relative liver and spleen weights at the high dose were 

accompanied by histopathological increases in hemosiderin in the spleen and in the Kupffer cells of 

the liver. The LOAEL is 125 mg/kg/day based on increased relative liver weight. There was no 

NOAEL. The MRL was obtained by dividing the LOAEL value by 300 (3 for a minimal LOAEL, 

10 for extrapolation from animals to humans, and 10 for human variability) and rounding to one 

significant figure. 

Anthracene. 

•	 An MRL of 10 mg/kg/day has been derived for intermediate-duration oral exposure 

(15-364 days) to anthracene. 
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2. HEALTH EFFECTS 

The MRL was based on a NOAEL of 1,000 mg/kg/day for liver effects (EPA 1989d). The objective 

of this study was to evaluate the toxicity of anthracene in a subchronic toxicity study. Four groups of 

male and female CD-1 mice (20/sex/group) were placed on study, and were dosed with 0, 250, 500, 

and 1,000 mg/kg/day fluorene in corn oil by gavage for 13 weeks. The mice were observed twice 

daily for clinical signs. Body weights and food consumption were reported weekly. Hematologic and 

serum chemistry evaluations were completed at final sacrifice. At final sacrifice, gross post-mortem 

examinations were completed, organ weights were taken, and histological examinations were 

subsequently done on the tissues collected from all organ systems. No treatment-related finding were 

noted in survival, clinical signs, mean body weights, food consumption, and ophthalmological 

examinations, hematology, clinical chemistry, organ weights , gross pathology, and histopathology. In 

summary, anthracene produced no discemable effects. This study was conducted under the same 

laboratory conditions as the 90-day study of acenaphthene (EPA 1989c), and under similar laboratory 

conditions as the 90-day studies of fluoranthene (EPA 1988e) and fluorene (EPA 1989e), from which 

intermediate-duration MRLs were derived, based on liver effects. In these studies (EPA 1988e, 1989c, 

1989e), many other treatment-related and dose-related effects were observed, including renal, 

hematological, and splenic, that lent support to the derivation of the MRL for each compound. Thus, 

although no toxic effects were noted even at the highest dose tested in the study cited for anthracene 

(EPA 1989d), this free-standing NOAEL has considerable credibility, based on the assumption that 

toxic effects would have been observed if present, as was seen for the other compounds using the 

same study design. The NOAEL is 1,000 mg/kg/day based on the absence of liver effects, and any 

other effects in the organ systems studied. The NOAEL was the highest dose used in the study. The 

MRL was obtained by dividing the NOAEL value by 100 (10 for extrapolation from animals to 

humans, and 10 for human variability). 

Dermal MRLs 

No acute-, intermediate-, or chronic-duration MRLs were derived for the 17 PAHs because of the lack 

of appropriate methodology for the development of dermal MRLs. 

Death. There have been no reports of death in humans following exposure to any of the PAHs. 

However, benzo[a]pyrene is fatal to mice following ingestion, and aeath in animals has been reported 

following parenteral exposure to a number of PAHs. The intraperitoneal LD50 values in mice for 

pyrene, anthracene, and benzo[a]pyrene are 514, >430, and 232 mg/kg, respectively (Salamone 1981). 
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Reduced survival time in “Ah-responsive” mice (those capable of producing increased levels of 

cytochrome P-450 enzymes) was observed following a single intraperitoneal dose of 500 mg/kg 

benzo[a]pyrene (Robinson et al. 1975). In contrast, oral exposure to 120 mg/kg/day benzo[a]pyrene 

results in reduced survival of “Ah-nonresponsive” mice (those whose P-450 enzymes are not induced 

by PAHs). 

While the results in animal studies indicate that exposure to high doses of PAHs is lethal, the majority 

of the data are from parenteral exposure. This route is not applicable to exposure routes humans may 

expect to encounter, so the relevance of these findings to public health is not known. Parenteral 

administration bypasses the first-pass effect in the liver that occurs following oral exposure (PAHs 

may be expected to be ultimately biotransformed to inactive metabolites more quickly in the liver than 

in other tissues). However, because death has been observed in animals following oral exposure as 

well, it can be assumed that acute exposure to high enough doses of the PAHs can be lethal. 

Systemic Effects. 

Respiratory Effects. Adverse noncancer respiratory effects, including bloody vomit, breathing 

problems, chest pains, chest and throat irritation, and abnormalities in chest X-rays have been reported 

in humans exposed to PAHs and respirable particles in a rubber factory (Gupta et al. 1993). Inhalation 

is a significant route of exposure to PAHs in humans. In vitro studies using human lung tumor cells 

demonstrate that the benzo[a]pyrene-induced cytotoxicity (as measured by protein incorporation or 

cloning efficiency) observed was most likely due to formation of such reactive products as the 7,8­

diol 9,10-epoxide metabolite of benzo[a]pyrene (Kiefer et al. 1988). Thus, human lung cells are 

capable of metabolizing PAHs to reactive intermediates. This implies that inhalation exposure to 

PAHs could result in toxicity in the respiratory tract. 

Adverse effects on the respiratory tissue of laboratory animals have also been observed. The effects of 

benzo[e]pyrene, pyrene, anthracene, benz[a]anthracene, and benzo[a]pyrene on respiratory mucosa 

were studied in tracheal explants in rats (Topping et al. 1978). The PAHs were incorporated into 

beeswax pellets that were placed into tracheal grafts that had been transplanted subcutaneously into the 

subscapular region of isogenic host rats and the pellets remained in place for 4 weeks. Approximately 

50-60% of the test substance was delivered to the tracheal tissue by the end of 4 weeks, in most 

instances. Benzo[e]pyrene induced only mild changes that included slight hyperplasia of the tracheal 
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epithelium. A more long-lasting epithelial hyperplasia was observed with pyrene, anthracene, and 

benz[a]anthracene, and tracheas implanted with pyrene also exhibited a more severe mucocilliary 

hyperplasia. In addition, undifferentiated epithelium and small areas of squamous metaplasia were also 

seen with these PAHs, effects that persisted at least 8 weeks after exposure. Severe and long-lasting 

hyperplasia and transitional hyperplasia as well as metaplasia were seen in tracheas exposed to 

benzo[a]pyrene, and after 8 weeks, 75% of the epithelium was still abnormal. Acute inflammation 

(edema and/or granulocyte infiltration), subacute inflammation (mononuclear infiltration and an 

increase in fibroblasts), and fibrosis and hyalinization in the second half of the experiment were seen 

with all PAHs. The authors concluded that all of the PAHs tested induced pathological changes in the 

respiratory mucosa of the transplanted tracheas. The effects were different for the noncarcinogenic 

PAHs (benzo[e]pyrene, pyrene, anthracene, benz[a]anthracene), as compared to the carcinogenic PAH 

(benzo[a]pyrene); the former induced changes that were short-lived while the latter produced more 

severe, long-lasting (metaplastic) changes. 

Cultured fetal hamster tracheal explants were exposed to two concentrations each of benz[a]anthracene, 

benzo[b]fluoranthene, benzo[k]fluoranthene, benzo[e]pyrene, and pyrene in vitro for 4 days, and the 

effects of these PAHs on the respiratory epithelium were evaluated by scanning electron microscopy 

(Richter-Reichhelm and Althoff 1986). Exposure to benzo[e]pyrene and pyrene, as well as the lower 

concentrations of benz[a]anthracene and benzo[b]fluoranthene, resulted in effects similar to those seen 

in the DMSO controls: up to a 10% incidence of focally slight inhibition of epithelial differentiation 

and/or metaplasia. When the concentrations of all of the PAHs except benzo[e]pyrene and pyrene 

were doubled, the frequency of these lesions increased to 50-l00%, and the incidence of dysplasia 

(including hyperplasia) 1 was also observed to occur in a dose-related manner in explants exposed to 

benz[a]anthracene, benzo[b]fluoranthene, and benzo[k]fluoranthene. These lesions were not seen in 

the cultures exposed to benzo[e]pyrene, pyrene, or DMSO. The authors note that these effects on 

respiratory epithelium seen in vitro are similar to the preneoplastic changes seen in vivo following 

exposure to PAHs, and thus, this system may serve as a good screen for assessing risk to the 

respiratory tract. 

These observations, coupled with the fact that the respiratory system appears to be a target for 

PAH-induced cancer in humans, suggest that the respiratory system may be a target organ for 

PAH-induced noncancer adverse effects in humans as well. 
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Cardiovascular Effects. PAHs are contained in cigarette smoke, and smoking is a well-established 

risk factor in the development of atherosclerosis. Arterial smooth muscle cell proliferation, collagen 

synthesis, lipid accumulation, and cellular necrosis are all involved in the pathogenesis of the 

atherosclerotic plaque. In vitro studies conducted using bovine, rabbit, and human smooth muscle 

cells from arteries demonstrated that benzopyrene affects some of the aforementioned processes. Cell 

proliferation was not affected by benzo[a]pyrene, but a decrease in collagen secretion and an increase 

in cellular toxicity were noted in both the animal and human cell cultures (Stavenow and 

Pessah-Rasmussen 1988). 

Male White Leghorn chickens (six/group) were given weekly intramuscular injections of 

benzo[a]pyrene, benzo[e]pyrene, anthracene, and dibenz[a,h]anthracene for 16 weeks prior to removal 

of the abdominal aorta to investigate the effects of benzo[a]pyrene on the development of 

arteriosclerotic plaques (Penn and Snyder 1988). Animals injected with DMSO (the vehicle) served as 

controls. Microscopic plaques were found in the aortas of all treated and control animals. However, 

the plaque volume index (PVI), which is a measure of both plaque cross-sectional area and plaque 

length, was nine times larger in the benzo[a]pyrene animals than the controls. Benzo[e]pyrene and 

dibenz[a,h]anthracene also caused an increase in plaque volume as compared to controls. However, 

the plaque sizes in the animals treated with anthracene were no different than controls. Therefore, the 

authors concluded that benzo[a]pyrene, benzo[e]pyrene, and dibenz[a,h]anthracene “promoted” the 

development of preexisting atherosclerotic plaques in male chickens as opposed to initiating the 

development of new plaques. The ability to promote plaque development was not correlated with the 

mutagenicity or carcinogenicity of the PAH tested. Similarly, administration of benzo[a]pyrene or 

benzo[e]pyrene into atherosclerosis-susceptible or atherosclerosis-resistant pigeons for 3-5 months of 

treatment indicated that benzo[a]pyrene, but not benzo[e]pyrene, enhanced the formation of arterial 

lesions in female, but not male, birds (Hough et al. 1993). Female pigeons were also infertile, and 

showed ovarian abnormalities. 

These results, therefore, suggest that PAHs may contribute to the pathogenesis of atherosclerosis in 

humans. This is a particularly relevant health risk for those individuals who are exposed to high levels 

of PAHs in the environment and who also smoke cigarettes. 

Gastrointestinal Effects. Anthracene has been associated with gastrointestinal toxicity in humans. 

Humans that consumed laxatives that contained anthracene (anthracene concentration not specified) for 
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prolonged periods were found to have an increased incidence (73.4%) of melanosis of the colon and 

rectum as compared to those who did not consume anthracene-containing laxatives (26.6%) (Badiali et 

al. 1985). The authors suggested that the melanosis observed may be attributed to the consumption of 

anthracene laxatives and not to intestinal stasis. This study is severely limited because of confounding 

factors such as the existence of other predisposing factors for melanosis and lack of follow-up. 

Several PAHs discussed in this profile have been shown to alter enzyme activity in the intestinal 

mucosa of animals following oral administration, which could conceivably lead to increased production 

of reactive intermediates and tissue injury. Given the selectivity of PAHs for rapidly proliferating 

tissues such as gastrointestinal mucosa and the results discussed above, exposure to PAHs (particularly 

oral) by humans could lead to adverse gastrointestinal effects. 

Hematological Effects. Adverse hematological effects have been observed in animals following 

exposure to PAHs. For example, administration of a single intraperitoneal dose of benzo[a]pyrene to 

mice resulted in a small spleen, marked cellular depletion, prominent hemosiderosis, and follicles with 

large lymphocytes. These pathological lesions were associated with death (Shubik and Porta 1957). 

Death due to adverse hematological effects (e.g., aplastic anemia and pancytopenia resulting in 

hemorrhage) has also been observed in mice following intermediate-duration oral exposure to 

benzo[a]pyrene (Robinson et al. 1975). Fluoranthene and fluorene administered by gavage to male 

and female mice for 13 weeks caused hematological effects including decreased packed cell volume 

and decreased hemoglobin content (EPA 1988e, 1989e). In addition, it has been shown that 

benzo[a]pyrene is toxic to cultured bone marrow cells when applied directly (Legraverend et al. 1983) 

PAHs appear to affect other blood elements, as well. The influence of several PAHs on calcium 

ionophore-induced activation of isolated rabbit platelets was studied (Yamazaki et al. 1990). The 

activation of the platelets was assessed by measuring thromboxane B2 synthesis in response to 

stimulation by the calcium ionophore, A-23 187. The authors reported that thromboxane B2 synthesis 

was inhibited by incubation of the stimulated platelets with benz[a]anthracene, chrysene, 

benzo[a]pyrene, and benzo[g,h,i]perylene, and stimulated by incubation with anthracene and pyrene. 

However, no statistical analysis was performed on these data, and the changes reported are generally 

within ±10% of control values. In addition, the effects of the PAHs on thromboxane B2 synthesis are 

bidirectional, and in many instances, the same compound induced both inhibition and stimulation at 

different concentrations. 
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As discussed above, PAHs tend to exert their adverse effects on rapidly proliferating tissues, such as 

the bone marrow blood forming elements. It is likely that PAH-induced toxicity in this tissue is due 

to a specific attack on DNA of cells in the S or synthetic phase of mitosis (EPA 1988a). 

Although the human data available on PAH-induced hematological toxicity are flawed by confounding 

factors, they, together with the animal data and the propensity for PAHs to attack rapidly proliferating 

tissues, indicate that humans exposed to PAHs may be at risk for developing hematological toxicity. 

Hepatic Effects. No adverse hepatic effects have been reported in humans following exposure to 

PAHs. However, hepatic effects have been observed in animals following acute oral, intraperitoneal, 

or subcutaneous administration of various PAHs. These effects include the induction of preneoplastic 

hepatocytes, known as y-glutamyl transpeptidase foci, induction of carboxylesterase and aldehyde 

dehydrogenase activity, an increase in liver weight, and stimulation of hepatic regeneration (an 

indication of a proliferative effect) (Danz et al. 1991; Gershbein 1975; Kemena et al. 1988; Robinson 

et al. 1975; Shubik and Porta 1957; Torronen et al. 1981; Tsuda and Farber 1980). These hepatic 

changes are not considered serious adverse effects, but their incidence and severity have been shown to 

correlate with the carcinogenic potency of particular PAHs. Thus, monitoring of liver function and 

tissue integrity may prove useful in the evaluation of PAH exposure. 

More serious effects indicative of hepatic injury have been observed in animals. For example, an 

acute intraperitoneal injection of phenanthrene to rats resulted in liver congestion with a distinct 

lobular pattern, and an increase in serum aspartate aminotransferase, gamma-GT, and creatinine 

(Yoshikawa et al. 1987). Similarly, a single intraperitoneal injection of pyrene resulted in minimal 

swelling of the liver but no significant alterations in serum chemistry. Longer-term administration of 

PAHs has also been reported to result in adverse hepatic effects in animals. For example, increased 

absolute and relative liver weight correlated with hepatocellular hypertrophy was seen in male and 

female mice given 350 mg/kg/day acenaphthene by gavage for 13 weeks (EPA 1989c). Increased liver 

weight and dose-related centrilobular pigmentation accompanied by an increase in liver enzymes was 

observed in both male and female mice receiving 250 mg/kg/day fluoranthene by gavage for 13 weeks 

(EPA 1988e). 

Renal Effects. Adverse renal effects associated with PAHs have not been reported in humans. A 

single injection of anthracene or fluorene had no adverse effect on the kidneys of mice (Shubik and 
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Porta 1957). Dilated tubules were observed in the kidneys of mice administered pyrene in the diet for 

25 days (Rigdon and Giannukos 1964); the toxicological significance of this effect is not known. 

Renal tubular regeneration, and interstitial lymphocytic infiltrates and/or fibrosis were observed after 

13-week oral administration of fluoranthene to female mice at 250 mg/kg/day, and to male mice at 

500 mg/kg/day (EPA 1988e). Given the lack of renal toxicity in humans and the limited value of the 

observations made in animals, the risk to humans for renal toxicity following exposure to PAHs is not 

known. 

Endocrine Effects. There is suggestive evidence that PAHs may adversely affect endocrine function 

as well. The number of thymic glucocorticoid receptors in 6-week-old rats treated once with 2 mg/kg 

benzo[a]pyrene was measured (Csaba et al. 1991). It is assumed that the route of exposure was by 

oral gavage, but this was never explicitly stated. The number of these receptors was decreased by 

40% in females and was unaffected in males relative to the vehicle control animals. The statistical 

significance of these effects was not indicated, nor was the adversity of a decrease in receptor number 

assessed by examination of functional parameters. 

Dermal Effects. The skin is susceptible to PAH-induced toxicity in both humans and animals. 

Regressive verrucae were reported following intermediate-duration application of benzo[a]pyrene to 

human skin (Cottini and Mazzone 1939). Although reversible and apparently benign, these changes 

were thought to represent neoplastic proliferation. Benzo[a]pyrene application also apparently 

exacerbated skin lesions in patients with pre-existing skin conditions (pemphigus vulgaris and 

xeroderma pigmentosum) (Cottini and Mazzone 1939). Workers exposed to substances that contain 

PAHs (e.g., coal tar) experienced chronic dermatitis and hyperkeratosis (EPA 1988a). Coal tar 

preparations containing PAHs are used in the therapeutic treatment of some skin disorders. Adverse 

reactions have been noted in these patients, also. 

Adverse dermatological effects have also been noted in animals in conjunction with acute and 

intermediate-duration dermal exposure to PAHs. These effects include destruction of sebaceous 

glands, skin ulcerations, hyperplasia, and hyperkeratosis (Bock and Mund 1958), and alterations in 

epidermal cell growth (Albert et al. 1991b; Elgjo 1968). 

The observation that PAHs adversely affected the skin in both humans and animals is not surprising. 

The skin undergoes rapid cell turnover and is thus a likely target for PAH attack on DNA synthesis. 
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Given the information discussed above, the ubiquitous nature of PAHs in the environment, and the 

susceptibility of the skin to PAH-induced toxicity, adverse skin effects may occur in individuals 

exposed to these chemicals by the dermal route. 

Immunological and Lymphoreticular Effects. Humoral immunity was depressed in male 

iron foundry workers exposed to benzo[a]pyrene (Szczeklik et al. 1994). IgG, and IgA were depressed 

in those workers exposed to high levels. There are reports in the literature concerning the 

immunotoxicity of PAHs following dermal and parenteral exposure in animals. The carcinogenic 

PAHs as a group have an immunosuppressive effect. There are limited data that suggest that the 

degree of immunosuppression correlates with the carcinogenic potency. For example, using spleen cell 

cultures from C3H/Anf mice, suppression of humoral immunity (as measured by the plaque-forming 

cell [PFC] response to sheep red blood cells) and cell-mediated immunity (as measured by the 

one-way mixed lymphocyte response) were observed following incubation with 10-5-10-7 mol 

benzo[a]pyrene (Urso et al. 1986). There was no loss in cell viability at these concentrations. These 

immunological responses were unaffected by treatment with equivalent concentrations of 

benzo[e]pyrene. These findings led the authors to speculate that carcinogenic PAHs alter immune 

function, thereby allowing the induction of carcinogenesis while noncarcinogenic PAHs do not affect 

immune function. In addition, benzo[a]pyrene, but not benzo[e]pyrene, in the presence of S9 

metabolic activation mix, has been shown to inhibit interferon induction by viruses by 60-70% in 

cultured LLC-MK2 cells (Hahon and Booth 1986). 

Benzo[a]pyrene has been shown to markedly inhibit the immune system, especially T-cell dependent 

antibody production by lymphocytes exposed either in vivo or in vitro (Blanton et al. 1986; Lyte and 

Bick 1985; White and Holsapple 1984). These effects are generally seen at high dose relative to those 

that can induce cancer in animals. 

The effects of benzo[a]pyrene on several parameters of cell-mediated immune function in isolated and 

T-cell enriched mononuclear cell populations from three strains of mice (C57, C3H, and DBA) given a 

single intraperitoneal injection of 10-50 mg/kg benzo[a]pyrene or benzo[e]pyrene following 

stimulation with phytohemagglutinin (PHA) were studied (Wojdani and Alfred 1984). Neither 

benzo[a]pyrene nor benzo[e]pyrene had an inhibitory effect on lymphocyte blastogenesis induced by 

PHA; blastogenesis was slightly stimulated at 2.5 and 10 mg/kg of either PAH. Dose-related 

suppression of cell-mediated cytotoxicity of allosensitized lymphocytes was observed in all strains of 
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mice treated with benzo[a]pyrene, but no effect on this parameter was observed following treatment 

with benzo[e]pyrene. The percentage and adherence of macrophages from benzo[a]pyrene-treated mice 

were increased. The authors suggest that benzo[a]pyrene, but not benzo[e]pyrene, causes alterations in 

cell-mediated immune function that could compromise the animal’s immune function allowing the 

development of PAH-induced tumors. A major limitation of this study was the lack of statistical 

analysis, thereby making it difficult to determine the validity of the changes seen. As mentioned 

previously, relatively high doses of benzo[a]pyrene were employed in these studies. 

Benzo[a]pyrene-induced immune suppression was reported in male B6C3F1 mice (Lyte and Bick 

1985) and in the offspring of C3H/Anf mice treated intraperitoneally with benzo[a]pyrene (Urso and 

Gengozian 1980). Cell-mediated and humoral immune function of the liver, thymus, and spleen were 

evaluated in both maternal animals and the offspring of C3H mice administered one intraperitoneal 

dose of benzo[a]pyrene (150 mg/kg) during “mid-pregnancy” (Urso et al. 1992). The offspring were 

evaluated at 1 week and 18 months of age. Suppression of these various aspects of the immune 

system was observed in both the mothers and the offspring at these relatively high doses. However, 

the study lacked sufficient detail to adequately assess either the protocol or the results. 

Groups of four B6C3Fl female mice were administered single injections of 0, 50, or 200 mg/kg 

benzo[a]pyrene in corn oil to study the correlation between DNA adduct formation (as measured by 

32P-postlabelling analysis) and the suppression of polyclonal immune responses (3

H-TdR incorporation 

following stimulation by Escherichia coli lipopolysaccharide [LPS] and concanavalin A [Con A] and 

IgM secretion) and decreased cell viability in splenic lymphocytes harvested from the treated mice 

(Ginsberg et al. 1989). Spleen weight was significantly decreased (18%, p<0.05) at 50 mg/kg. The 

polyclonal response to LPS and Con A was suppressed by 30-45%, and this suppression was 

statistically significant at 200 mg/kg. IgM secretion was also significantly depressed (42%) at 

200 mg/kg. These immunosuppressive effects were accompanied by high levels of 

benzo[a]pyrene/DNA adducts. The authors speculated that the immunosuppressive effects of 

benzo[a]pyrene were due to a cytotoxic mechanism (as supported by in vitro experiments) that in turn 

resulted partially from the genotoxic effects of benzo[a]pyrene (i.e., the formation of 

benzo[a]pyrene/DNA adducts). 

Benzo[a]pyrene exerts its inhibitory effects on antibody production through alterations on the normal 

functioning of macrophages, T cells, and B cells (Blanton et al. 1988; Zhao et al. 1990). In contrast, 
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benzo[a]pyrene has no effect on most cellular immune responses before the appearance of tumors 

(Dean et al. 1983b), although benzo[a]pyrene exposure does inhibit IL-2-dependent proliferation 

(Myers et al. 1988). 

Benzo[a]pyrene may also induce autoimmune responses. Groups of eight female Sprague-Dawley rats 

were administered a single subcutaneous injection of 2 mg benzo[a]pyrene or benzo[e]pyrene 

(11.1 mg/kg) in sesame oil in the right thigh (Faiderbe et al. 1992). The animals were observed for up 

to 150 days and blood samples were taken at regular intervals to measure anti-phosphatidylinositol 

(PtdIns) antibodies. Serum levels of anti-PtdIns in animals treated with benzo[a]pyrene exceeded those 

of the oil-injected controls after day 10, and the difference became statistically significant (p<0.05) 

after day 40. The levels reached a peak at day 60 after which time they reached a plateau. The 

anti-PtdIns were of the IgG type and specific to phosphatidylinositol. Malignant sarcomas developed 

at the injection site in the animals treated with benzo[a]pyrene within 100-120 days. Serum levels of 

anti-PtdIns in animals treated with benzo[e]pyrene did not differ from those of the oil-injected 

controls. No malignant sarcomas developed at the injection site in 100% of the animals administered 

benzo[e]pyrene within 100-120 days. The authors speculated that constant stimulation of lymphocytes 

reactive for PtdIns by an endogenous antigen, of which PtdIns could be a part, was responsible for the 

increased serum levels of anti-PtdIns. The authors suggested that PtdIns metabolism is altered in 

rapidly proliferating malignant cells (the neoplasia being stimulated by benzo[a]pyrene), resulting in 

the synthesis of the PtdIns-containing antigen. The lack of an autoimmune response to benzo[e]pyrene 

was due to the fact that benzo[e]pyrene was not carcinogenic; there was no neoplastic transformation 

occurring that could result in the production of PtdIns-containing antigens such as was seen with 

benzo[a]pyrene. Therefore, this study provides evidence that benzo[a]pyrene-induced neoplasia may 

cause an alteration in the metabolism of endogenous substances, resulting in the production of 

autoimmune antibodies to those substances. 

The immunotoxic effects of benzo[a]pyrene have been noted in vitro as well (e.g., Ladies et al. 1991), 

and these studies provide some insight into the mechanism of action of benzo[a]pyrene-induced 

immunological effects. Splenic lymphocytes from B6C3F1 mice were incubated with various 

concentrations of benzo[a]pyrene for either 2 hours or the entire culture period (Ginsberg et al. 1989). 

A dose- and duration-related decrease in splenic lymphocyte viability (as measured by 3

H-TdR 

incorporation) and immune response (as measured by IgM secretion) was observed in the absence of 

S9 activation. Addition of S9 enhanced this effect after acute-duration exposure. However, there was 
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very little formation of benzo[a]pyrene/DNA adducts at benzo[a]pyrene concentrations of l-200 μmol 

in the splenic lymphocytes; this lack of effect was accompanied by a very low level of benzo[a]pyrene 

metabolism to DNA-adducting metabolites. Benzo[a]pyrene/DNA adducts were measured in liver and 

lung; however, in the in vivo experiment. This led the authors to suggest that benzo[a]pyrene-induced 

immunotoxicity as expressed by splenic lymphocytes was the result of a cytotoxic effect that was 

mediated, in part, by a genotoxic mechanism involving the formation of benzo[a]pyrene/DNA adducts 

remote from the spleen and a direct cytotoxic effect not requiring activation of benzo[a]pyrene to the 

reactive intermediate. 

Incubation of human lymphocytes with 0.1-01.0 μg/mL benzo[a]pyrene resulted in a suppression of 

lymphokine-activated killer cell (LAK) activity against tumor targets after 3 and 7 days (Lindemann 

and Park 1989). LAK DNA synthesis was also inhibited after 3 or 7 days of incubation with 

benzo[a]pyrene. However, benzo[a]pyrene had no effect on LAK binding with tumor targets, and 

benzo[a]pyrene did not interfere with the cytotoxic effect of natural killer cells added to the incubation 

medium. Based on these results, the authors concluded that benzo[a]pyrene interferes with the 

development of the immunological defense killer cells. 

Benzo[a]pyrene has also been shown to affect immune responses to viral infection. Benzo[a]pyrene 

can reversibly inhibit the induction of viral interferon in 32 different mammalian cell lines but only in 

the presence of S9 metabolic activation (Hahon and Booth 1988). This inhibition must occur at an 

early level and not affect viral interferon interactions because the activity of exogenous interferon was 

unaffected. In addition, influenza virus multiplication was also inhibited by activated benzo[a]pyrene. 

Benzo[e]pyrene had no effect on interferon induction. The authors suggest that benzo[a]pyrene’s 

inhibition of interferon induction may be an early step in compromising the host’s immune function, 

thereby allowing the induction of carcinogenesis. 

There is evidence to suggest that PAHs may alter the levels of brain neurotransmitters, which in turn 

affects the function of the immune system. The levels of two catecholamines, dopamine and 

norepinephrine, were determined in discrete brain areas in mice in which fibrosarcomas had been 

induced following a single subcutaneous injection of benzo[a]pyrene (Dasgupta and Lahiri 1992). 

Both dopamine and norepinephrine levels were significantly decreased in some brain regions (e.g., the 

corpus striatum and the hypothalamus), and these decreases were evident in both early and late tumor 

development. The authors state that since immunological function is compromised during 
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carcinogenesis and certain alterations in brain catecholamines impair immune function, the mechanism 

by which carcinogens such as benzo[a]pyrene cause immunosuppression and subsequent carcinogenesis 

may be via depression of brain catecholamines. 

Very little information is available on the immunological effects of other PAHs. Mice treated with 

high doses of dibenz[a,h]anthracene exhibited a reduced serum antibody level in response to antigenic 

challenge by comparison to controls (Malmgren et al. 1952). The immunosuppressive effects of 

dibenz[a,h]anthracene were studied in AHH-inducible mice (C57BL/6) and AHH-noninducible mice 

(DBA/2N) by intraperitoneal and oral administration (Lubet et al. 1984). Immunosuppression occurred 

in both strains and was more pronounced in the C57BL/6 mice than in the DBA/2N mice, following 

intraperitoneal administration. However, the DBA/2N mice were more susceptible to 

immunosuppression following oral administration. These results suggest that PAHs are rapidly 

metabolized and excreted following oral administration in AHH-inducible mice, whereas in 

AHH-noninducible mice, the PAHs are absorbed and distributed to target organs. Based on these 

results, the authors concluded that AHH inducibility plays an important role in the immunosuppressive 

activity of PAHs. 

B-cell lymphopoiesis in mouse bone marrow has been shown to be inhibited by incubation with 

fluoranthene in vitro at concentrations of ≥5 μg/mL (25 μmol). This effect on B-cell precursors may 

be mediated in part by a stimulation of programmed cell death; as demonstrated by the increase in 

DNA fragmentation induced by fluoranthene 15-17 hours after addition to the incubation medium. 

Furthermore, fluoranthene-induced DNA fragmentation always preceded fluoranthene-induced B-cell 

precursor death. Another mechanism for fluoranthene-induced inhibition of B-cell lymphopoiesis may 

be alterations in cell growth rates (fluoranthene was shown to slow the rate of B-cell precursor growth 

at concentrations <5 μg/mL) and/or altered cell survival (Hinoshita et al. 1992). 

The lymphoid system, because of its rapidly proliferating tissues, is susceptible to PAH-induced 

toxicity. The mechanism of action for this effect is most likely inhibition of DNA synthesis. No 

adverse effects on this system associated with PAH exposure have been reported in humans, but 

several accounts of lymphoid toxicity in animals are available. A single intraperitoneal injection of 

benzo[a]pyrene to mice resulted in a small spleen with marked cellular depletion, prominent and 

edematous trabeculae, and large lymphocytes. These lesions resulted in death (Shubik and Porta 

1957). The Shubik and Porta (1957) study was severely limited by the following: the benzo[a]pyrene 
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was only partly in solution, only one dose was employed, there was a small sample size, the purity of 

benzo[a]pyrene was not specified, only one sex was tested, and the presence of benzo[a]pyrene in the 

peritoneal cavity indicates inadequate absorption. No other similar studies were found in the literature. 

Even though these effects have not been noted in humans, and the data in animals are contained in 

only one study, the rapidly proliferating nature of this tissue suggest that humans exposed to PAHs 

may be a risk for the development of adverse effects on the lymphoid system. 

Given the high potential for exposure to PAHs in the vicinity of hazardous waste sites, the evidence 

from animal studies, and the heterogeneity of human genotypes with regard to enzyme induction 

capabilities, it would be prudent to consider that PAHs may pose an immunotoxic risk to humans 

living in areas surrounding hazardous waste sites. 

Neurological Effects. No information is available on the short- or long-term neurotoxic effects of 

exposure to PAHs in humans and animals. Acute-, intermediate-, or chronic-duration studies 

conducted in animals do not indicate that any of the PAHs tested showed evidence of neurotoxicity, 

although these tests were not designed to detect subtle neurological changes. 

However, there is evidence to suggest that PAHs may alter the levels of brain neurotransmitters. The 

levels of two catecholamines, dopamine and norepinephrine, were determined in discreet brain areas in 

mice in which fibrosarcomas had been induced following a single subcutaneous injection of 

benzo[a]pyrene (Dasgupta and Lahiri 1992). The mice were divided into two groups: early tumor 

development and late tumor development 3-4 months after administration of the benzo[a]pyrene. Both 

dopamine and norepinephrine levels were significantly decreased in some brain regions (e.g., the 

corpus striatum and the hypothalamus), and these decreases were evident in both early and late tumor 

development. The authors state that because immunological function is compromised during 

carcinogenesis and because certain alterations in brain catecholamines impair immune function, the 

mechanism by which carcinogens such as benzo[a]pyrene cause immunosuppression and subsequent 

carcinogenesis may be via depression of brain catecholamines. 

Reproductive Effects. In both prospective and retrospective studies, a decrease in fecundity was 

observed in women who were exposed prenatally to cigarette smoke (i.e., their mothers smoked when 

pregnant) (Weinberg et al. 1989; Wilcox et al. 1989). This association was apparent even after 
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adjustment for age, frequency of intercourse, current smoking status, age at menarche, childhood 

exposure to cigarette smoking, educational level, reproductive history, body weight, and consumption 

of alcohol and caffeine. On the other hand, an increase in fecundity was observed in a retrospective 

study of women who were exposed to cigarette smoke in early childhood. These apparently opposite 

effects may be partially explained by the fact that when a woman smokes during pregnancy, her fetus 

is exposed to the components of the cigarette smoke that cross the placenta as well as changes in fetal 

and placental oxygenation and metabolism that are secondary to changes in maternal metabolism 

resulting from smoking. However, childhood exposure involves direct inhalation of cigarette smoke. 

The authors could offer no explanation as to why fecundity should be increased as a result of 

childhood exposure to cigarette smoke (Weinberg et al. 1989; Wilcox et al. 1989). 

The testes and ovaries contain rapidly proliferating cells and therefore should be considered susceptible 

to damage by PAHs. The reproductive toxicity data in animals for the PAHs are limited. The 

available animal studies exclusively discuss the reproductive effects of benzo[a]pyrene. Adverse 

effects such as decreased fertility and total sterility in F1 progeny of CD-1 mice (Mackenzie and 

Angevine 1981) and decreased incidence of pregnant female rats at parturition (Rigdon and Rennels 

1964) were reported following oral exposure to benzo[a]pyrene. However, no adverse reproductive 

effects were observed in Swiss mice fed benzo[a]pyrene in their diet (Rigdon and Neal 1965). The 

metabolic differences and method of benzo[a]pyrene administration could account for the differential 

response to benzo[a]pyrene induced toxicity in these studies. A single intraperitoneal injection of 

benzo[a]pyrene to female C57BL/6N mice decreased the number of corpora lutea (Swartz and 

Mattison 1985). The, antiestrogenic effects causing decreased uterine weights in pseudopregnant 

Sprague-Dawley rats were reported following daily subcutaneous injections of benzo[a]pyrene during 

days 6-11 of pseudopregnancy (Bui et al. 1986). Similar treatment to pregnant rats during gestation 

caused resorptions, reduced percentage of viable litters, and decreased uterine weights (Bui et al. 1986; 

Cervello et al. 1992). Female mice exhibited decreased ovary weights after 13 weeks oral exposure to 

700 mg/kg/day acenaphthene (EPA 1989c). The studies conducted on the reproductive effects of 

benzo[a]pyrene via parenteral routes are briefly discussed below. 

Single intraperitoneal injection of benzo[a]pyrene to female C57BL/6N mice at doses as high as 

500 mg/kg body weight produced a dose- and time-dependent decrease in the number of corpora lutea 

(Swartz and Mattison 1985). The NOAEL in this study was 1 mg/kg/day. Groups of 20 C57BL/6N 

mice were given single intraperitoneal injections of 0-500 mg/kg benzo[a]pyrene (Miller et al. 1992) 
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and were killed at various intervals after injection. Total ovarian volume, total corpora lutea volume, 

and total number of corpora lutea per ovary were significantly reduced by doses of benzo[a]pyrene 

>5 mg/kg in a dose-related manner. These effects resolved in a dose- and time-dependent fashion, so 

that after 4 weeks, most changes were only evident in the animals treated with the 100 or 500 mg/kg 

benzo[a]pyrene. Individual corpora lutea volume actually increased in the treated animals, indicating 

that compensatory hypertrophy was probably occurring. The authors concluded that based on these 

findings and previous findings, benzo[a]pyrene impairs corpora lutea formation by destroying follicles. 

In another study, DBA/2N (D2), C57BL/6N (B6), and (DBA/2N x C57BL/6N)F1 (F1) mice 

(7-8/group) were injected with 10 μg of either benzo[a]pyrene or one of three different metabolites of 

benzo[a]pyrene ([+]7,8-oxide, [-]7,8-dihydrodiol, or [+]-diol-epoxide-2) into the right ovary (Mattison 

et al. 1989). The left ovary served as a control, and an additional control group injected with the 

vehicle (DMSO) also served as controls. Ovarian volume, wet weight, and small, growing, and large 

follicle number were measured in both the treated and contralateral control ovaries. Benzo[a]pyrene 

and one or more of its’metabolites caused decreases in the treated ovarian weight, the ovarian volume, 

and the small, growing, and large follicles in one or more strains. In most instances, the contralateral 

untreated ovary exhibited a compensatory response; ovarian weight and volume increased as compared 

to the DMSO controls. This study shows that benzo[a]pyrene and some of its metabolites are toxic to 

the ovaries of mice, and that the ovary is capable of metabolizing benzo[a]pyrene into reactive 

metabolites. Similarly, administration of benzo[a]pyrene or benzo[e]pyrene into atherosclerosissusceptible or atherosclerosis-resistant pigeons for 3-5 months of treatment indicated that 

benzo[a]pyrene, but not benzo[e]pyrene, rendered female pigeons infertile, with ovarian abnormalities 

(Hough et al. 1993). Cumulatively, these results demonstrate the sensitivity of integrated 

hypothalamic-pituitary-ovarian function to adverse effects of benzo[a]pyrene. 

Daily subcutaneous injection of benzo[a]pyrene beginning on day 6 of gestation for 6 days as opposed 

to 3 days significantly increased the number of resorptions, and decreased the fetal survival and uterine 

weights in Sprague-Dawley rats (Bui et al. 1986). In pseudopregnant (i.e., condition occurring 

following sterile matings in which anatomical and physiological changes occur similar to those of 

pregnancy) rats, similar benzo[a]pyrene treatment during days 6-l1 of pseudopregnancy significantly 

decreased the cyclic nucleotide levels and uterine weights suggesting an antiestrogenic effect (Bui et 

al. 1986). Use of a single dosage level precluded the assessment of dose response in these studies. 
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Pregnant Sprague-Dawley rats were administered subcutaneous injections of benzo[a]pyrene (in DMSO 

and corn oil) or the vehicle alone on gestation days 7, 9, 11, 13, and 15 (Cervello et al. 1992). The 

animals were sacrificed on gestation day 16. There were no maternal deaths or signs of maternal 

toxicity. However, the number of fetuses per litter and number of live fetuses per litter were 

significantly decreased in the animals treated with benzo[a]pyrene, and the number of resorptions was 

significantly increased. In addition, uterine weight, and whole uterine gravid weight were significantly 

decreased and increased, respectively. These results demonstrate the reproductive toxicity of 

benzo[a]pyrene, but a dose-response relationship could not be established because only one dose was 

tested. 

These results suggest that the potential for adverse reproductive effects may be increased in humans 

exposed to benzo[a]pyrene in the workplace or at hazardous waste sites. 

Developmental Effects. The developmental toxicity data for PAHs are mostly limited to in utero 

exposure of pregnant animals to benzo[a]pyrene via various routes of exposure. The placental transfer 

of benzo[a]pyrene has been shown in mice following oral and intravenous exposure of dams 

(Shendrikova and Aleksandrov 1974) and in rats after intratracheal administration (Srivastava et al. 

1986). The available data from oral studies in animals indicate that exposure of pregnant dams to 

benzo[a]pyrene produced resorptions and malformations in fetuses (Legraverend et al. 1984) and 

sterility in F1 mouse progeny (Mackenzie and Angevine 1981). Investigations by Legraverend et 

al. (1984) suggest that benzo[a]pyrene metabolites generated in the fetus rather than in the maternal 

tissues are responsible for these adverse effects. Also, the genetic differences observed in this study 

using the oral route were contrary to those induced by intraperitoneal administration of benzo[a]pyrene 

(Hoshino et al. 1981; Shum et al. 1979), thus emphasizing the importance of route of administration in 

benzo[a]pyrene metabolism and resulting toxicity. 

The developmental effects of benzo[a]pyrene have also been investigated in animals using the 

parenteral route of administration. Intraperitoneal injection of benzo[a]pyrene to pregnant mice 

produced stillbirths, resorptions and malformations at a greater incidence in Ah-responsive mice than 

in Ah-nonresponsive mice (Shum et al. 1979); testicular changes including atrophy of seminiferous 

tubules with lack of spermatids and spermatozoa; interstitial cell tumors (Payne 1958); 

immunosuppression (Urso and Gengozian 1980); and tumor induction (Bulay and Wattenberg 1971; 

Soyka 1980). Adverse effects observed following subcutaneous injection of benzo[a]pyrene include 
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increased fetal resorptio’ns in rats (Wolfe and Bryan 1939) and lung tumor induction in mice 

(Nikonova 1977). Decreased fetal survival (Wolfe and Bryan 1939) and lung tumor development 

(Rossi et al. 1983) were reported in Swiss mice following direct intra-embryonal injection of 

benzo[a]pyrene. 

Results of in vitro studies suggest that benzo[a]pyrene may affect a number of enzyme and hormone 

activities in the human placenta. The effects of benzo[a]pyrene on the binding of epidermal growth 

factor (EGF) and receptor autophosphorylation were studied in human placental cell cultures from 

early and late gestation placentas (Guyda et al. 1990). In a subsequent study, the effects of 

benzo[a]pyrene on the uptake of aminoisobutyric acid (AIB) by early and late gestational human 

placental cells was also studied (Guyda 1991). Benzo[a]pyrene decreased binding of EGF (37-60%) 

to the early gestation placental cells, but not the late gestation placental cells. The decrease in binding 

was due to a decrease in the number of high-affinity EGF binding sites. This effect was specific for 

EGF receptor sites and not due to a nonspecific effect of benzo[a]pyrene on the membranes because 

benzo[a]pyrene had no effect on the binding of 125I-labeled insulin and insulin-like growth factors. 

The authors concluded that the effects of benzo[a]pyrene on EGF binding were specific and related to 

gestational age. Benzo[a]pyrene stimulated AIB uptake by both early and late gestational cells and 

enhanced EGF-stimulated AIB uptake in spite of a decrease in the number of EGF receptors. The 

implications of these finding are that benzo[a]pyrene could alter EGF-induced secretion of human 

chorionic gonadotrophin and human placental lactogen secretion as well as metabolic functions, 

thereby affecting the regulation of cell growth and differentiation in human placentas. 

The activity of quinone reductase, a major protective enzyme, was increased 2-3-fold in first trimester 

human placental extracts in vitro when incubated for 6 hours with benz[a]anthracene, 

dibenz[a,h]anthracene, and chrysene at a concentration of 50 μmol (Avigdor et al. 1992). Based on 

these results, it can be postulated that the early placenta is capable of metabolizing certain toxic 

xenobiotics such as PAH quinone metabolites to inactive intermediates thereby protecting the 

developing embryo. 

Benzo[a]pyrene (50 μmol) has been shown to stimulate human gonadotropin release by first trimester 

human placental explants in vitro (Bamea and Shurtz-Swirski 1992). This stimulation was evident 

following static exposure for 24 hours and also in cultures that were superfused, meaning the 

benzo[a]pyrene had a delayed effect and did not need to be present for this effect to be expressed. 
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The implication of these findings is that benzo[a]pyrene can alter human placental endocrine function 

early in pregnancy. 

Results of animal and in vitro studies suggest that benzo[a]pyrene may produce adverse effects in the 

offspring of women exposed during pregnancy. Furthermore, the results of genetic studies conducted 

via oral and intraperitoneal routes emphasize the importance of route of administration in 

benzo[a]pyrene metabolism and resulting toxicity and the severity of the effect may vary depending 

upon the genotype of the individual exposed (see Section 2.7, Populations That Are Unusually 

Susceptible). Based on these observations, it is therefore prudent to consider that the genetically 

heterogeneous human population may show variation in response to in utero exposure to 

benzo[a]pyrene. 

Other PAHs such as anthracene, benzanthracene, chrysene, and dibenz[a,h]anthracene have also been 

tested for developmental effects via parenteral routes. Of these compounds, dibenz[a,h]anthracene 

produced fetolethal effects in rats (Wolfe and Bryan 1939), while chrysene produced liver tumors in 

the mouse progeny (Buening et al. 1979a; Grover et al. 1975) 

Genotoxic Effects. As the results presented in Tables 2-4 and 2-5 indicate, benzo[a]pyrene has 

been thoroughly studied in genetic toxicology test systems. It induces genetic damage in prokaryotes, 

eukaryotes, and mammalian cells in vitro and produces a wide range of genotoxic effects (gene 

mutations in somatic cells, chromosome damage in germinal and somatic cells, DNA adduct formation, 

UDS, sister chromatid exchange, and neoplastic cell transformation). In cultured human cells, 

benzo[a]pyrene binds to DNA and causes gene mutations, chromosome aberrations, sister chromatid 

exchange, and UDS. 

The results of in vivo studies indicate that many of the same types of adverse effects observed in vitro 

were seen in mice, rats, and hamsters exposed to benzo[a]pyrene via the oral, dermal, or 

intraperitoneal routes. The available data also indicate that benzo[a]pyrene is genotoxic in both 

somatic and germinal cells of intact animals (Table 2-4). The only study that was found regarding 

genotoxic effects in humans following exposure to benzo[a]pyrene reported no correlation between 

aluminum plant workers’ exposure to PAHs, including benzo[a]pyrene, and sister chromatid exchange 

frequency (Becher et al. 1984). The findings from assays using human cells as the target, in 
























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conjunction with the data from whole animal experiments, suggest that benzo[a]pyrene would probably 

have similar deleterious effects on human genetic material. 

Because the genotoxic activity of benzo[a]pyrene is well established, it is frequently used as a positive 

control to demonstrate the sensitivity of various test systems to detect the genotoxic action of unknown 

compounds. It also serves as the model compound for PAHs, and the available information on the 

formation of metabolites and structure of benzo[a]pyrene can theoretically be used to predict potential 

genotoxicity/carcinogenicity of other PAHs that have not been as extensively studied. 

Benzo[a]pyrene is generally considered to be biologically inert but can be metabolized by enzyme 

systems into at least 27 identified metabolites; however, only a few of these metabolites are reactive 

species that can damage DNA (De Bruin 1976). Benzo[a]pyrene 7,8-diol-9,10-epoxide is thought to 

be the ultimate mutagenic/carcinogenic metabolite. The primary metabolic pathway leading to the 

formation of the genotoxic/carcinogenic diol epoxides is assumed to be cytochrome P-450-dependent 

mixed-function oxidases (MFO), which in the case of PAHs are called AHHs. AHH is an ubiquitous 

enzyme system and has been found in a variety of tissues including liver, lung, and gastrointestinal 

tract of rats, mice, hamsters, and monkeys. AHH has also been detected in human liver, lung, 

placenta, lymphocytes, monocytes, and alveolar macrophages (Singer and Grunberger 1983). The 

evidence indicating that a variety of human tissues including human lymphocytes (GAO 1991; 

Wiencke et al. 1990), human lymphoblasts (Danheiser et al. 1989), and human mammary epithelial 

cells (Mane et al. 1990) can serve as a source of exogenous metabolic activation tends to support the 

role of AHH systems in initiating the conversion of benzo[a]pyrene to genotoxic forms. However, 

human erythrocytes, which do not contain an effective cytochrome P-450 system, were more efficient 

than induced rat liver fractions in converting benzo[a]pyrene to a genotoxin as indicated by higher 

sister chromatid exchange and micronuclei frequencies observed in human lymphocytes co-cultivated 

with human erythrocytes (Lo Jacono et al. 1992). The findings, while unconfirmed, suggest that 

enzymatic systems other than AHH may yield reactive intermediates. Similar evidence that uninduced 

lung, kidney, or spleen from Sprague-Dawley rats or BALB/c mice did not convert benzo[a]pyrene to 

a mutagen in S. typhimurium was reported by Ampy et al. (1988) who concluded that these tissues 

may, therefore, not be at risk from exposure. Superficially, the data from the studies conducted by 

Phillipson and Ioannides (1989) (indicating that neither benzo[a]pyrene nor benz[a]anthracene were 

mutagenic in S. typhimurium TA100 in the presence of noninduced hepatic fractions from rats, mice, 

pigs, or humans) would tend to support the data reported by Lo Jacono et al. (1992) and Ampy et al. 
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(1988) (suggesting either that other enzyme systems are involved in the bioactivation of these 

compounds or that certain tissues are not at risk). However, it is important to note, as pointed out by 

Phillipson and Ioannides (1989), that the level of P-450 isoenzyme proteins in unexposed animals is 

relatively low. Support for this statement was provided by the data demonstrating that uninduced 

hamster liver fractions, which contain high cytochrome P-448 levels, converted both benzo[a]pyrene 

and benz[a]anthracene to mutagens. By inference, it can reasonably be assumed that repeated 

exposures are required to induce the requisite enzyme systems to metabolize these promutagens to 

ultimate mutagenic/carcinogenic forms. It can further be assumed that the tissues of any species, 

including humans, that contain the appropriate inducible enzyme system are at risk. 

The function of other enzyme systems in the biotransformation of benzo[a]pyrene should not be ruled 

out. However, the evidence that cytochrome P-448 plays a major role in this process was further 

substantiated by the observation that rat liver enzymes induced by PAHs such as 3-MC or 

dibenz[a,h]anthracene were more efficient in metabolizing benzo[a]pyrene, dibenz[a,h]anthracene, and 

benz[a]anthracene to mutagenic metabolites for S. typhimurium than was phenobarbital (Teranishi et al. 

1975). This finding is consistent with the well-documented observation that various inducing agents 

such as phenobarbital and 3-MC cause the preferential synthesis of specific forms of cytochrome 

P-450. In the case of 3-MC, cytochrome P-448 is the principal form of induced cytochrome (Singer 

and Grunberger 1983). 

Epoxidation is thought to be the major pathway for benzo[a]pyrene metabolism pertinent to 

macromolecular interaction. The metabolic attack consists of the cytochrome P-450/P-448-dependent 

MFO system converting the benzo[a]pyrene molecule into an epoxide; the epoxide is acted upon by 

epoxide hydrolase to form a dihydrodiol, and a second cytochrome MFO reaction gives rise to the 

ultimate mutagenic/carcinogenic form, benzo[a]pyrene 7,8-diol-9,10-epoxide. One of the unique 

structural features of the diol epoxide is that it appears to form in the area of the PAH molecule 

referred to as the bay region (i.e., a deep-pocketed area formed when a single benzo ring is joined to 

the remainder of the multiple ring system to form a phenanthrene nucleus). The location of the bay 

region(s) for the various PAHs in this profile is depicted in Chapter 3 (Table 3-l). 

An additional feature of bay region diol epoxides is the ease of carbonium ion formation, which 

renders the PAH molecule highly reactive and susceptible to attack by nucleophiles (Jerina 1980; 

Singer and Grunberger 1983). The carbonium ion is more likely to form in structures where the 
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epoxide is part of the bay region of a saturated terminal angular ring than in an area where the diol 

epoxide is not associated with a bay region. Further enhancement of bay region epoxides can occur 

by the formation of an intramolecular hydrogen bond between the oxygen molecule of the epoxide and 

an associated hydroxyl group. These metabolites are also more resistant to enzymatic detoxification 

by epoxide hydrolase and glutathione transferase. The increased reactivity conferred by intramolecular 

hydrogen bonding and the decreased rate of further metabolism favor the interaction with DNA. 

Analysis of the bay region diol epoxides and their contribution to the DNA binding, genotoxicity, and 

carcinogenicity of various PAHs has provided the basis for the bay region hypothesis (Wood et al. 

1979a). For example, DNA adducts formed with non-bay region diol epoxides of benzo[a]pyrene have 

low mutagenic potential (MacLeod et al. 1994). The hypothesis further predicts that structures with 

more reactive bay regions would probably be more genotoxic and more carcinogenic. The body of 

evidence on the mutagenic and tumorigenic activity of the PAHs that form bay region diol epoxides 

(benzo[a]pyrene, benz[a]anthracene, chrysene, dibenz[a,h]anthracene; benzo[b]fluoranthene, 

benzo[j]fluoranthene, benzo[k]fluoranthene, and indeno[ 1,2,3-c,d]pyrene) supports this hypothesis. 

Based on these considerations, the available genetic toxicology results from studies conducted with the 

other PAHs in this profile are discussed relative to the bay region hypothesis. It is cautioned, 

however, that while the use of structural relationships to predict potentially reactive compounds is a 

powerful tool, it is not infallible, nor does it replace in vitro or in vivo testing. The formation of bay 

region epoxides is not an absolute requirement for carcinogenic activity because several PAHs that 

cannot form bay region epoxides are known to be carcinogens. It can, nevertheless, serve as a 

warning system to alert regulatory agencies to a potential health hazard and to enable investigators to 

establish priority lists for testing PAHs. 

There is no convincing evidence that the PAHs lacking a bay region structure (acenaphthene, 

acenaphthylene, and fluorene) are genotoxic; the results for acenaphthene and acenaphthylene are 

consistently negative. The induction of chromosome aberrations only, at a single dose in Chinese 

hamster lung cells exposed to fluorene (Matsuoka et al. 1991), is not sufficient to conclude that 

fluorene is a clastogen. However, none of these compounds have been extensively studied in in vitro 

assays, and they have not been tested in vivo. 
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The majority of the data for anthracene and pyrene were negative. Although isolated positive results 

were obtained, particularly in microbial systems, neither compound produced consistent genotoxic 

effects in mammalian cells in vitro, and both were negative in the limited in vivo studies that have 

been performed. 

The results obtained with fluoranthene, the remaining PAH without a bay region configuration, 

illustrate the need to apply the bay region hypothesis judiciously. There is ample evidence indicating 

that fluoranthene induced gene mutations in bacteria and human lymphoblasts and sister chromatid 

exchange in Chinese hamster ovary cells. Based on the evidence of a powerful response in the 

Escherichia coli PQ37 SOS DNA repair assay (SOS chromotest), Mersch-Sundermann et al. (1992a) 

predicted that fluoranthene has a relatively high probability of being genotoxic (80%). However, 

fluoranthene did not induce sister chromatid exchanges in mouse bone marrow cells (Palitti et al. 

1986). The work of Busby et al. (1984) in newborn mice, suggests that fluoranthene should be 

classified as a carcinogen. 

The occurrence of a bay region structure on the phenanthrene molecule suggests that this compound is 

genotoxic. However, the overall findings from the genetic toxicology studies do not support such a 

prediction. Similarly, the reported observation that the intraperitoneal injection of phenanthrene 

resulted in sister chromatid exchange induction in Chinese hamster bone marrow cells was not 

convincing (Bayer 1979; Roszinsky-Kocher et al. 1979). In both studies, the sister chromatid 

exchange increase over background was less than 1.5 fold and comparable doses did not cause 

chromosome aberrations. As stated earlier, the occurrence of a bay region on the molecule in 

conjunction with the reactivity of the bay region appear to be the determinants of 

genotoxic/carcinogenic activity. It is, therefore, probable that the bay region on phenanthrene is not 

very reactive, which would account for the lack of genotoxicity and for the low carcinogenicity index 

(<2) assigned to this compound (Arcos et al. 1968). Similarly, quantum mechanical calculations 

indicate a low probability of carbonation formation for the bay region diol epoxide of phenanthrene 

(Jerina 1980). 

The lack of genotoxicity for phenanthrene is thought to be related to the metabolism of this substance 

to its 9,10-dihydrodiol. However, specific methylated phenanthrenes, which direct the metabolic fate 

of this tricyclic hydrocarbon towards the formation of a classical bay region dihydrodiol epoxide, have 

exhibited significant genotoxicity (LaVoie et al. 1983a). It was demonstrated that the presence of a 
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methyl group at or adjacent to the K-region of phenanthrene can inhibit the formation of the 

9,10-dihydrodiol and produce powerful mutagens in Salmonella. Additionally, the presence of a 

halogen at the K-region site produced similar results, which further support the association between 

inhibition of 9,10-dihydrodiol formation and mutagenic potency of substituted phenanthrenes. The 

study authors concluded that derivatives of phenanthrene that can inhibit metabolism at this site have a 

greater probability of exerting genotoxic effects. Additionally, methylated derivatives of phenanthrene 

may act as tumor initiators, as shown on mouse skin (LaVoie et al. 1981b). 

The weight of evidence from the in vitro and in vivo studies conducted with benz[a]anthracene, 

dibenz[a,h]anthracene, and chrysene indicates that these three agents are genotoxic and that they exert 

their genotoxic effects through the binding of bay region diol epoxides to cellular DNA. Similarly, 

there is a substantive body of evidence that confirms the hypothesis implicating the formation of bay 

region diol epoxides as the major mechanism of action for both the genotoxicity and carcinogenicity 

induced by these PAHs (Cheung et al. 1993; Fuchs et al. 1993a, 1993b; Glatt et al. 1993; Lecoq et al. 

1989, 1991a; Wood 1979). It is also of note that these three compounds, as well as benzo[a]pyrene, 

induced neoplastic cell transformation in at least one cell line (see Table 2-5). 

Both benzo[b]fluoranthene and indeno[ 1,2,3-c,d]pyrene are known to exhibit mutagenic activity in 

S. typhimurium TA100 in the presence of rat liver homogenate (Amin et al. 1984; Hermann 1981; 

LaVoie et al. 1979; Rice et al. 1985b). Both of these agents were positive for in vitro cell 

transformation (Emura et al. 1980; Greb et al. 1980) and were shown to bind to mouse skin DNA in 

vivo (Hughes et al. 1993; Weyand 1989; Weyand et al. 1987). In addition, benzo[b]fluoranthene 

formed DNA adducts in the lungs and livers of adult rats (Ross et al. 1992) and newborn mice 

(Weyand et al. 1993b) and formed DNA adducts as well as induced sister chromatid exchange in 

peripheral blood lymphocytes of treated rats (Ross et al. 1992). These data are consistent with reports 

on the tumorigenic activity of these PAHs in rodents and their potential to act as carcinogens in 

humans. 

Studies on the mutagenic activity of benzofilfluoranthene and benzo[k]fluoranthene have also indicated 

that these nonalternant PAHs are mutagenic in S. typhimurium TAl00 (Amin et al. 1985b; LaVoie et 

al. 1979, 1980a; Weyand et al. 1988, 1992). There is also evidence that application of either of these 

PAHs to mouse skin results in DNA adduct formation (Hughes et al. 1993; LaVoie et al. 1991a, 

1991b; Weyand et al. 1987, 1993a). The relative extent of binding to mouse skin DNA (i.e., 
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benzo[j]fluoranthene > benzo[k]fluoranthene) parallels the relative tumorigenic potency of these 

hydrocarbons on mouse skin (i.e., benzo[j]fluoranthene is more potent than benzo[k]fluoranthene as a 

tumor initiator) (LaVoie et al. 1982). 

Benzo[g,h,i]perylene has been reported to be mutagenic in S. typhimurium and to cause DNA damage 

in E. coli. Benzo[g,h,i]perylene has been shown to be responsible for the formation of DNA adducts 

isolated after topical application of pharmaceutical grade coal tar to the skin of mice (Hughes et al. 

1993). However, the few studies that were found were insufficient to draw meaningful conclusions. 

The final compound in this group, benzo[e]pyrene, contains two equivalent bay regions. In vivo, 

benzo[e]pyrene induced a marginal increase in sister chromatid exchanges but did not cause structural 

chromosome aberrations in bone marrow cells harvested from Chinese hamsters receiving two daily 

intraperitoneal doses of 450 mg/kg (Roszinsky-Kocher et al. 1979). Similarly, sister chromatid 

exchange frequencies were not increased in V-79 cells co-cultivated with rat mammary epithelial cells 

as the source of exogenous metabolic activation (Mane et al. 1990). Benzo[e]pyrene was, however, 

reported to be weakly mutagenic in S. typhimurium (Andrews et al. 1978) and weakly genotoxic in the 

E. coli SOS chromotest (Mersch-Sundermann et al. 1992a, 1992b). 

The weak genotoxicity and the very weak carcinogenicity of benzo[e]pyrene appear to contradict the 

bay region diol epoxide hypothesis. Quantum mechanical analysis of the ringed structure strongly 

suggests the likelihood of carbonium ion formation and an associated chemical reactivity equivalent to 

the bay region diol epoxide of dibenz[a,h]anthracene (Wood et al. 1979a). Similarly, synthetically 

prepared bay region tetrahydro-epoxides of benzo[e]pyrene were found to be highly mutagenic in 

bacteria and mammalian cells, suggesting that bay region diol epoxide(s), if formed, would also be 

mutagenic. However, the parent compound was not metabolized to a reactive state by Aroclor 

1254 S9 or by purified cytochrome P-450 derived from rat livers induced with Aroclor 1254. From 

these results, Wood et al. (1979a) concluded that the lack of mutagenicity for benzo[e]pyrene may be 

associated with the failure of the cytochrome P-450-dependent monooxygenase system to catalyze the 

critical oxidations necessary to form the bay region diol. Specifically, there was very little formation 

of the bay region 9,10-dihydrodiol and low conversion of authentic 9,10-dihydrodiol to the bay region 

diol epoxide. Subsequent studies with authentic bay region diol epoxides of benzo[e]pyrene showed 

that they had relatively low mutagenic and tumorigenic activity as a result of the diaxial conformation 

of the diol. The diaxial conformation of the 9,10-dihydrodiol also provides an explanation for the low 
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formation of the bay region diol epoxide since metabolism is shifted away from the adjacent isolated 

double bond (Chang et al. 1981; Wood et al. 1980). Since structural-activity analysis suggests that the 

bay region diol epoxides would have biological activity, Wood et al. (1979a) caution that extrapolation 

of these findings to in vivo metabolic events in species other than rats should be approached with 

caution. It is conceivable that benzo[e]pyrene would be genotoxic in species capable of carrying out 

the appropriate enzymatic steps. 

In summary, several general conclusions can be reached for the unsubstituted PAHs evaluated in this 

profile. The formation of diol epoxides that covalently bind to DNA appears to be the primary 

mechanism of action for both genotoxicity and carcinogenicity of several of the unsubstituted PAHs 

that are genotoxins (benzo[a]pyrene, benz[a]anthracene, dibenz[a,h]anthracene, chrysene, 

benzo[b]fluoranthene, benzo[j]fluoranthene). There was insufficient evidence to draw meaningful 

conclusions regarding the genotoxic potential of benzo[g,h,i]perylene, although some evidence does 

exist. 

With regard to the unsubstituted PAHs that either lack a bay region configuration (acenaphthene, 

acenaphthylene, anthracene, fluorene, and pyrene) or appear to have a weakly reactive bay region 

(phenanthrene), there is no compelling evidence to suggest that they interact with or damage DNA. 

The five PAHs that appear to be exceptions to the bay region diol epoxide hypothesis are fluoranthene, 

benzo[k]fluoranthene, benzo[j]fluoranthene, and indeno[1,2,3-cd]pyrene (no bay region), and 

benzo[e]pyrene (two bay regions). The evidence does suggest, however, that fluoranthene possesses 

genotoxic properties while benzo[e]pyrene is either weakly mutagenic or nonmutagenic. 

Cancer. Evidence exists to indicate that mixtures of PAHs are carcinogenic in humans. The 

evidence in humans comes primarily from occupational studies of workers exposed to mixtures 

containing PAHs as a result of their involvement in such processes as coke production, roofing, oil 

refining, or coal gasification (e.g., coal tar, roofing tar, soot, coke oven emissions, soot, crude oil) 

(Hammond et al. 1976; Lloyd 1971; Maclure and MacMahon 1980; Mazumdar et al. 1975; Redmond 

et al. 1976; Wynder and Hoffmann 1967). PAHs, however, have not been clearly identified as the 

causative agent. Cancer associated with exposure to PAH-containing mixtures in humans occurs 

predominantly in the lung and skin following inhalation and dermal exposure, respectively. Some 

ingestion of PAHs is likely because of swallowing of particles containing PAHs subsequent to 

mucocilliary clearance of these particulates from the lung. 
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Certain PAHs are carcinogenic to animals by the oral route (e.g., benz[a]anthracene, benzo[a]pyrene, 

and dibenz[a,h]anthracene) (Berenblum and Haran 1955; Chu and Malmgren 1965; Klein 1963; 

McCormick et al. 1981; Neal and Rigdon 1967; Rigdon and Neal 1966; Snell and Stewart 1963; 

Spamins et al. 1986; Wattenberg and Leong 1970). The results of dermal studies indicate that 

benz[a]anthracene, benzo[a]pyrene, benzo[b]fluoranthene, benzo[j]fluoranthene, benzo[k]fluoranthene, 

chrysene, dibenz[a,h]anthracene, and indeno[l,2,3-c,d]pyrene are tumorigenic in mice following dermal 

exposure (Albert et al. 1991b; Cavalieri 1988b; Habs et al. 1984; Levin et al. 1984; Warshawsky and 

Barkley 1987; Wilson and Holland 1988; Wynder and Hoffmann 1959b). The sensitivity of mouse 

skin to PAH tumorigenesis forms the basis for the extensive studies performed using dermal 

administration. This tumorigenicity can be enhanced or modified with concomitant exposure to more 

than one PAH, long straight-chain hydrocarbons (i.e., dodecane), or similar organic compounds 

commonly found at hazardous waste sites. Thus, humans exposed to PAHs in combination with these 

substances could be at risk for developing skin cancer. 

For many of the carcinogenic PAHs discussed in this profile, it appears that the site of tumor induction 

is influenced by the route of administration and site of absorption, i.e., forestomach tumors are 

observed following ingestion, lung tumors following inhalation, and skin tumors following dermal 

exposure. However, the observations (discussed below) that (1) mammary tumors are induced 

following intravenous injection in Sprague-Dawley rats, (2) the susceptibility to tumor development on 

the skin after dermal application is not similar in rats and mice, and (3) oral cavity tumors are not 

observed when benzo[a]pyrene is administered in the diet, suggest that the point of first contact may 

not always be the site of PAH-induced tumors. The results of carcinogenicity studies conducted with 

the 17 PAHs discussed in this profile by parenteral routes of exposure are summarized in Table 2-6. 




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2.5 BIOMARKERS OF EXPOSURE AND EFFECT 

Biomarkers are broadly defined as indicators signaling events in biologic systems or samples. They 

have been classified as markers of exposure, markers of effect, and markers of susceptibility 

(NAS/NRC 1989). 

A biomarker of exposure is a xenobiotic substance or its metabolite(s), or the product of an interaction 

between a xenobiotic agent and some target molecule(s) or cell(s) that is measured within a 

compartment of an organism (NRC 1989). The preferred biomarkers of exposure are generally the 

substance itself or substance-specific metabolites in readily obtainable body fluid(s) or excreta. 

However, several factors can confound the use and interpretation of biomarkers of exposure. The 

body burden of a substance may be the result of exposures from more than one source. The substance 

being measured may be a metabolite of another xenobiotic substance (e.g., high urinary levels of 

phenol can result from exposure to several different aromatic compounds). Depending on the 

properties of the substance (e.g., biologic half-life) and environmental conditions (e.g., duration and 

route of exposure), the substance and all of its metabolites may have left the body by the time samples 

can be taken. It may be difficult to identify individuals exposed to hazardous substances that are 

commonly found in body tissues and fluids (e.g., essential mineral nutrients such as copper, zinc, and 

selenium). Biomarkers of exposure to PAHs are discussed in Section 2.5.1. 

Biomarkers of effect are defined as any measurable biochemical, physiologic, or other alteration within 

an organism that, depending on magnitude, can be recognized as an established or potential health 

impairment or disease (NAS/NRC 1989). This definition encompasses biochemical or cellular signals 

of tissue dysfunction (e.g., increased liver enzyme activity or pathologic changes in female genital 

epithelial cells), as well as physiologic signs of dysfunction such as increased blood pressure or 

decreased lung capacity. Note that these markers are not often substance specific. They also may not 

be directly adverse, but can indicate potential health impairment (e.g., DNA adducts). Biomarkers of 

effects caused by PAHs are discussed in Section 2.5.2. 

A biomarker of susceptibility is an indicator of an inherent or acquired limitation of an organism’s 

ability to respond to the challenge of exposure to a specific xenobiotic substance. It can be an 

intrinsic genetic or other characteristic or a preexisting disease that results in an increase in absorbed 
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dose, a decrease in the biologically effective dose;or a target tissue response. If biomarkers of 

susceptibility exist, they are discussed in Section 2.7, Populations That Are Unusually Susceptible. 

2.5.1 	 Biomarkers Used to Identify or Quantify Exposure to Polycyclic Aromatic 

Hydrocarbons 

PAHs and their metabolites can be measured in the urine of exposed individuals. In workers exposed 

to PAHs and dermatology patients treated with coal tar, the PAH metabolite 1-hydroxypyrene has been 

detected in the urine at concentrations of 0-40 μg/g creatinine or 290 ng/g creatinine, respectively 

(Jongeneelen et al. 1985). The amount of 1-hydroxypyrene detected in urine samples taken during the 

weekend was less than that detected during the weekdays, when the exposure was presumably higher 

than on the weekends. No correlation was found between occupational exposure levels and urine 

levels, so it is not known whether urinary metabolites could be detected following exposure to low 

levels of PAHs (as might be expected to occur in individuals living in the vicinity of hazardous waste 

sites). The presence of 1-hydroxypyrene in urine has also been demonstrated in workers exposed to 

PAHs in several different environments (creosote-impregnating plant, road workers laying asphalt, and 

workers exposed to diesel exhaust fumes) (Jongeneelen et al. 1988). In another study, the levels of 

urinary 1-hydroxypyrene significantly correlated with the environmental levels of pyrene and 

benzo[a]pyrene in coke plants, steel plants, and several Chinese cities where coal burning occurs (Zhao 

et al. 1990). The usefulness of monitoring urinary 1-hydroxypyrene concentration by liquid 

chromatography in occupationally exposed individuals as a biomarker for exposure to environmental 

PAHs was assessed. Postshift 1-hydroxypyrene urinary levels were significantly increased over 

preshift 1-hydroxypyrene levels in exposed workers as compared to nonexposed controls (the net mean 

change was 17-fold higher in the exposed workers as opposed to the nonexposed controls), and 

smoking status did not affect this result. In addition, in this work setting (an aluminum production 

plant), environmental levels of pyrene were strongly correlated with the environmental levels of total 

PAHs, indicating that pyrene is an appropriate environmental PAH marker in this situation. Thus, 

1-Hydroxypyrene levels in urine may be used as a biomarker of exposure to PAHs in certain situations 

(Tolos et al. 1990). Additional studies have evaluated the usefulness of determining PAH or 

metabolite levels in human urine as a measure of exposure in industrial and environmental exposure 

settings (Granella and Clonfero 1993; Hansen et al. 1993; Herikstad et al. 1993; Kanoh et al. 1993; 

Likhachev et al. 1993; Ovrebo et al. 1994; Santella et al. 1993; Strickland et al. 1994; Van Hummelen 

et al. 1993; Van Rooij et al. 1993a, 1993b; Viau et al. 1993). Based on these results, the identification 
PAHs 171 

2. HEALTH EFFECTS 

of PAH metabolites in the urine could serve as a method of biological monitoring of exposed workers, 

and possibly individuals living in the vicinity of hazardous waste sites where PAHs have been detected 

although it would be very difficult to distinguish exposures resulting from hazardous waste sites form 

those resulting from normal human activities. 

Autopsies performed on cancer-free patients found PAH levels of 1l-2,700 ppt in fat samples. 

Several PAHs were detected, including anthracene, pyrene, benzo[e]pyrene, benzo[k]fluoranthene, 

benzo[a]pyrene, and benzo[g,h,i]perylene, with pyrene being detected in the highest concentrations 

(Obana et al. 1981). A similar study done on livers from cancer-free patients found levels of 

6-500 ppt of all of the same PAHs except benzo[e]pyrene, which was not detected in the liver. As in 

the fat samples studies, pyrene appeared in the highest concentrations in the liver, but the overall 

levels were less than in fat (Obana et al. 1981). However, because of the ubiquitous nature of PAHs 

in the environment, detection of PAH metabolites in the body tissues or fluids is not specific for 

exposure to PAHs from hazardous waste sites. In addition, it is impossible to determine from these 

biological media whether exposure was to high or low levels of PAHs or if the exposure duration was 

acute, intermediate, or chronic. Benzo[a]pyrene and 1-nitropyrene were determined in excised lung 

samples from Chinese and Japanese cancer patients (Tokiwa et al. 1993). Exposure to the two marker 

compounds was from burning coal (Chinese, benzo[a]pyrene) or oil (Japanese, 1-nitropyrene). 

Compound levels in the lung correlated with individual exposure history. Using a large sample 

population, total PAH levels in lung tissue has also been shown to correlate to cancer incidence (Seto 

et al. 1993). 

PAHs form DNA adducts that can be measured in body tissues or blood in both humans and 

laboratory animals following exposure to PAHs or mixtures containing PAHs (e.g., Assennato et al. 

1993; Bjelogrlic et al. 1994; Chou et al. 1993; Culp and Beland 1994; Day et al. 1990; Fuchs et al. 

1993a, 1993b; Garg et al. 1993; Garner et al. 1988; Gallagher et al. 1993; Herberg et al. 1990; Hughes 

and Phillips 1990; Hughes et al. 1993; Jones et al. 1993; Khanduja and Majid 1993; Lee et al. 1993; 

Lewtas et al. 1993; Likhachev et al. 1993; Lu et al. 1993; Mass et al. 1993; Mumford et al. 1993; 

Newman et al. 1988, 1990; Nowak et al. 1992; Oueslati et al. 1992; Roggeband et al. 1994; Ross et al. 

1990, 1991; Shamshuddin and Gan 1988; Van Schooten et al. 1991, 1992; Weston et al. 1988, 1993a; 

Weyand et al. 1993a, 1993b). PAHs also form adducts with other cellular macromolecules, such as 

hemoglobin, globin, and other large serum proteins (e.g., Bechtold et al. 1991; Sherson et al. 1990; 

Weston et al. 1988). Again, these PAH-DNA and PAH-protein adducts are not specific for any 
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particular source of PAHs, and the adducts measured could have been from exposure to other sources 

of PAHs, such as complex mixtures that contain PAHs (e.g., crude oils, various high-boiling point 

distillates, complex petroleum products, coal tars, creosote, and the products of coal liquification 

processes), as well as vehicle exhausts, wild fires, agricultural burning, tobacco smoke, smoke from 

home heating of wood, cereals, grains, flour, bread, vegetables, fruits, meat, processed or pickled 

foods, beverages, and grilled meats. It is impossible to determine from these adducts whether 

exposure was to high or low levels of PAHs or if the exposure duration was short or long. 

In another study, an evaluation of mutations in peripheral lymphocytes was conducted in workers in or 

near an iron foundry; data were examined for correlations to benzo[a]pyrene exposure determined by 

personal monitors. Exposure levels for benzo[a]pyrene were 2-60 ng/m3

, which are the lowest levels 

yet analyzed in foundry workers. Mutations at the hypoxanthine guanine phosphoribosyl transferase 

(HPRT) and glycophorin A (GPA) loci, which are measures of molecular effects in lymphocytes and 

erythrocytes, respectively, were assessed to demonstrate their relationship to external exposure at these 

low levels. The rate of mutation was also compared to PAH-DNA adducts in the blood (Santella et al. 

1993). Workers were classified into three exposure categories, low (<5), medium (5-12), and high 

(>12). HPRT mutant frequencies for these groups were 1.04, 1.13, and 1 .82x10-6 cells, respectively, 

and demonstrated an upward trend that was marginally significant. In contrast, HPRT mutations were 

highly correlated with PAH-DNA adducts (Santella et al. 1993). GPA variants were not correlated 

with PAH exposure. These results support the use of both biomonitoring and personal environmental 

monitoring in the determination of exposure. 

Three methods were evaluated for their usefulness as biomarkers of exposure to benzo[a]pyrene in 

Wistar rats administered a single dose of l-200 mg/kg (Willems et al. 1991). These three methods 

were mutagenicity observed in urine and fecal extracts, chromosome aberrations and sister chromatid 

exchanges in peripheral blood lymphocytes, and DNA adduct formation in peripheral blood 

lymphocytes and liver. Mutagens were measured in urine and feces at levels of 10 and 1 mg/kg, 

respectively. DNA adduct formation (measured by 32P-postlabelling) could be detected at doses of 

≥10 mg in lymphocytes and ≥100 mg in liver, and the levels were twice as high in the lymphocytes as 

in the liver. Only a slight increase in sister chromatid exchanges and no enhanced frequency of 

chromosomal aberrations were seen. These results indicate that mutagenicity observed in excreta and 

DNA adducts in lymphocytes are both useful biomarkers of exposure in the rat, with mutagenic 

activity in feces being the more sensitive. 
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The ability of phenanthrene, fluoranthene, pyrene, benz[a]anthracene, chrysene, benzo[a]pyrene, 

benzo[b]fluoranthene, benzo[j]fluoranthene, indeno[l,2,3-c,d]pyrene, and dibenz[a]anthracene to bind 

to mouse hemoglobin and serum proteins after tail vein injection was investigated (Singh and Weyand 

1994). Urinary excretion of these compounds was also investigated. A direct correlation between 

urinary excretion and hydrocarbon molecular weight was observed. Binding to both globin and serum 

proteins was detectable, with binding to serum proteins 10-fold higher than to globin. These results 

provide an assessment of the potential usefulness of various PAHs as biomarkers of exposure to 

complex mixtures. 

2.5.2 	 Biomarkers Used to Characterize Effects Caused by Polycyclic Aromatic 

Hydrocarbons 

The available genotoxicity data indicate that several of the 17 PAHs discussed in this profile are 

genotoxic in both nonmammalian and mammalian systems and are indirect mutagens (i.e., requiring 

the presence of an exogenous mammalian metabolic system). There were no tests reported for humans 

exposed to benzo[a]pyrene (the most widely tested PAH) in vivo, but several types of cultured human 

tissue cells demonstrated positive results for benzo[a]pyrene-induced genotoxicity (as evidenced by the 

induction of chromosomal aberrations, sister chromatid exchange) and binding of benzo[a]pyrene to 

DNA. The measurement of DNA adduct formation as well as the induction of sister chromatid 

exchange in human lymphocytes has been proposed as a biomarker of benzo[a]pyrene-induced effects 

for human monitoring programs (Wiencke et al. 1990). It is probable, however, that the analysis of 

DNA adducts would be the more sensitive diagnostic tool since hundreds of benzo[a]pyrene-DNA 

adducts per nucleus would be required to yield a detectable increase in the sister chromatid exchange 

frequency for an exposed population. Although these results are exclusively from in vitro tests and the 

limited genotoxicity tests conducted on urine obtained from humans exposed to creosote (a complex 

mixture containing PAHs) have been negative, the genotoxic effects observed in human tissue cells, 

particularly DNA adduct formation, may serve as a biomarker of effects for at least one of the PAHs, 

benzo[a]pyrene. It would not be possible to identify the source.of the benzo[a]pyrene, however. 

PAHs have been shown to cause noncancer adverse effects on rapidly proliferating tissues such as the 

hematopoietic system, the lymphoid system, and the skin in both humans and animals. The skin is 

susceptible to PAH-induced toxicity in both humans and animals. Regressive verrucae were reported 

following intermediate-duration application of benzo[a]pyrene to human skin (Cottini and Mazzone 
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1939). Although reversible and apparently benign, these changes were thought to represent neoplastic 

proliferation. Benzo[a]pyrene application also apparently exacerbated skin lesions in patients with 

pre-existing skin conditions (pemphigus vulgaris and xeroderma pigmentosum) (Cottini and Mazzone 

1939). Workers exposed to substances that contain PAHs (e.g., coal tar) experienced chronic 

dermatitis and hyperkeratosis (EPA 1988a). However, none of these end points is specific to PAHs, 

and all can be seen with other agents. No other biomarkers of effect (specific or otherwise) have been 

identified following exposure to PAHs. 

For more information on biomarkers for renal and hepatic effects of chemicals see ATSDR/CDC 

Subcommittee Report on Biological Indicators of Organ Damage (1990) and for information on 

biomarkers for neurological effects see OTA (1990). Additional information can be found in a series 

of reports on biomarkers issued by the National Research Council (NRC 1989, 1992). 

2.6 INTERACTIONS WITH OTHER SUBSTANCES 

Because humans are usually exposed to PAHs in complex mixtures rather than to individual PAHs, it 

is important to understand the potential interactions between the PAHs and other components of the 

mixture. Interactions may occur among chemicals in a mixture prior to exposure, or may occur after 

exposure as a result of differing effects of the mixture components on the body. Synergistic and/or 

antagonistic interactions with regard to the development of health effects, particularly carcinogenesis, 

may occur. 

The extent of human exposures to PAH mixtures in occupational settings is generally not known in 

quantitative terms. However, exposures to complex chemical mixtures that include PAHs, such as use 

of tobacco products and exposure to roofing tar emissions, coke oven emissions, coal tar, and shale 

oils, have been associated with adverse health effects in humans. The biological consequences of 

human exposure to complex mixtures of PAHs depend on the interaction of the various strongly 

carcinogenic, weakly carcinogenic, or noncarcinogenic PAHs. For example, there is evidence to 

suggest that PAHs in cigarette smoke require other components in the smoke in order to exert their 

tumorigenic effect (Akin et al. 1976). 

The interaction between noncarcinogenic and carcinogenic PAHs has been extensively examined in 

animals. Noncarcinogenic PAHs exhibit co-carcinogenic potential and tumor-initiating and promoting 
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activity when applied with benzo[a]pyrene to the skin of mice. Simultaneous administration of weakly 

carcinogenic or noncarcinogenic PAHs including benzo[e]pyrene, benzo[g,h,i]perylene, fluoranthene, or 

pyrene significantly elevated the benzo[a]pyrene-induced tumor incidence. Benzo[e]pyrene, 

fluoranthene, and pyrene were more potent co-carcinogens than benzo[g,h,i]perylene (Van, Duuren and 

Goldschmidt 1976; Van Duuren et al. 1973). Benzo[e]pyrene, fluoranthene, and pyrene have also 

demonstrated weak tumor-promoting activity following initiation with benzo[a]pyrene, and these 

compounds increased benzo[a]pyrene-DNA adduct formation (Di Giovanni et al. 1982; Lau and Baird 

1992; Rice et al. 1984, 1988; Slaga et al. 1979; Smolarek et al. 1987; Van Duuren and Goldschmidt 

1976; Van Duuren et al. 1973). 

Interactions between selected noncarcinogenic PAHs and carcinogenic benzo[a]pyrene have also been 

documented to reduce the carcinogenic potential of benzo[a]pyrene in animals. Benzo[a]fluoranthene, 

benzo[k]fluoranthene, chrysene, perylene, and a mixture of anthracene, phenanthracene, and pyrene 

significantly inhibited benzo[a]pyrene-induced injection-site sarcomas. However, other PAHs 

including anthracene, benzo[g,h,i]perylene, fluorene, and indeno[ 1,2,3-c,d]pyrene had no antagonistic 

effects (Falk et al. 1964). Coexposure of tracheal explants to benzo[e]pyrene and benzo[a]pyrene 

resulted in an increased incidence of tracheal epithelial sarcomas over that seen with either PAH alone 

(Topping et al. 1981). Phenanthrene administration with benzo[a]pyrene decreased the DNA adduct 

formation in mice (Rice et al. 1984). 

There is evidence to suggest that benz[a]anthracene may serve as an anticarcinogen when administered 

with benzo[a]pyrene. Coadministration of benz[a]anthracene and benzo[a]pyrene decreased 

benzo[a]pyrene metabolism, benzo[a]pyrene-DNA adduct formation, and reduced the mutagenic 

activity of benzo[a]pyrene on hamster embryo cells. It has been postulated that the antimutagenic 

effect of benz[a]anthracene results from competition with benzo[a]pyrene for MFO enzymes, rather 

than the induction of detoxifying enzymes (Smolarek et al. 1986). 

The synergistic effect of individual PAHs on the mutagenicity of benzo[a]pyrene has also been 

demonstrated. Anthracene and benzo[e]pyrene enhanced the mutagenicity of benzo[a]pyrene, the 

maximal increase being obtained with anthracene. Benzo[e]pyrene (at a ratio of 2:l) had no effect on 

benzo[a]pyrene-induced mutation frequencies in V79 cells, but at a ratio of 15:1, benzo[e]pyrene 

inhibited the benzo[a]pyrene-induced mutations by approximately 10-fold. Benzo[e]pyrene inhibited 

the metabolism of benzo[a]pyrene by cultured hamster embryo only at high doses, but at both low and 
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high doses, the proportion of metabolites formed was altered by benzo[e]pyrene (Baird et al. 1984). 

The percentage of water-soluble metabolites was decreased, whereas the percentage of diols was 

increased. The authors postulated that benzo[e]pyrene alters the activity of other PAHs by inhibiting 

the conversion of the proximate carcinogenic diol of a particular PAH to a diol epoxide. 

Benzo[a]pyrene and dibenz[a,h]anthracene in combination with 10 noncarcinogenic PAHs were less 

potent tumor-inducers than was dibenz[a,h]anthracene alone or in combination with benzo[a]pyrene. 

The noncarcinogenic or weakly carcinogenic PAHs include benzo[e]pyrene, phenanthrene, anthracene, 

pyrene, fluoranthene, chrysene, perylene, benzo[g,h,i]pyrene, and coronene. Dose-response 

relationships for tumor incidences were observed for benzo[a]pyrene and dibenz[a,h]anthracene either 

alone or in combination with the 10 noncarcinogenic PAHs; however, no treatment-related sarcoma 

incidences were observed for any of the 10 noncarcinogenic PAHs (Pfeiffer 1977). 

Phenanthrene, a noncarcinogenic PAH, demonstrated a dose-related inhibition of 

dibenz[a,h]anthracene-induced carcinogenicity in mice. Phenanthrene significantly reduced the 

incidence of injection-site sarcomas elicited by dibenz[a,h]anthracene, especially at low doses. 

However, when triethylene glycol was the vehicle administered in combination with phenanthrene and 

dibenz[a,h]anthracene, a substantial increase (50%) in the rate of tumor induction was observed (Falk 

et al. 1964). 

Several experiments have shown that most PAH mixtures are considerably less potent than individual 

PAHs. Various combustion emissions and benzo[a]pyrene have been examined for carcinogenic 

potency and tumor initiation activity on mouse skin. In all cases, PAH mixtures were much less 

potent than benzo[a]pyrene. The authors calculated relative potency estimates that ranged from 

0.007 for coke oven emissions extract to less than 0.002 for diesel engine exhaust extract, using 

papillomas per mouse per milligram of the mixture as the end point (Slaga et al. 1980b). Another 

study demonstrated that the relative tumorigenicities, as compared to benzo[a]pyrene, of automobile 

exhaust .condensate (AEC), diesel emission condensate, and a representative mixture of carcinogenic 

PAHs were 0.0053, 0.00011, and 0.36, respectively, following chronic application to mouse skin 

(Misfeld 1980). AEC has also exhibited an antagonistic influence on benzo[a]pyrene carcinogenicity 

when subcutaneously administered to mice; this effect was particularly augmented at higher 

benzo[a]pyrene concentrations (Pott et al. 1977). 
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Carcinogenic and noncarcinogenic PAHs, comprising a quantitative fraction of automobile exhaust gas 

condensate, were selected for carcinogenicity testing via dermal exposure of female NMRI mice. The 

purpose was to identify interactions between mixtures of the carcinogenic and noncarcinogenic PAHs 

(Schmahl et al. 1977). The carcinogenic PAHs were benzo[a]pyrene, dibenz[a,h]anthracene, 

benz[a]anthracene, benzo[b]fluoranthene; and the noncarcinogenic PAHs were phenanthrene, 

anthracene, fluoranthene, pyrene, chrysene, benzo[e]pyrene, and benzo[g,h,i]perylene. Treatment was 

carried out twice a week, for the natural lifetime of the animals. Although the carcinogenic action 

observed could be attributed almost entirely to the action of the carcinogenic PAHs, in relatively small 

doses, addition of the noncarcinogenic PAHs did not inhibit carcinogenesis, but had an additive effect. 

Predicting the toxicity of a complex mixture on the basis of one of several of its components may be 

misleading, because the interactions among the components may modify toxicity. Since PAHs require 

metabolic activation by monoxygenases to elicit carcinogenic effects, any alteration in these metabolic 

pathways will influence the observed toxicity. There are two primary mechanisms by which chemicals 

interact with PAHs to influence toxicity. A compound may compete for the same metabolic activating 

enzymes and thereby reduce the toxicity of carcinogenic PAHs, or it may induce the metabolizing 

enzyme levels to result in a more rapid detoxification of the carcinogenic PAHs (Levin et al. 1982). 

Chaloupka et al. (1993) showed that a mixture of PAHs, produced as by-products from a manufactured 

gas plant, was 706 times more potent than expected, based on its benzo[a]pyrene content (0.17%) at 

inducing mouse hepatic microsomal ethoxyresorufin O-deethylase. Alternatively, compounds may 

compete for a deactivating pathway, thereby increasing the toxicity of PAHs (Furman et al. 1991). 

Many monooxygenase inducers are ubiquitous in the environment, and they may have an effect on the 

toxicity of PAHs. For example, environmental contaminants such as tetrachlorodibenzo-p-dioxin 

(TCDD), polychlorinated biphenyls (PCBs) and 3,3’,4,4’-tetrachlorobiphenyl (TCBP) can increase 

microsomal enzyme activity and consequently affect PAH toxicity (Jacob et al. 1987; Kouri et al. 

1978). The dermal absorption of benzo[a]pyrene was measured in the presence or absence of complex 

organic mixtures derived from coal liquefaction processes (Dankovic et al. 1989). The dermal half-life 

of benzo[a]pyrene was 3.0 hours when applied alone, 6.7 hours when measured as a component of a 

mixture, and ranged from 7.8 to 29.7 hours in the presence of different mixtures. The authors 

proposed that these mixtures inhibit the dermal absorption of benzo[a]pyrene by inhibiting the 

metabolism of benzo[a]pyrene at the application site. Interactions. can thus play important modulatory 

roles in the expression of PAH toxicity that may not be adequately reflected based on the toxicity of a 

single PAH. 
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The majority of human exposures to PAHs occur in the presence of particles that affect the 

pharmacokinetics of PAHs in a manner that can enhance their carcinogenicity. Coadministration of 

benzo[a]pyrene and particulate material, such as hematite (Fe2O3) and arsenic trioxide (As2O3), greatly 

increases respiratory tract tumor yields in laboratory animals following intratracheal instillation 

(Pershagen et al. 1984; Saffiotti et al. 1972; Stenback and Rowland 1979; Stenback et al. 1976). The 

effects of particles on the potential human carcinogenicity of PAHs are likely to be similar. When 

benzo[a]pyrene is particle-bound, clearance from hamster lungs is slower than that of pure 

benzo[a]pyrene aerosol, increasing the length of time the lungs are exposed and increasing the dose to 

the gastrointestinal tract as a result of mucocilliary clearance. Respirable benzo[a]pyrene-containing 

particulates such as diesel exhaust, when coated with the phospholipid component of a pulmonary 

surfactant, are genotoxic (Wallace et al. 1987). Dusts can increase the rates of pulmonary cell 

proliferation (Harris et al. 1971; Stenback and Rowland 1979; Stenback et al. 1976), which in turn 

increases the cells’ susceptibility to an initiation event in the presence of a carcinogen. 

Environmental exposure to PAHs can also occur along with exposure to other environmental 

pollutants. The effects of exposure to SO2 (either by inhalation or systemically with endogenous 

sulfite/bisulfite anions that accumulated as a result of induced sulfite oxidase deficiency) on 

benzo[a]pyrene-induced lung tumors were studied in male Sprague-Dawley rats (Gunnison et al. 1988). 

The animals, were administered benzo[a]pyrene (1 mg) by weekly intratracheal instillation for 

15 weeks during which time they were exposed daily by inhalation to 30 or 60 ppm SO2 or were 

maintained on a high tungsten to molybdenum diet. There were no statistically significant differences 

between the benzo[a]pyrene only and the benzo[a]pyrene + SO2 or benzo[a]pyrene + sulfite/bisulfite 

groups with respect to the incidence of squamous cell carcinomas of the lung, latency for tumor 

development, or rate of appearance. Although benzo[a]pyrene alone induced almost 100% tumor 

incidence leaving little room for an SO2-induced enhanced response, a shortened latency or increased 

rate of appearance would have suggested that SO2 potentiates the carcinogenicity of benzo[a]pyrene, 

and this did not occur. Therefore, the authors concluded that SO2 does not potentiate the 

carcinogenicity of benzo[a]pyrene in the lung. 

Concomitant exposure to solvents may also occur, particularly in an occupational setting. It has been 

demonstrated that pretreatment of rats with toluene (1 g/kg intraperitoneally) inhibits the total 

cytochrome P-450 content in microsomes isolated from the lungs (Furman et al. 1991). In addition, 

formation of 3-hydroxybenzo[a]pyrene (a nontoxic metabolite) was inhibited by 36% by the 
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microsomes in vitro whereas the formation of several diols (reactive intermediates) was unaffected by 

toluene pretreatment. These results indicate that toluene alters the balance between toxification and 

detoxification of benzo[a]pyrene by cytochrome P-450 in the lung, favoring the formation of reactive 

genotoxic/carcinogenic intermediates. Therefore, the authors suggested that concomitant exposure to 

solvents, such as toluene, and PAHs may result in an increased risk for lung cancer. 

Asbestos exerts a synergistic influence on cigarette smoke (which contains several PAHs) in the 

development of bronchopulmonary cancers. This has important implications for workers 

occupationally exposed to asbestos, who also smoke. The interaction between cigarette smoke and 

asbestos may be explained partly by differences in the kinetics of PAH cell uptake when PAHs are 

preadsorbed on asbestos (Foumier and Pezerat 1986). Plutonium oxide (PuO2) has also been shown to 

enhance benzo[a]pyrene-induced lung carcinogenesis following simultaneous inhalation of both 

compounds (Metivier et al. 1984). 

Another component of cigarette smoke, nicotine, may also affect the toxicokinetics of PAHs. When 

introduced in the perfusion medium with benzo[a]pyrene, nicotine inhibited the elimination of 

benzo[a]pyrene from the lung (Foth et al. 1988a). 

Naturally occurring compounds have been found to induce the enzymes that metabolize PAHs, leading 

to either increased or decreased toxicity. Compounds that exert a protective effect against the 

carcinogenicity of PAHs and are enzyme inducers include plant flavonoids, plant phenols, antioxidants, 

retinoids (vitamin A), garlic oil, selenium, molybdenum, turmeric extracts, nitrates, soy sauce, and 

Chinese herbs. Plant flavonoids can induce microsomal monooxygenases and reduce the 

carcinogenicity of benzo[a]pyrene (Weibel 1980). Flavones administered orally or dermally increased 

benzo[a]pyrene hydroxylase activity in the small intestine and skin, respectively, and prevented the 

formation of pulmonary adenomas and forestomach and skin tumors initiated by benzo[a]pyrene 

(Rahimtula et al. 1977; Wattenberg and Leong 1970). A series of flavonoids and isoflavonoids, 

compounds that are found in fruits and vegetables, were tested for their ability to inhibit metabolism of 

benzo[a]pyrene in cultured hamster embryo cells (Chae et al. 1992). The results indicated that the 

flavonoids are generally more active than the isoflavonoids, and that two hydroxyl, two methoxyl, or 

methyl and hydroxyl substituents at the 5- and 7-positions and a 2,3-double bond are the structural 

characteristics required for inhibition of benzo[a]pyrene metabolism to reactive intermediates. Two of 

these compounds, acacetin and kaempferide also inhibited benzo[a]pyrene-induced mutation in Chinese 
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hamster V79 cells. Therefore, the authors proposed that the protective effect of these compounds may 

be due to their ability to inhibit metabolism of benzo[a]pyrene to reactive intermediates. Similar 

results have been obtained with pine cone extracts, tested in the Salmonella typhimurium assay (Lee et 

al. 1993). These results suggest that these compounds may be useful as potential chemopreventive 

agents in individuals exposed to genotoxic/carcinogenic PAHs. 

Dietary plant phenols, such as tannic acid, quercetin, myricetin, and anthraflavic acid exhibit a 

protective effect against the tumorigenicity of benzo[a]pyrene and other PAHs by altering the 

metabolic pathways that detoxify and activate PAHs to their ultimate carcinogenic metabolites, thus 

suppressing PAH metabolism and subsequent PAH-DNA adduct formation. It has been suggested that 

the possible mechanism for the anticarcinogenic effect of these plant phenols may be an inhibitory 

effect on the binding of the ultimate carcinogen to the target tissue DNA (Mukhtar et al. 1988). Oral 

administration of these compounds has been associated with a decrease in tumorigenesis induced by 

benzo[a]pyrene in mouse forestomach (Katiyar et al. 1993a, 1993b; Zheng et al. 1993). Antioxidants 

also affect benzo[a]pyrene hydroxylation by rat liver microsomal MFOs, and reduce the bacterial 

mutagenicity of benzo[a]pyrene in the presence of rat liver microsomes and cofactors (Rahimtula et al. 

1977). Antioxidants such as BHA, BHT, phenothiazine, phenothiazine methosulfate, and ethoxyquin 

all can reduce the quantitative yield of benzo[a]pyrene metabolites in incubations with rat liver 

microsomes (Sullivan et al. 1978). 

Retinoids, of which vitamin A is a member, have demonstrated an antagonistic effect on 

benzo[a]pyrene-induced carcinogenicity. (Vitamin A has been shown to prevent and/or reverse the 

genetic damage caused by benzo[a]pyrene.) Similarly, the ability of benzo[a]pyrene (75 mg/kg, oral 

administration) to induce micronuclei in vivo was completely inhibited in Swiss mice orally 

administered doses of vitamin A ranging from 750 to 1,500 mg/kg 1 hour prior to benzo[a]pyrene 

treatment (Rao and Nandan 1990). Although the protective mechanism has not been fully elucidated, 

it has been suggested that vitamin A interferes with the activation of benzo[a]pyrene to its reactive 

metabolites, thus reducing the amount of benzo[a]pyrene-DNA binding in rat liver and target tissues 

(i.e., stomach) to prevent the mutagenic action of benzo[a]pyrene. It has also been suggested that 

vitamin A can enhance DNA repair (McCarthy et al. 1987; Rao et al. 1986). Conversely, vitamin A 

deficiency enhances the mutagenicity and carcinogenic effect of cigarette smoke and benzo[a]pyrene. 

This activity is related to a decreased level of free radical scavengers like ascorbic acid and glutathione 

in the liver (Alzieu et al. 1987). Another study observed that vitamin A deficient animals exposed to 
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cigarette smoke via inhalation exhibited enhanced benzo[a]pyrene-DNA adduct formation (Gupta et al. 

1987, 1990). This has important implications for humans who smoke and consume diets deficient in 

vitamin A. Mammary tumor incidence was reduced by 27% (from 67% to 40%) in female rats 

receiving retinyl acetate before, during, and after the administration of either a single dose or eight 

weekly doses of benzo[a]pyrene. These results demonstrate that retinyl acetate is capable of inhibiting 

benzo[a]pyrene-induced mammary tumor formation in rats when given before, during, and after 

carcinogen treatment (McCormick et al. 1981). 

The oral gavage administration of 25 mg/kg n-acetylcysteine (NAC) prevented the formation of 

benzo[a]pyrene-diol-epoxide-DNA adducts in rats receiving benzo[a]pyrene by intratracheal instillation 

for 3 consecutive days (25 mg/kg in 2% Tween 80) (De Flora et al. 1991). Inhibition of DNA adduct 

formation was more efficient in the liver than in the lungs. Similarly, micronuclei induction in the 

benzo[a]pyrene-treated rats was completely reversed by NAC. These results suggest that NAC, which 

is a glutathione precursor, may be effective in preventing or reversing the binding of the reactive 

intermediates of PAHs to cellular macromolecules and, therefore, may prevent the subsequent toxic 

effects of PAHs. 

Coumarin, also known to be anticarcinogenic, inhibited benzo[a]pyrene-induced micronuclei in male 

ICR mice pretreated with 65 or 130 mg/kg/day coumarin for 6 days prior to the intraperitoneal 

administration of 150 mg/kg benzo[a]pyrene (Morris and Ward 1992). However, pretreatment with 

either dose of coumarin did not alter the genotoxicity of benzo[a]pyrene when females were included 

in the study. 

Garlic oil also exhibits an antagonistic effect on benzo[a]pyrene by inhibiting benzo[a]pyrene-induced 

skin carcinogenesis in Swiss mice during the initiation phase (Sadhana et al. 1988). A primary 

constituent of garlic oil, allyl methyl trisulfide (ATM), has also demonstrated an inhibitory effect on 

benzo[a]pyrene induced neoplasia of the forestomach in mice (Spamins et al. 1986). 

Selenium has been shown to reduce the mutagenicity of benzo[a]pyrene as well as AHH activity (Lee 

and Lin 1988). Selenium also inhibits the metabolism of benzo[a]pyrene in vitro (Bompart and 

Claments 1990). Several different salts of molybdenum inhibited the formation of certain metabolites 

of benzo[a]pyrene by lung and liver microsomes in vitro obtained from rats pretreated with 

3-methylcholanthrene (Bompart et al. 1989). In a later study, it was demonstrated that benzo[a]pyrene 
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metabolism in vitro by lung and liver microsomes isolated from rats that were exposed to 40 or 

80 mg/kg of ammonium heptamolybdate for 8 weeks was inhibited (Bompart 1990). These results 

suggest that molybdenum interferes with the cytochrome P-450 enzymes responsible for the activation 

of benzo[a]pyrene, and thus may have a protective effect against benzo[a]pyrene-induced 

toxicity/carcinogenicity. 

Ferric oxide has been shown to increase the metabolism of benzo[a]pyrene by hamster alveolar 

macrophages (Greife and Warshawsky 1993). Alveolar macrophages, the primary lung defense cell, 

have been shown to metabolize benzo[a]pyrene to a more biologically active form, and then release 

the metabolites. Concurrent exposure of hamster alveolar macrophages to benzo[a]pyrene-coated ferric 

oxide resulted in a significant increase in the amount of benzo[a]pyrene metabolites and superoxide 

anions, which have been shown to produce localized lipid peroxidation and edema in vivo. 

Lindane, an isomer of hexachlorocyclohexane, is an organochlorine pesticide which is extensively used 

in agricultural and public health programs in developing countries (Khan et al. 1993). Pretreatment of 

rat lungs with lindane by intratracheal injection inhibited benzo[a]pyrene hydoxlase activity in the 

lungs. Reduced elimination of intravenously administered benzo[a]pyrene from the lungs of rats after 

lindane pretreatment was also observed, suggesting that lindane may alter the clearance of 

benzo[a]pyrene from the lungs. 

Prostacyclin has been shown to significantly reduce genetic damage caused by benzo[a]pyrene to 

mouse bone marrow cells, using the micronucleus test (Koratkar et al. 1993). 

Chinese herbs commonly used in anti-cancer drugs have also been demonstrated to inh.ibit the 

mutagenicity of benzo[a]pyrene (Lee and Lin 1988). 

Aqueous extracts of turmeric, curcumin-free aqueous turmeric extract, and curcumin were tested for 

their ability to inhibit benzo[a]pyrene-induced mutagenicity in the S. typhimurium assay and the bone 

marrow micronucleus test in Swiss mice (Azuine et al. 1992). A dose-dependent inhibition of 

benzo[a]pyrene-induced mutagenicity was observed in two strains of Salmonella in the presence of 

Aroclor-1254-induced rat liver homogenate, and 3 mg/kg of these three extracts also significantly 

inhibited benzo[a]pyrene-induced bone marrow micronuclei formation by 43%, 76%, and 65%. 

Female Swiss mice were treated with either aqueous turmeric extract (3 mg/day), curcumin-free 
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aqueous tumeric extract (1 mg/day), or curcumin (1 mg/day) 5 days/week for 2 weeks prior to 

receiving twice weekly gavage administrations of 20 mg/kg benzo[a]pyrene for 4 weeks. The 

treatment with the turmeric extracts continued for another 2 weeks after the cessation of benzopyrene 

treatment, and the animals were observed until they were 180 days old. A group receiving just 

benzo[a]pyrene served as a control. The benzo[a]pyrene-only animals exhibited 100% tumor incidence 

with 9.1±0.6 papillomas/mouse. All three extracts significantly (p<0.00l) inhibited the formation of 

forestomach papillomas by benzo[a]pyrene by 53%, and the average numbers of papillomas per mouse 

was also significantly decreased. The authors suggested that turmeric extracts may be useful as 

chemopreventive agents, but that there are probably several mechanisms of action for these inhibitory 

effects. 

Pumark, a mixture of solvent extracts of tumeric, betel leaf, and catechu, was tested for its 

chemopreventative activity against benzo[a]pyrene-induced DNA damage (Ghaisas and Bhide 1994) 

Sister chromatid exchange and micronuclei formation in human lymphocyte culture were used as 

markers to assess the protective effect of Pumark. Pumark gave 50-60% protection against 

benzo[a]pyrene-induced chromosomal damage. 

Other environmentally ubiquitous substances, such as nitrites and nitrates, have been shown to interact 

with PAHs. Pyrene is not mutagenic in the Salmonella typhimurium assay. However, when injected 

intraperitoneally into mice at doses of 10-200 mg/kg in combination with inhalation exposure to 50 or 

100 ppm nitrous oxide (NO2) mutagenic metabolites of pyrene were recovered from the urine (Kanoh 

et al. 1990). In addition, 1-hydroxypyrene (the major urinary metabolite of pyrene) administration in 

combination with NO2 exposure also produced mutagenic metabolites. These results suggest that 

combined exposure to pyrene, a prevalent environmental PAH, and nitrogen compounds could result in 

the formation of nitrogenated mutagenic metabolites of pyrene. The effects of nitrite (in drinking 

water) and/or soy sauce (in food) on the forestomach tumors induced by twice weekly gavage 

administration of eight total doses of benzo[a]pyrene were studied in mice (Benjamin et al. 1988). 

The combination of nitrite and soy sauce resulted primarily in a significant reduction in the number of 

tumors per animal induced by benzo[a]pyrene, and also a reduction in tumor incidence. Neither 

substance alone had much effect on the carcinogenicity of benzo[a]pyrene alone. The mechanism for 

this protective effect of the combination of nitrite and soy sauce is not known. 
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The fungicide (prochloraz), the topical antifungal agent (miconazole) and 7,8-benzoflavone (a strong 

inhibitor of cytochrome P-450) limited the conversion of benzo[a]pyrene to a mutagen in S. 

typhimurium. The order of inhibitory action was 7,8-benzoflavone ≥ prochloraz ≥ miconazole 

(Antignac et al. 1990). 

Diet (i.e., dietary fat levels) can also have an effect on the disposition and toxicity of PAHs. The 

metabolism of benzo[a]pyrene in hepatocytes in vitro from rats fed high-fat (as corn oil) diets was 

decreased (Zaleski et al. 1991). This effect was not due to a decrease in the activity of AHH. The 

authors postulated that the high-fat diets allowed benzo[a]pyrene, which is highly lipophilic, to become 

sequestered in lipid droplets and, therefore, become inaccessible to the P-450 enzymes. Another study 

suggests that caloric restriction, per se, can reduce the metabolic activation capacity of the liver, 

thereby reducing the production of mutagenic metabolites from PAHs (Xiao et al. 1993). 

Organosulphur compounds, e.g., S-methyl cysteine sulphoxide (SMCSO) and its metabolite methyl 

methane thiosulphinate, both naturally occurring compounds present in Brassica vegetables (broccoli, 

cabbage) were found to inhibit benzo[a]pyrene-induced micronucleus formation in mouse bone marrow 

by 31 and 33%, respectively, after oral administration (Marks et al. 1993). 

o-Cresol, often found in conjunction with PAHs in industrial waste from coking, oil processing, shale 

processing, and other industries, was found to protect mice from benzo[a]pyrene-induced forestomach 

tumors after oral administration, when o-cresol was administered before or after benzo[a]pyrene 

(Yanysheva et al. 1993). 

Topical application of 1% croton oil twice weekly at 4 weeks of age had an inhibitory effect on tumor 

formation in offspring of ICF/Ha female mice treated with intraperitoneal injections of benzo[a]pyrene 

in sesame oil on the llth, 13th and 15th days of pregnancy (Bulay and Wattenberg 1971). 

Administration of a diet containing 3% myo-inositol to mice beginning one week after oral 

benzo[a]pyrene administration reduced the number of pulmonary adenomas by 40% but did not 

prevent forestomach tumors (Estensen and Wattenberg 1993). Under the same conditions, 

administration of 0.5% dexamethasone in the diet inhibited pulmonary adenomas by 57% and also 

inhibited forestomach tumor formation to a similar degree. A combination of the two compounds 

resulted in additive chemoprevention. 
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The ubiquitous nature of PAHs in the environment, particularly as constituents of complex mixtures 

such as automobile emissions, coal tar, coke oven emissions, and combustion products of tobacco, 

increases the likelihood that the type of interactions discussed will occur. Thus, interactions may play 

a decisive role in the expression of toxicity and the development of cancer in exposed populations. 

2.7 POPULATIONS THAT ARE UNUSUALLY SUSCEPTIBLE 

A susceptible population will exhibit a different or enhanced response to PAHs than will most persons 

exposed to the same level of PAHs in the environment. Reasons include genetic make-up, 

developmental stage, age, health and nutritional status (including dietary habits that may increase 

susceptibility, such as inconsistent diets or nutritional deficiencies), and substance exposure history 

(including smoking). These parameters result in decreased function of the detoxification and excretory 

processes (mainly hepatic, renal, and respiratory) or the pre-existing compromised function of target 

organs (including effects or clearance rates and any resulting end-product metabolites). For these 

reasons we expect the elderly with declining organ function and the youngest of the population with 

immature and developing organs will generally be more vulnerable to toxic substances than healthy 

adults. Populations who are at greater risk due to their unusually high exposure are discussed in 

Section 5.6, Populations With Potentially High Exposure. 

Data suggest that specific subsections of the population may be susceptible to the toxic effects 

produced by exposure to PAHs. These include people with various conditions, such as aryl 

hydrocarbon hydroxylase (AHH) that is particularly susceptible to induction, nutritional deficiencies, 

genetic diseases that influence the efficiency of DNA repair, and imrnunodeficiency due to age or 

disease. Other subsections of the population that may be susceptible to the toxic effects of PAHs are 

people who smoke, people with a history of excessive sun exposure, people with liver and skin 

diseases, and women, especially of child-bearing age. Human fetuses may also be particularly 

susceptible to the toxic effects produced by exposure to PAHs. Data also indicate that the general 

population may be at increased risk of developing lung cancer following prolonged inhalation of 

PAH-contaminated air, and skin cancer following concurrent dermal exposure to PAHs and sunlight. 

There is some limited evidence that indicates that all people could be sensitive at some point to the 

toxic effects of environmental contaminants, such as PAHs, as a result of stress and/or circadian 

rhythms. 
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Pre- and post-natal exposure to PAHs could produce adverse reproductive and developmental effects in 

human fetuses. Most PAHs and their metabolites cross the placenta because of their lipid solubility 

(Calabrese 1978; Shendrikova and Aleksandrov 1974). Fetuses are susceptible to the toxic effects 

produced by maternal exposure to PAHs, such as benzo[a]pyrene, because of an increased permeability 

of the embryonic and fetal blood-brain barrier and a decreased liver-enzyme conjugating function 

(Calabrese 1978; Shendrikova and Aleksandrov 1974). Because of PAH exposure, higher incidences 

of embryo- and fetolethality, stillbirths, resorptions, and malformations of the kidney and bladder have 

been observed in animals (Legraverend et al. 1984; Urso and Gengozian 1980; Urso and Johnson 

1987). Delayed effects have been observed in the progeny of mothers exposed to PAHs, such as 

benzo[a]pyrene (Urso and Gengozian 1980). These .delayed effects include sterility of progeny, 

immune suppression, possible alteration of endocrine function, and cancer in rodents (Csaba and 

Inczefi-Gonda 1992; Csaba et al. 1991; Legraverend et al. 1984; Mackenzie and Angevine 1981). 

Tobacco smoke contains both PAHs and particulate matter. These could interact synergistically in 

pregnant women who smoke to produce decreased birth weight, increased perinatal morbidity and 

mortality, and other diseases of the newborn (NRC 1983). PAHs in cigarette smoke, such as 

benzo[a]pyrene, have been associated with the induction of AHH activity in human placental tissue 

and a decrease in placental size (NRC 1983). Results of in vitro studies indicate that benzo[a]pyrene 

alters human placental endocrine and metabolic function (Avigdor et al. 1992; Bamea and 

Shurtz-Swirski 1992; Guyda 1991; Guyda et al. 1990). 

People with AHH that is particularly susceptible to induction may also be susceptible to the possible 

carcinogenic effect of exposure to PAHs. This enzyme is mixed function oxidase (MFO) that is 

involved in the metabolism of PAHs to certain reactive intermediates that can cause cell 

transformation, mutagenicity, and cytotoxicity. The incidence of this inducible genetic trait is low in 

53%, intermediate in 37%, and high in 10% of the white population in the United States (Calabrese 

1978). It has been proposed that genetically expressed AHH inducibility is related to the development 

of bronchogenic carcinoma in persons exposed to PAHs contained in tobacco smoke. On the other 

hand, individuals with a greater ability to induce AHH may be able to rapidly detoxify PAHs and 

eliminate them, thus making them less susceptible to the toxic effects of PAHs. Based on the 

population frequency of genetically controlled AHH induction, some scientists predict that 

approximately 45% of the general population are considered to be at high risk, and 9% of the 45% are 

considered to be at very high risk, of developing bronchogenic carcinoma following exposure to PAHs 

(Calabrese 1978). 
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Certain nutritional deficiencies have been associated with an increased cancer incidence in 

PAH-exposed animals. These include deficiencies in vitamins A and C, iron, and riboflavin 

(Calabrese 1978). It is estimated that at least 25% of all children between the ages of 7 and 12 years 

and all children of low-income families consume less than the recommended dietary allowance (RDA) 

of vitamin A. It has also been estimated that between 10% and 30% of the infants, children, and 

adults of low-income groups consume less than the RDA for vitamin C, 98% of all children consume 

less than the RDA for iron, and 30% of women and 10% of men between the ages of 30 and 60 years 

consume less than two-thirds of the RDA for riboflavin (Calabrese 1978). Other nutrients such as 

vitamin D, selenium, and protein can also influence the cancer incidence in animals exposed to PAHs 

(NRC 1983; Prasanna et al. 1987). Several studies have been conducted to investigate the interaction 

between nutrition and PAH exposure by administering benzo[a]pyrene to laboratory animals. The 

nutritional factors listed above either reduced the amount of benzo[a]pyrene binding to DNA in rat 

liver or forestomach tissue (McCarthy et al. 1987), prevented or reversed genetic damage (Rao et al. 

1986), or reduced the activity of AHH (Prasanna et al. 1987). It has also been observed that fasted 

rats showed altered toxicokinetics of PAHs resulting from benz[a]anthracene and chrysene exposure. 

This altered toxicokinetics included reduced hepatic clearance because of decreased AHH activity and 

the acceleration of the depletion of cytochrome P-450 and other microsomal enzymes required to 

metabolize PAHs (Fiume et al. 1983). 

Individuals who undergo a rapid reduction in body fat may be at risk from increased toxicity because 

of the systemic release and activation of PAHs that had been stored in fat. The metabolism of 

benzo[a]pyrene in hepatocytes in vitro from rats fed high-fat (as corn oil) diets was decreased (Zaleski 

et al. 1991). This effect was not due to a decrease in the activity of AHH. The authors postulated 

that the high-fat diets allowed benzo[a]pyrene, which is highly lipophilic, to become sequestered in 

lipid droplets and, therefore, become inaccessible to the P-450 enzymes. 

People exposed to PAHs in conjunction with particulates from tobacco smoke, fossil-fuel combustion, 

coal fly ash, and asbestos fibers are at increased risk of developing toxic effects, primarily cancer. 

Even people not susceptible to the toxic effects of PAHs may become affected when exposure occurs 

in conjunction with exposure to particulates (NRC 1983). This enhanced effect results from the 

adsorption of PAHs onto the particulates. They are vacuolized into cells, and distributed differently in 

tissues depending on the size and type of particulate matter. This increased PAH uptake may result in 

more efficient induction of AHH activity at low PAH concentrations. This activity also increases the 
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dose to the gastrointestinal tract as a result of mucocilliary clearance (NRC 1983). This synergistic 

action between PAHs and particulate matter in air pollution has been associated with the occurrence of 

stomach cancer in humans (Fraumeni 1975). 

Immunocompetence is an important factor in decreasing or preventing human susceptibility to toxicity 

and disease after exposure to environmental contaminants. Small children have an immature immune 

system, and the elderly may have a deficient immune system due to age, genetic factors, or disease 

(Calabrese 1978; NRC 1983). It is possible that individuals whose immune systems are compromised 

could be at an increased risk of carcinogenesis, including that produced by PAHs. Some genetic 

diseases that may predispose a person to immune deficiency include ataxia telangiectasia, 

Wiskott-Aldrich syndrome, Bloom’s syndrome, common variable immunodeficiency, selective IgA 

deficiency, Bruton’s agammaglobulinemia, severe combined immunodeficiency, selective IgM 

deficiency, AIDS, and immunodeficiency with normal or increased immunoglobulins (NRC 1983). 

Genetic diseases that reduce DNA-repair capabilities also increase an individual’s susceptibility to 

PAH-related malignancy by reducing the efficiency of DNA repair. In fact, the level of 

benzo[a]pyrene/DNA adducts in peripheral lymphocytes was slightly but significantly higher in 

22 lung cancer patients who had at least one “first-degree” relative with lung cancer than in 30 healthy 

controls (Nowak et al. 1992). This finding led the authors to speculate that altered metabolic 

activation and deactivation and increased formation of adducts may indicate a genetic predisposition 

for lung cancer. Some of the diseases that reduce DNA repair capability are also associated with an 

abnormality of the immune system (NRC 1983). Diseases that may be associated with DNA-repair 

deficiencies are classical and variant xeroderma pigmentosum, ataxia telangiectasia, Bloom’s 

syndrome, Fanconi’s anemia, familial retinoblastoma, D-deletion retinoblastoma, progeria 

(Hutchinson-Gilford) syndrome, Down’s syndrome, dyskeratosis congenita, Cockayne’s syndrome, 

actinia keratosis, and cutaneous malignant melanoma (NRC 1983). 

Women may be at increased risk of reproductive dysfunction following exposure to high levels of 

PAHs. Data from animal studies indicate that oocyte and follicle destruction occurs following dosing 

with PAHs (Mattison et al. 1989; Miller et al. 1992; Urso and Gengozian 1980; Urso and Johnson 

1987). Benzo[a]pyrene exposure may reduce fertility and the ability to bear children (Mackenzie and 

Angevine 1981; Rigdon and Rennels 1964). Exposure may also reduce fertility of exposed women by 
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causing ovarian dysfunction (Swartz and Mattison 1985). However, the doses that produced the 

effects discussed above are high relative to expected environmental exposures to PAHs. 

Subsections of the population that suffer from liver and skin diseases may be at increased risk of 

developing adverse effects from exposure to PAHs. People with pre-existing skin conditions, such as 

pemphigus vulgaris and xeroderma pigmentosum, and those with normal skin may be at increased risk 

of developing adverse dermal effects ranging from rashes to cancer following exposure to some PAHs, 

such as benzo[a]pyrene, anthracene, benz[a]anthracene, and pyrene (Bingham and Falk 1969; Cavalieri 

et al. 1977, 1988b; Cottini and Mazzone 1939; Forbes et al. 1976; Habs et al. 1980; Shubik and Porta 

1957). Exposure to more than one PAH may enhance or reduce tumor development (Slaga et al. 

1979; Van Duuren and Goldschmidt 1976; Van Duuren et al. 1973). Skin cancer induction in 

laboratory animals has been associated with exposure to benz[a]anthracene, benzo[a]pyrene, 

benzo[b]fluoranthene, benzohlfluoranthene, benzo[k]fluoranthene, chrysene, dibenz[a,h]anthracene, and 

indeno[ 1,2,3-c,d]pyrene. 

People with significant exposure to ultraviolet radiation, such as from sunlight, may also be at 

increased risk of developing skin cancer due to PAH exposure. Ultraviolet radiation has a synergistic 

influence on PAH-induced skin cancer following dermal exposure. It enhances benzo[a]pyreneinduced 

skin carcinogenesis in the mouse, which is dependent on the dose of benzo[a]pyrene (Gensler 

1988). Combined exposure to anthracene and sunlight could produce mutagenic lesions (Blackbum 

and Taussig 1975; Forbes et al. 1976). Laboratory animals exposed concurrently to chronic ultraviolet 

irradiation and to PAHs were at a higher risk of skin tumor induction (Mukhtar et al. 1986). 

2.8 METHODS FOR REDUCING TOXIC EFFECTS 

This section will describe clinical practice and research concerning methods for reducing toxic effects 

of exposure to PAHs. However, because some of the treatments discussed may be experimental and 

unproven, this section should not be used as a guide for treatment of exposures to PAHs. When 

specific exposures have occurred, poison control centers and medical toxicologists should be consulted 

for medical advice. 
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2.8.1 Reducing Peak Absorption Following Exposure 

General recommendations to reduce absorption follo.wing acute exposure to PAHs include removing 

the individual from the source of exposure, removing contaminated clothing, and decontaminating 

exposed areas of the body. It has been suggested that contaminated skin should be washed with soap 

and water, and eyes exposed to PAHs should be flushed with water or saline (Stutz and Janusz 1988). 

Administration of activated charcoal following ingestion of PAHs is recommended; however, it has not 

been proven to reduce absorption of PAHs in the gastrointestinal system (Stutz and Janusz 1988). The 

use of emetics as a means of gastrointestinal decontamination of PAHs is not recommended (Bronstein 

and Currance 1988). There is a risk of causing chemical pneumonitis in the patient by the aspiration 

of the PAHs. 

2.8.2 Reducing Body Burden 

There are no known methods currently available for reducing the body burden of PAHs. Evidence 

from acute-duration studies in experimental animals indicates that PAHs are rapidly metabolized and 

conjugated to form water-soluble metabolites that are essentially completely eliminated in the urine 

and feces within a matter of days (see Sections 2.3.3 and 2.3.4). No data are available on the kinetics 

of PAHs following chronic exposure, so it is not known if PAHs or their metabolites bioaccumulate in 

these exposure situations. Given the relatively rapid and complete excretion observed following 

short-term exposures, it is not likely that PAHs bioaccumulate to an appreciable degree. However, 

PAHs are lipophilic, so it is conceivable that unmetabolized parent compound could accumulate in 

tissue fat stores. In fact, diet (i.e., dietary fat levels) may have an effect on the disposition and 

toxicity of PAHs. The metabolism of benzo[a]pyrene in hepatocytes in vitro from rats fed high-fat (as 

corn oil) diets was decreased (Zaleski et al. 1991). This effect was not due to a decrease in the 

activity of AHH. The authors postulated that the high-fat diets allowed benzo[a]pyrene, which is 

highly lipophilic, to become sequestered in lipid droplets and, therefore, become inaccessible to the 

P-450 enzymes. Therefore, high-fat diets may favor the accumulation of parent PAHs in lipids so that 

they are not metabolized to reactive intermediates or water-soluble conjugates. Alternatively, rapid fat 

loss may result in the release of sequestered parent PAHs, making them available to the P-450 

enzymes to be metabolized to reactive intermediates as well as water-soluble conjugates that can be 

easily excreted. Thus, modulating body fat content may reduce body burden of PAHs by hastening 
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their metabolism to water-soluble conjugates. However, the result may also be an increase in toxicity 

due to increased metabolism to reactive intermediates. 

2.8.3 Interfering with the Mechanism of Action for Toxic Effects 

As discussed in Sections 2.3.5 and 2.4, it is currently believed that the toxic and carcinogenic effects 

of PAHs are mediated by reactive diol-epoxide intermediates that interact directly with DNA and 

RNA, producing adducts. The formation of these adducts leads to neoplastic transformation as well as 

interfering with the normal functioning of rapidly proliferating tissues. As discussed above, these 

reactive intermediates are formed when PAHs are biotransformed by the P-450 enzymes. Interference 

with these metabolic pathways, by inactivation of the activated diol epoxides, reduction in tissue levels 

of cytochrome P-450, and direct inhibition of the cytochrome P-450 enzymes responsible for the 

formation of the reactive intermediates, could reduce the toxic and carcinogenic effects of PAHs. A 

number of drugs, such as cobaltous chloride, SKF-525-A, and 6-nitro- 1,2,3-benzothioadiazole, have 

been reported to inhibit P-450 enzymes. In addition, as discussed in Section 2.7, other compounds 

that exert a protective effect against the carcinogenicity of PAHs by interfering with cytochrome P-450 

enzymes include plant flavonoids, plant phenols, antioxidants (such as BHA, BHT, phenothiazine, 

phenothiazine methosulfate, and ethoxyquin), retinoids (vitamin A), garlic oil, selenium, organosulphur 

compounds, o-cresol, myo-inositol, lindane, and molybdenum (Bompart 1990; Bompart et al. 1989; 

Chae et al. 1992; Estensen and Wattenberg 1993; Ghaisas and Bhide 1994; Katiyar et al. 1993a, 

1993b; Khan et al. 1993; Lee and Lin 1988; Lee et al. 1993; Marks et al. 1993; Mukhtar et al. 1988; 

Rahimtula et al. 1977; Rao and Nandan 1990; Sullivan et al. 1978; Weibel 1980, Yanysheva et al. 

1994; Zheng et .al. 1993). Reduced caloric intake has also been shown to cause decreased metabolism 

of PAHs by liver microsomes, thus protecting against genotoxic effects (Xiao et al. 1993). P-450 

metabolism also results in products that can be more readily eliminated than can the parent compound. 

Hence, any side products of the drugs or substances listed above, along with their potential to increase 

the biological half-life of the PAHs, would also need to be considered in any protocol. Further 

research to determine which cytochrome P-450 isozymes are involved in the metabolism to the 

reactive intermediates, as well as which isozymes are involved in enhancing the elimination of PAHs, 

could lead to the development of strategies to selectively inhibit specific isozymes and, thus, reduce 

the toxic effects of PAHs. 
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It has also been suggested that some of these compounds may act at other points in the activation, 

macromolecular binding sequence described above. For instance, vitamin A can also enhance DNA 

repair (McCarthy et al. 1987; Rao et al. 1986). 

Because PAHs are detoxified by conjugation with substances such as glutathione (see Sections 2.3.3 

and 2.3.4), sufficient glutathione stores in the body may reduce the chances of toxic effects following 

acute exposure to PAHs. For example, the oral gavage administration of NAC prevented the 

formation of benzo[a]pyrene-diol-epoxide-DNA adducts in rats receiving benzo[a]pyrene (De Flora et 

al. 1991). Inhibition of DNA adduct formation was more efficient in the liver than in the lungs. 

Similarly, micronuclei induction in the benzo[a]pyrene-treated rats was completely reversed by NAC. 

These results suggest that NAC, which is a glutathione precursor, may be effective in preventing or 

reversing the binding of the reactive intermediates of PAHs to cellular macromolecules and, therefore, 

may prevent the subsequent toxic effects of PAHs. 

2.9 ADEQUACY OF THE DATABASE 

Section 104(i)(5) of CERCLA, as amended, directs the Administrator of ATSDR (in consultation with 

the Administrator of EPA and agencies and programs of the Public Health Service) to assess whether 

adequate information on the health effects of PAHs is available. Where adequate information is not 

available, ATSDR, in conjunction with the National Toxicology Program (NTP), is required to assure 

the initiation of a program of research designed to determine the health effects (and techniques for 

developing methods to determine such health effects) of PAHs. 

The following categories of possible data needs have been identified by a joint team of scientists from 

ATSDR, NTP, and EPA. They are defined as substance-specific informational needs that if met would 

reduce the uncertainties of human health assessment. This definition should not be interpreted to mean 

that all data needs discussed in this section must be filled. In the future, the identified data needs will 

be evaluated and prioritized, and a substance-specific research agenda will be proposed. 

2.9.1 Existing Information on Health Effects of Polycyclic Aromatic Hydrocarbons 

The existing data on health effects of inhalation, oral, and dermal exposure of humans and animals to 

PAHs are summarized in Figure 2-5. The purpose of this figure is to illustrate the existing 

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information concerning the health effects of PAHs. Each dot in the figure indicates that one or more 

studies provide information associated with that particular effect. The dot does not imply anything 

about the quality of the study or studies. Gaps in this figure should not be interpreted as “data needs.” 

A data need, as defined in ATSDR’s Decision Guide for Identifying Substance-Specific Data Needs 

Related to Toxicological Profiles (ATSDR 1989), is substance-specific information necessary to 

conduct comprehensive public health assessments. Generally, ATSDR defines a data gap more 

broadly as any substance-specific information missing from the scientific literature. 

The vast majority of literature reviewed concerning the health effects of PAHs in humans described 

case reports and chronic-duration studies in workers linking the occurrence of lung and skin cancer 

and adverse noncancer skin effects with exposure to PAH-containing mixtures such as coke oven 

emissions, roofing-tar emissions, shale oils, and soot, and exposure to cigarette smoke. The 

predominant routes of exposure in the studies are inhalation and dermal, but the possibility of some 

degree of oral exposure cannot be ruled out, especially in light of muco-cilliary clearance and 

ingestion following inhalation exposure. Because of the lack of quantitative exposure information and 

the presence of other potentially carcinogenic substances in these mixtures, it is impossible to evaluate 

the contribution of an individual PAH or even the PAHs as a class to the effects observed. 

The database for the health effects of PAHs in experimental animals consists primarily of older animal 

studies that would be considered inadequate by current standards, and two-stage dermal carcinogenesis 

studies. As can be seen in Figure 2-5, very little information is available on the effects of inhalation 

exposure to PAHs in animals. However, oral and dermal exposures to relatively high doses of PAHs 

have been shown in numerous studies to induce skin, lung, and forestomach tumors in animals, and 

noncancer adverse effects in rapidly proliferating tissues such as bone marrow, lymphoid organs, 

gonads, and intestinal epithelium. Benzo[a]pyrene is by far the most extensively studied of the PAHs; 

therefore, the adverse effects of other less-studied PAHs must generally be inferred from the results 

obtained with benzo[a]pyrene. This may over- or underestimate the health risk associated with the 

various PAHs. 

2.9.2 Identification of Data Needs 

Acute-Duration Exposure. Little is known regarding the adverse health effects associated with 

acute-duration inhalation exposure to any of the PAHs in either humans or animals. Limited 
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information is available on the effects of acute-duration oral and dermal exposures to PAHs in 

animals; the skin and the liver have been identified as target organs of PAH toxicity in animals (Iwata 

et al. 1981; Nousiainen et al. 1984). Available information is insufficient to derive an acute inhalation 

or oral MRL. Identification of target organs (other than the developing fetus) from acute-duration 

animal studies following inhalation and oral exposures would be helpful in order to assess the risk 

associated with the acute inhalation of contaminated air or ingestion of PAH-contaminated water or 

soils by humans living in areas surrounding hazardous waste sites. Additional inhalation and oral 

studies in animals involving a range of exposure concentrations and employing sensitive histological 

and biochemical measurements of injury to a comprehensive set of end points would be useful for 

establishing dose-response relationships and identifying thresholds for these effects. This information 

would be useful for determining levels of significant exposure to PAHs that are associated with 

adverse health effects. Both routes are considered significant for individuals living in the vicinity of 

hazardous waste sites because exposures to particulate PAHs in air and PAHs bound to soil particles, 

sediments in water, and contaminated food are possible in such areas. Furthermore, the 

pharmacokinetic data on PAHs are insufficient to determine whether similar effects may be expected 

to occur across different routes of exposure. Additional studies should be conducted on the effects of 

PAHs after acute-duration dermal exposure, since dermal exposure may be important to populations 

around hazardous waste sites. It is known that acute-duration dermal exposure to PAHs results in 

adverse dermal effects. Further studies determining the relative importance of exposure by this route, 

with regard to subsequent toxicity, would be useful. Studies describing dermal and oral absorption of 

PAHs from complex mixtures, including soil and other mixtures that may actually be the vehicles of 

human exposure, would be useful in furthering understanding of the toxicity of these compounds. The 

studies should be conducted with benzo[a]pyrene, because this PAH has been found at the highest 

number of NPL sites and it is a representative altemant PAH; studies should also be conducted with a 

representative nonaltemant PAH such as fluoranthene, benzo[b]fluoranthene, or benzo[j]fluoranthene. 

In addition, since PAHs rarely occur alone, but are found in the environment in combinations and 

mixtures, studies, particularly feeding studies, addressing mixtures of PAHs should be conducted. In 

all cases, research should be cOnducted to determine the most appropriate animal species for 

extrapolation to humans. 

Intermediate-Duration Exposure. Little is known regarding the adverse health effects associated 

with intermediate-duration inhalation exposure to any of the PAHs in either humans or animals. One 

inhalation study in rats failed to establish an effect level (Wolff et al. 1989c). Information is available 
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on the effects of intermediate-duration exposures to some of the PAHs in humans (dermal) and in 

animals (oral and dermal). Regressive verrucae and other epidermal changes were noted in the skin of 

human volunteers treated with topically applied benzo[a]pyrene (Cottini and Mazzone 1939). 

Intermediate-duration dermal exposure to benzo[a]pyrene in patients with preexisting dermal conditions 

of pemphigus vulgaris and xeroderma pigmentosum was associated with an exacerbation of the 

abnormal skin lesions (Cottini and Mazzone 1939). Target organs identified in animal studies with 

some of the PAHs were the skin, the liver, and the hemolymphatic system (Legraverend et al. 1983; 

Old et al. 1963; Robinson et al. 1975). The available information is insufficient to derive an 

intermediate inhalation MRL for PAHs because no intermediate-duration inhalation animal studies 

exist that adequately describe the effects of inhalation exposure to PAHs. Intermediate duration MRLs 

have been derived for acenaphthene, fluoranthene, fluorene, and anthracene, based on 90-day gavage 

studies in mice (EPA 1988e; 1989c; 1989d, 1989e). For acenaphthene, fluoranthene, and fluorene, 

liver effects, supported by effects in other organ systems were identified as the target toxicity. For 

anthracene, no effect was seen in the liver or any other organ system, even at the highest dose of 

1,000 mg/kg/day. The PAHs in these studies were administered by gavage, a route that does not 

mimic the potential exposure of people living near hazardous waste sites. Identification of target 

organs from intermediate-duration animal studies following inhalation and oral (drinking water) 

exposures would be useful in order to assess the risk associated with the intermediate-duration 

inhalation of contaminated air or ingestion of PAH-contaminated water or soils by humans living in 

areas surrounding hazardous waste sites. Ninety-day studies in animals by the inhalation and oral 

(drinking water) routes would be helpful to establish dose-response relationships and to identify other 

possible target organs or systems in individuals living around hazardous waste sites who can be 

exposed to low levels of PAHs for an intermediate-duration period of time. Both routes are 

considered important for individuals living in the vicinity of hazardous waste sites because exposure to 

particulate PAHs in air and PAHs bound to soil particles, sediments in water, and contaminated food 

are significant routes of exposure for individuals living in the vicinity of hazardous waste sites. 

Furthermore, the pharmacokinetic data on PAHs are insufficient to determine whether similar effects 

may be expected to occur across different routes of exposure. Additional studies should be conducted 

on the effects of PAHs after intermediate-duration dermal exposure, since dermal exposure may be 

important to populations around hazardous waste sites. It is known that acute-duration dermal 

exposure to PAHs results in adverse dermal effects. Further studies determining the relative 

importance of exposure by this route, with regard to subsequent toxicity, would be useful. Studies 

describing dermal and oral absorption of PAHs from complex mixtures, including soil and other 
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mixtures that may actually be the vehicles of human exposure, would be useful in furthering the 

understanding of the toxicity of these compounds. The studies should be conducted with 

benzo[a]pyrene, because it has been found at the hi.ghest number of NPL sites and it is a representative 

altemant PAH, and with a representative nonaltemant PAH such as fluoranthene, benzo[b]fluoranthene, 

or benzofilfluoranthene. In addition, since PAHs rarely occur alone, but are found in the environment 

in combinations and mixtures, studies, particularly feeding studies, addressing mixtures of PAHs 

should be conducted. In all cases, research should be conducted to determine the most appropriate 

animal species for extrapolation to humans. 

Chronic-Duration Exposure and Cancer. Few controlled epidemiological studies have been 

reported in humans on the effects of exposure to PAHs or PAH-containing mixtures; such studies 

would be difficult to conduct because of the presence of too many confounding factors. However, 

information is available on the effects of chronic-duration dermal exposures to PAH-containing 

mixtures in humans. Workers exposed to substances that contain PAHs (e.g., coal tar) experience 

chronic dermatitis and hyperkeratosis (EPA 1988a). Several chronic ingestion, intratracheal 

installation, and skin-painting studies have been conducted in animals using various PAHs, but none 

identified adverse effects other than cancer. Therefore, threshold levels for chronic-duration inhalation 

and oral exposure have not been thoroughly investigated, and no MRLs have been developed from this 

database. Although the existing animal studies are inadequate to establish threshold levels and 

dose-response relationships for toxic effects resulting from chronic exposure to PAHs, the data from 

90-day studies recommended above should be evaluated before chronic studies are conducted. 

Inhalation and ingestion are probably the most significant routes of exposure for individuals living in 

the vicinity of hazardous waste sites contaminated with PAHs. Low dose chronic studies are needed 

to mimic these exposures. 

Human data on the carcinogenicity of PAHs are available only for mixtures containing PAHs. Animal 

carcinogenicity data are available for only benzo[a]pyrene following inhalation exposure, for a limited 

number of PAHs following ingestion, and for almost all of the 17 PAHs following dermal exposure. 

A large database on carcinogenicity exists on complex mixtures that contain PAHs (such as crude oils, 

various high boiling point distillates, complex petroleum products, coal tars, creosote, and the products 

of coal liquification processes). It is difficult to ascertain the carcinogenicity of the component PAHs 

in these mixtures because of the potential interactions that could occur and the presence of other 

carcinogenic substances in the mixtures. Furthermore, the levels of PAHs were not quantified in any 
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of these reports. However, most of the available information on the carcinogenicity of PAHs in 

humans must be inferred from studies that reported the effects of exposure to complex mixtures that 

contain PAHs. Epidemiologic studies have shown increased mortality due to lung cancer in humans 

exposed to coke oven emissions (Lloyd 1971; Mazumdar et al. 1975; Redmond et al. 1976), 

roofing-tar emissions (Hammond et al. 1976), and cigarette smoke (Maclure and MacMahon 1980; 

Wynder and Hoffmann 1967). Despite the limitations inherent in these studies, reports of this nature 

provide qualitative evidence of the potential for mixtures containing PAHs to cause cancer in humans, 

and more definitive studies in humans on individual PAHs are not recommended at this time. 

Inhalation exposure to benzo[a]pyrene has been shown to induce respiratory tract tumors in hamsters 

(Thyssen et al. 1981). Certain PAHs are carcinogenic to animals by the oral route (e.g., 

benz[a]anthracene, benzo[a]pyrene, and dibenz[a,h]anthracene), and tumors have been noted in the 

liver, mammary gland, and respiratory and gastrointestinal tracts following oral administration of these 

compounds (Neal and Rigdon 1967; Rigdon and Neal 1969). However, only a few PAHs have been 

assayed by the oral route. The results of dermal studies indicate that benz[a]anthracene, 

benzo[a]pyrene, benzo[b]fluoranthene, benzo[j]fluoranthene, benzo[k]fluoranthene, chrysene, 

dibenz[a,h]anthracene, and indeno[l,2,3-c,d]pyrene are tumorigenic in rats and mice following dermal 

exposure. Although many of these studies would be considered inadequate by current standards, the 

results nevertheless indicate that these PAHs can induce skin tumors as well as act as tumor initiators 

and promoters (Habs et al. 1984; Warshawsky and Barkley 1987; Wynder and Hoffmann 1959b). 

Therefore, additional studies on the carcinogenicity of PAHs in animals are probably not necessary at 

this time. 

Genotoxicity. The genotoxic potential of several of the PAHs (both altemant and nonaltemant) has 

been extensively investigated using both in vivo and in vitro assays. All but three of the PAHs 

(acenaphthene, acenaphthylene, and fluorene) were reported to be mutagenic in at least one in vitro 

assay with the bacteria S. typhimurium. No further genotoxicity data are considered necessary at this 

time. 

Reproductive Toxicity. No data were located regarding reproductive effects of PAHs in humans, 

and the available information regarding reproductive effects of PAHs in animals is limited; data exist 

on only one of the PAHs (benzo[a]pyrene), and these data are conflicting. Adverse effects such as 

decreased fertility and total sterility in F1 progeny of CD-1 mice (Mackenzie and Angevine 1981) and 
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decreased incidence of pregnant female rats at parturition (Rigdon and Rennels 1964) were reported 

following oral exposure to benzo[a]pyrene. However, no adverse reproductive effects were observed 

in Swiss mice fed benzo[a]pyrene in their diet (Rigdon and Neal 1965). The metabolic differences 

and method of benzo[a]pyrene administration could account for the differential response to 

benzo[a]pyrene-induced toxicity in these studies. Parenteral studies in animals have also demonstrated 

adverse reproductive effects (Bui et al. 1986; Cervello et al. 1992; Mattison et al. 1992; Miller et al. 

1992; Swartz and Mattison 1985). The limited animal data suggest that PAHs may be reproductive 

toxicants, but these data are not extensive enough to draw firm conclusions. Furthermore, the testes 

and ovaries contain rapidly proliferating cells and therefore should be considered susceptible to 

damage by PAHs. The 90-day studies identified above should be conducted with special emphasis on 

reproductive organ pathology. If reproductive effects are observed in these studies, multigeneration 

animal studies could then be conducted to evaluate properly the relevance of this end point. In 

addition, since PAHs rarely occur alone, but are found in the environment in combinations and 

mixtures, studies, particularly feeding studies, addressing mixtures of PAHs should be conducted. In 

all cases, research should be conducted to determine the most appropriate animal species for 

extrapolation to humans. Future epidemiological studies should give special emphasis to evaluation of 

reproductive toxicity. 

Developmental Toxicity. No studies were located regarding developmental effects in humans 

exposed to PAHs by any route. However, results of in vitro studies suggest that human placental 

endocrine and hormonal function may be affected by exposure to benzo[a]pyrene (Avigdor et al. 1992; 

Bamea and Shurtz-Swirski 1992; Guyda 1991). Only limited data are available in animals on a few 

PAHs (mostly benzo[a]pyrene). These data indicate that ingested or parenterally administered PAHs 

have a potential to induce adverse developmental effects such as resorptions and malformations 

(Legraverend et al. 1984; Shum et al. 1979), sterility in F1 progeny (Mackenzie and Angevine 1981), 

testicular changes including atrophy of seminiferous tubules with lack of spermatids and spermatozoa, 

interstitial cell tumors (Payne 1958), immunosuppression (Urso and Gengozian 1980), and tumor 

induction (Bulay and Wattenberg 1971; Soyka 1980). However, another study found no 

developmental effects when benzo[a]pyrene was administered orally to mice (Rigdon and Neal 1965). 

The available animal data suggest that PAHs may be developmental toxicants. However, most of the 

data are from parenteral routes of exposure, and there are no inhalation data. The oral data are limited 

because of conflicting results across studies, the use of inconsistent protocols (e.g., varying numbers of 

animals, administration of the test compound during different times of gestation), the use of only one 
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dose, lack of study details, and the fact that data are available only on benzo[a]pyrene. Furthermore, 

some studies have shown that the toxic manifestations of benzo[a]pyrene are dependent on the route of 

exposure. Therefore, a two-species developmental toxicity study would be helpful to assess fully the 

potential for PAHs to affect development in humans. The route of exposure should be determined 

following evaluation of the reproductive organs in the 90-day studies to see if any particular route of 

exposure has a greater effect. The pharmacokinetic data on PAHs are insufficient,to determine 

whether similar effects may be expected to occur across different routes of exposure. Developmental 

toxicity should also be assessed in future animal reproductive toxicology testing and human 

epidemiological studies. In addition, since PAHs rarely occur alone, but are found in the environment 

in combinations and mixtures, studies, particularly feeding studies, addressing mixtures of PAHs 

should be conducted. In all cases, research should be conducted to determine the most appropriate 

animal species for extrapolation to humans. 

Immunotoxicity.No studies were located regarding immunological effects in humans after 

exposure to PAHs by any route, or in animals following inhalation exposure. In the one oral exposure 

study in animals that was located, a single dose of fluorene failed to affect thymus or spleen weight 

(Danz and Brauer 1988). However, there is information available in animals on the immunotoxicity of 

several PAHs following dermal exposure (contact hypersensitivity) (Old et al. 1963) and 

intraperitoneal or subcutaneous administration (suppression of both humoral and cellular immunity) 

(Blanton et al. 1986, 1988; Lubet et al. 1984; Lyte and Bick 1985; White and Holsapple 1984). In 

general, the degree of immunosuppression correlates with the individual PAH’s carcinogenic potency. 

Because of the information in animals that suggests that PAHs may affect the immune system, Tier I 

testing to assess PAH-induced immunotoxicity, as recently defined by the NTP (Luster et al. 1988) is 

recommended. The parameters that should be measured include immunopathology, humoral-mediated 

immunity, cell-mediated immunity, and nonspecific immunity. Although relatively high doses of 

PAHs must be used to obtain immunotoxicity in animals, much information could be gained from 

these studies. In addition, since PAHs rarely occur alone, but are found in the environment in 

combinations and mixtures, studies, particularly feeding studies, addressing mixtures of PAHs should 

be conducted. In all cases, research should be conducted to determine the most appropriate animal 

species for extrapolation to humans. Future epidemiologic studies should also place emphasis on 

evaluation of this end point. 
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Neurotoxicity. The potential for short- or long-term neurotoxic effects following exposure to 

PAHs by any route has not been specifically studied in humans or animals. Although acute-, 

intermediate-, and chronic-duration studies conducted in animals do not indicate that any of the PAHs 

tested showed gross evidence of neurotoxicity, these tests were not designed to detect subtle 

neurological changes. It is recommended that neurobehavioral as well as neuropathological end points 

be included in future 90-day toxicity testing of PAHs. If these preliminary data indicate that any of 

the PAHs are neurotoxicants, then a more comprehensive neurotoxicity battery, using sensitive 

functional and neuropathological tests, could be conducted to characterize further the neurotoxic 

potential of these PAHs. In addition, since PAHs rarely occur alone, but are found in the environment 

in combinations and mixtures, studies, particularly feeding studies, addressing mixtures of PAHs 

should be conducted. In all cases, research should be conducted to determine the most appropriate 

animal species for extrapolation to humans. 

Epidemiological and Human Dosimetry Studies. There are no epidemiological studies 

available that have investigated the effects of single PAHs by any route of exposure. Most of the 

available information on the effects of PAHs in humans comes from reports of occupational exposures 

to PAH-containing mixtures. For example, epidemiologic studies have shown increased mortality due 

to lung cancer in humans exposed to coke oven emissions (Lloyd 1971; Mazumdar et al. 1975; 

Redmond et al. 1976), roofing-tar emissions (Hammond et al. 1976), and cigarette smoke (Maclure 

and MacMahon 1980; Wynder and Hoffmann 1967). Each of these mixtures contains benzo[a]pyrene, 

chrysene, benz[a]anthracene, benzo[b]fluoranthene, and dibenz[a,h]anthracene as well as other 

potentially carcinogenic PAHs and other carcinogenic and potentially carcinogenic chemicals, tumor 

promoters, initiators, and co-carcinogens such as nitrosamines, coal tar pitch, and creosote. 

Limitations inherent in these studies include unquantified exposure concentrations and durations, as 

well as concomitant exposure to other potentially toxic substances. Despite their inadequacies, studies 

in humans suggest that PAH-containing mixtures are dermal irritants and carcinogens following 

inhalation and/or derrnal exposure. If either worker or general populations with appropriate exposure 

can be identified, epidemiologic studies should be undertaken with special emphasis placed upon 

evaluation of cancer (of the skin and other organs) and other adverse skin effects, reproductive/ 

developmental toxicity, and immunotoxicity. However, such studies would be difficult to conduct. 

With a group of chemicals that are as ubiquitous as PAHs, it would be extremely difficult to 

distinguish between exposed and nonexposed populations. The more these groups overlap, the higher 

the chance for misclassification bias. In addition, the statistical power of an epidemiological study 
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depends partially on the variance of the exposure measurements. If there is enormous variation in the 

exposure levels among the exposed and nonexposed groups, then the population size needed to obtain 

statistical significance in the study would be unmanageable and would most likely not be found in any 

one occupational setting or hazardous waste site. Furthermore, because of the size of the population 

needed, it would be very difficult to control for confounding factors such as smoking, geographical 

location, lifestyle. 

Biomarkers of Exposure and Effect. 

Exposure. Sensitive analytical methods are available to quantify PAH exposure in humans. Although 

PAHs can be detected in the body fluids and tissues, because of the ubiquitous nature of PAHs in the 

environment, these biomarkers are not specific for any particular source of PAH exposure. PAHs and 

their metabolites (e.g., 1-hydroxypyrene) can be measured in the urine of exposed individuals. PAHs 

form DNA adducts that can be measured in body tissues or blood following exposure to PAHs and 

mixtures that contain PAHs. Studies attempting to identify suitable and reliable biomarkers from 

phenanthrene, chrysene, and fluoranthene have been conducted (e.g., Grimmer et al. 1988). However, 

no other biomarkers (specific or otherwise) that have practical utility have been identified following 

exposure to PAHs. Further work on developing biomarkers that enable exposure to be quantified 

would be useful to ascertain whether individuals have been exposed to potentially toxic levels of 

PAHs. In addition, since PAHs rarely occur alone, but are found in the environment in combinations 

and mixtures, studies addressing mixtures of PAHs should be conducted. 

Effect. The available biomarkers of effect for PAHs are not specific for effects induced by PAHs. 

The available genotoxicity data indicate that several of the 17 PAHs considered here are genotoxic in 

both nonmammalian and mammalian systems and are indirect mutagens (i.e., requiring the presence of 

an exogenous mammalian metabolic system). There were no tests reported for humans exposed to 

benzo[a]pyrene (the most widely tested PAH) in vivo, but several types of cultured human tissue cells 

demonstrated positive results for benzo[a]pyrene-induced genotoxicity (as evidenced by the induction 

of chromosomal aberrations, sister chromatid exchange) and binding of benzo[a]pyrene to DNA. 

Thus, although these results are exclusively from in vitro tests and the limited genotoxicity tests 

conducted on urine obtained from humans exposed to PAHs have been negative, these genotoxic 

effects observed in human tissue cells may serve as a biomarker of effects for at least one of the 

PAHs, benzo[a]pyrene. The formation of benzo[a]pyrene-DNA adducts has been demonstrated, and 
PAHs 203 

2. HEALTH EFFECTS 

this may serve as a biomarker of PAH-induced carcinogenicity. Additional studies on the relative 

sensitivity of DNA adducts and sister chromatid exchanges to identify threshold levels of exposure that 

could be detected in human populations would be useful. In addition, since PAHs rarely occur alone, 

but are found in the environment in combinations and mixtures, studies addressing mixtures of PAHs 

should be conducted. 

Absorption, Distribution, Metabolism, and Excretion. The quantitative data on the 

toxicokinetics of PAHs are based, to a large extent, on short-term exposure to benzo[a]pyrene in 

animals. Occupational exposure to PAHs generally occurs as a mixture. Therefore, inhalation, oral, 

and dermal studies exploring how PAHs interact with each other to affect their disposition would be 

more representative of exposures in humans. 

The presence of PAHs and their metabolites in human urine and blood following inhalation, oral, and 

dermal exposures indicates that PAH absorption occurs in humans (Becher and Bjorseth 1983; Buckley 

and Lioy 1992; Hecht et al. 1979). However, there was no quantitative information on the extent and 

rate of PAH absorption in humans. Most of the information regarding the pulmonary and oral 

absorption and distribution of PAHs in animals is based on acute-duration exposures (Chang 1943; 

Hecht et al. 1979; Weyand and Bevan 1986, 1987b, 1988; Withey et al. 1991; Wolff et al. 1989c). 

PAHs appear to be widely distributed in tissues of animals following oral and inhalation exposure; 

peak tissue concentrations occurred earlier with higher exposure levels. Studies on the absorption and 

distribution of PAHs following long-term exposures would indicate whether the kinetics are similar to 

acute-duration exposures. The dermal study conducted by Storer et al. (1984) revealed that several 

PAHs in a crude coal tar mixture were absorbed, but that benzo[a]pyrene was not. In contrast, animal 

studies indicate that benzo[a]pyrene was dermally absorbed (Ng et al. 1992; Wester et al. 1990; Yang 

et al. 1989); however, tissue distribution was not discussed. 

PAH metabolism has been extensively reviewed in human and animal tissue homogenates, cultures, 

and perfused systems (Autrup et al. 1978; Cavalieri et al. 1987; Cohen et al. 1976; Kiefer et al. 1988; 

Leung et al. 1988). However, these studies are limited to the biotransformation of individual 

compounds. Since most metabolic pathways have been identified or can be predicted for the 

individual PAHs, it is now important to understand how these metabolic pathways are affected when 

the PAHs compete. The carcinogenic and toxic potential of PAHs is associated with their metabolites. 

Alterations in rates of metabolism and metabolite profiles may affect the toxic consequences of PAHs. 
PAHs 204 

2. HEALTH EFFECTS 

Available data on several of the nonaltemant tumorigenic PAHs discussed in this profile indicate that 

they exert their adverse effects by mechanisms that differ from those that have been more recently 

elucidated for altemant PAHs (Amin et al. 1982, 1985b; Rice et al. 1987b). The mechanisms by 

which benzo[b]fluoranthene and benzo[j]fluoranthene are metabolically activated to genotoxic agents 

have been elucidated (LaVoie et al. 1993b; Marshall et al. 1993; Weyand et al. 1993a, 1993b). 

Additional studies designed to assess the potential toxic effects of these reactive metabolites in various 

species and at various organ sites would be useful. 

No studies were located that monitored the rate and extent of PAH excretion in humans. Most studies 

in animals concentrated on the extent of PAH excretion and the distribution of the compound and its 

metabolites in urine, feces, and bile following short-term exposures (Bevan and Weyand 1988; 

Grimmer et al. 1988; Petridou-Fischer et al. 1988; Weyand and Bevan 1986; Wolff et al. 1989c; 

Yamazaki and Kakiuchi 1989). Data regarding the excretion pattern and rate following long-term 

exposure to PAHs would be useful to determine if bioaccumulation occurs. 

In addition, since PAHs rarely occur alone, but are found in the environment in combinations and 

mixtures, studies, particularly feeding studies, addressing mixtures of PAHs should be conducted. In 

all cases, research should be conducted to determine the most appropriate animal species for 

extrapolation to humans. 

Comparative Toxicokinetics. Occupational studies provide evidence that inhaled PAHs are 

absorbed. Animal studies also show that pulmonary absorption of benzo[a]pyrene occurs, but the 

extent of absorption is not known. Ingestion of benzo[a]pyrene is low in humans while oral 

absorption in animals varies among the PAH compounds depending on the lipophilicity. The 

absorption and distribution of PAHs in various species would be expected to be similar based on the 

lipophilicity of the compounds. In general, percutaneous absorption of PAHs in several animal species 

appears to be rapid and high (Ng et al. 1992; Sanders et al. 1986; Wester et al. 1990; Yang et al. 

1989). This suggests that dermal absorption in humans may also occur rapidly; however, the extent of 

absorption may vary depending on the vehicle. 

There was no information available on the distribution of PAHs in humans. In general, tissue 

distribution of benzo[a]pyrene following inhalation exposure is qualitatively similar for different 

species (Bevan and Weyand 1988; Weyand and Bevan 1986, 1987a, 1988; Wolff et al. 1989c). In 
PAHs 205 

2. HEALTH EFFECTS 

general, orally absorbed PAHs were rapidly and widely distributed in the rat (Bartosek et al. 1984; 

Withey et al. 1991; Yamazaki and Kakiuchi 1989). Qualitative similarities in distribution among 

species suggest that distribution in humans would also be similar. Placental transfer of PAHs in mice 

and rats appears to be limited (Neubert and Tapken 1988; Withey et al. 1992); therefore, human 

fetuses may be exposed to PAHs, but levels would not be as high as maternal levels. 

Qualitatively, metabolism and excretion would be relatively similar in humans and animals, but 

variability in specific activities of enzymes will alter the metabolic profiles among the species. 

Knowledge of these differences in enzyme activity in various species would assist in predicting which 

pathways and metabolites would prevail. For instance, AHH activity is not induced by PAHs in some 

strains of mice. Therefore, it would be useful to examine the metabolism of those less-well-studied 

PAHs in several species (i.e., rodent and nonrodent) so that the carcinogenic potential of PAHs in 

various species could be predicted. The feces (via the bile) appears to be the major excretion route, 

but the extent of elimination of PAHs varies among species (Bevan and Weyand 1988; Grimmer et al. 

1988; Ng et al. 1992; Petridou-Fischer et al. 1988; Sanders et al. 1986; Weyand and Bevan 1986; 

Wolff et al. 1989c). Further comparative studies on excretion would be useful because differences in 

human and animal excretion rates are not known. In addition, many of the toxicity tests have used 

mice, while a larger proportion of toxicokinetic studies have used rats. Thus, more kinetic studies 

should be conducted in mice to provide data to correspond to the toxicity data. Further, since PAHs 

rarely occur alone, but are found in the environment in combinations and mixtures, studies, particularly 

feeding studies, addressing mixtures of PAHs should be conducted. In all cases, research should be 

conducted to determine the most appropriate animal species for extrapolation to humans. 

Methods for Reducing Toxic Effects. Efforts are currently underway to develop ways to 

mitigate the adverse effects of PAHs, especially with regard to natural products. Efforts to reduce or 

eliminate cigarette smoking in the general population also contribute toward reducing exposure to and 

toxic effects of PAHs. The target organs of PAHs have been identified (i.e., the skin and rapidly 

proliferating tissue such as the hematopoietic and lymphoid systems). Furthermore, several PAHs are 

considered to be carcinogenic. The mechanism of action for altemant PAH-induced carcinogenicity is 

fairly well understood. However, additional information would be useful to understand the mechanism 

of nonaltemant PAH-induced carcinogenicity, how PAHs exert their adverse effects on rapidly 

proliferating tissue, and how various interactions between PAHs can affect their toxicity and 

carcinogenicity. 
PAHs 206 

2. HEALTH EFFECTS 

2.9.3 Ongoing Studies
 

Ongoing research on the health effects and toxicokinetics of PAHs is summarized in Table 2-7.
 


PAHs 209 

3. CHEMICAL AND PHYSICAL INFORMATION 

3.1 CHEMICAL IDENTITY
 

Information regarding the chemical identity of PAHs is located in Table 3-l.
 

3.2 PHYSICAL AND CHEMICAL PROPERTIES
 

Information regarding the physical and chemical properties of PAHs is located in Table 3-2.
 













PAHs 223 

4. PRODUCTION, IMPORT/EXPORT, USE, AND DISPOSAL 

4.1 PRODUCTION 

The commercial production of PAHs is not a significant source of these compounds in the 

environment. The primary source of many PAHs in air is the incomplete combustion of wood and 

fuel (Perwak et al. 1982). PAHs are a ubiquitous product of combustion from common sources such 

as motor vehicles and other gas-burning engines, wood-burning stoves and furnaces, cigarette smoke, 

industrial smoke or soot, and charcoal-broiled foods (IARC 1983). Natural sources include volcanoes, 

forest fires, crude oil, and shale oil (HSDB 1994). 

Of the 17 PAHs included in this profile, only three are produced commercially in the United States in 

quantities greater than research level: acenaphthene, acenaphthylene, and anthracene. Acenaphthene is 

manufactured by passing ethylene and benzene or naphthalene through a red hot tube or by heating 

tetrahydroacenaphthene with sulfur to 180 oC. It can also be made from acenaphthenone or 

acenaphthenequinone by high-pressure hydrogenation in decalin with nickel at 180-240 oC 

(Windholz 1983). Another manufacturing process involves the isolation and recovery of acenaphthene 

from a concentrated tar-distillation fraction (Grayson 1978). Technical grades of acenaphthene are 

typically 98% pure (HSDB 1994). Acenaphthylene is produced by catalytic degradation of 

acenaphthene (Grayson 1978). Toxic Release Chemical Inventory (TRI) production data for 

acenaphthylene and acenaphthene are not available (TRI92 1994) and no other data on the production 

volumes for these compounds in the United States could be found. 

Anthracene is produced commercially by recovery from the coal tar distillation fraction known as 

“anthracene oil” or “green oil.” Purification techniques, including heating and vacuum distillation, are 

required to remove the major contaminant, potassium carbazole (IARC 1985). Zone melting of solid 

anthracene and crystallization from benzene followed by sublimation are also effective purification 

techniques (Hampel and Hawley 1973). Technical grades of anthracene are typically 90-98% pure 

(HSDB 1994). Table 4-l shows the number of facilities per state that manufacture or process 

anthracene, as well as a range of the maximum amounts of anthracene present at the facilities 

(TR192 1994). The following companies have been cited as current U.S. manufacturers of anthracene 

for sale/distribution: ABC Coke Division (Tarrant City, Alabama); Granite City Steel (Granite City, 

Illinois); Citizens Gas and Coke Utility (Indianapolis, Indiana); National Steel Corporation (Encore, 

PAHs 225 

4. PRODUCTION, IMPORT/EXPORT, USE, AND DISPOSAL 

Missouri); Koch Refining Company (Pine Bend, Montana); Amerada Hess Corporation (Port Reading, 

New Jersey); Reilly Industries, Incorporated (Cleveland, Ohio, and Lone Star, Texas); New Boston 

Coke Corporation (New Boston, Ohio); Bethlehem Steel Structural (Bethlehem, Pennsylvania); Geneva 

Steel (Vineyard, Utah); and Hess Oil Virgin Islands Corporation (St. Croix, Virgin Islands). It should 

be noted that another source (SRI 1994) lists no anthracene processing or manufacturing facilities. In 

1982, more than 2,270 kg of anthracene were produced in the United States; more recent production 

data are not available (HSDB 1994). 

The following compounds are not produced commercially in the United States: benz[a]anthracene, 

benzo[b]fluoranthene, benzo[e]pyrene, benzo[j]fluoranthene, benzo[k]fluoranthene, 

benzo[g,h,i]perylene, benzo[a]pyrene, chrysene, dibenz[a,h]anthracene, fluoranthene, fluorene, 

indeno[l,2,3-c,d]pyrene, phenanthrene, and pyrene (HSDB 1994; IARC 1985). 

4.2 IMPORT/EXPORT 

The most recent data available on U.S. import and export volumes of individual PAHs are as follows: 

in 1986, 4,000 kg of acenaphthene were imported into the United States; in 1985, 5,730,000 kg of 

anthracene oil, 882,000 kg of anthracene (≥ 30% purity by weight), 1,040 kg of fluoranthene, and 

57,400 kg of pyrene were imported into the United States; in 1984, 79,200 kg of chrysene, 9,440 kg 

of fluoranthene, 9.1 kg of fluorene, and 551 kg of phenanthrene were imported into the United States; 

in 1984, 502,000 kg of anthracene oil and pitch of tar coke were exported from the United States, 

increasing to 3,890,000 kg by 1987 (HSDB 1994; USDOC 1985). In 1985, the United States imported 

a total of almost 12 million gallons of creosote oil from the Netherlands, France, West Germany, and 

other countries and almost 185 million pounds of coal tar pitch, blast furnace tar, and oil-gas tar from 

Canada, Mexico, West Germany, Australia, and other countries (USDOC 1985). The only relevant 

information found on more recent import and export volumes is for the following group of compounds 

which contains two PAHs: acenaphthene, chrysene, cymene, and indene (NTDB 1994). Annual 

import volumes for this group of compounds were 2x106

, 3x106, and 9x105 kg for the years 1991, 

1992, and 1993, respectively. While these numbers set an upper limit on the import volume for any 

one of the compounds in this group, they do not provide any further information on import volumes of 

acenaphthene and chrysene. 
PAHs 226 

4. PRODUCTION, IMPORT/EXPORT, USE, AND DISPOSAL 

4.3 USE 

There is no known use for acenaphthylene, benz[a]anthracene, benzo[b]fluoranthene, benzo[e]pyrene, 

benzo[j]fluoranthene, benzo[k]fluoranthene, benzo[g,h,i]perylene, benzo[a]pyrene, chrysene, 

dibenz[a,h]anthracene, indeno[ 1,2,3-c,d]pyrene, or pyrene except as research chemicals (Hawley 1987; 

HSDB 1994). 

Anthracene is used as an intermediate in dye production, in the manufacture of synthetic fibers, and as 

a diluent for wood preservatives. It is also used in smoke screens, as scintillation counter crystals, and 

in organic semiconductor research (Hawley 1987). Anthracene is used to synthesize the 

chemotherapeutic agent, Amsacrine (Wadler et al. 1986). Acenaphthene is used as a dye intermediate, 

in the manufacture of pharmaceuticals and plastics, and as an insecticide and fungicide (HSDB 1994; 

Windholz 1983). 

Fluorene is used as a chemical intermediate in many chemical processes, in the formation of 

polyradicals for resins, and in the manufacture of dyestuffs (Hawley 1993; HSDB 1994). 

Phenanthrene is used in the manufacture of dyestuffs and explosives and in biological research 

(Hawley 1987; HSDB 1994). Fluoranthene is used as a lining material to protect the interior of steel 

and ductile-iron drinking water pipes and storage tanks (NRC 1983). 

4.4 DISPOSAL 

PAHs serve as the basis for listing certain hazardous wastes under the Resource Conservation and 

Recovery Act (RCRA); they are listed as constituents for groundwater monitoring and are monitored in 

hazardous wastes as part of the RCRA land disposal restrictions (EPA 1989c). Specific regulations 

governing the generation, treatment, storage and disposal of hazardous wastes containing PAHs are 

listed in Chapter 7. 

Approximately two-thirds of the PAHs in surface waters are particle-bound and can be removed by 

sedimentation, flocculation, and filtration processes. The remaining one-third of the dissolved PAHs 

usually require oxidation for partial removal/transformation (EPA 1980). 
PAHs 227 

4. PRODUCTION, IMPORT/EXPORT, USE, AND DISPOSAL 

Acenaphthene, acenaphthylene, benz[a]anthracene; benzo[a]pyrene, benzo[b]fluoranthene, chrysene, 

dibenz[a,h]anthracene, indeno[ 1,2,3-c,d]pyrene, and fluoranthene are all good candidates for rotary kiln 

incineration at temperatures ranging from 820 to 1,600 oC and residence times of seconds for liquids 

and gases, and hours for solids (EPA 1981a; HSDB 1994). Benz[a]anthracene, benzo[a]pyrene, 

chrysene, dibenz[a,h]anthracene, indeno[ 1,2,3-c,d]pyrene, and fluoranthene all are good candidates for 

fluidized-bed incineration at a temperature of 450-980 oC and residence times of seconds for liquids 

and gases and longer for solids (EPA 1981a; HSDB 1994). Benz[a]anthracene also is a good 

candidate for liquid injection incineration at a temperature range of 650-1,600 oC and a residence time 

of 0.1 to 2 seconds (EPA 1981a; HSDB 1994). Liquids containing pyrene should be atomized in an 

incinerator. Combustion is improved by mixing with a more flammable solvent. Solids should be 

combined with paper or other flammable material prior to incineration (UN 1985). Benz[a]anthracene, 

chrysene, dibenz[a,h]anthracene, benzo[k]fluoranthene, benzo[j]fluoranthene, benzo[b]fluoranthene, and 

benzo[a]pyrene laboratory wastes can be oxidized using agents such as concentrated sulfuric acid, 

potassium dichromate, or potassium permanganate (IARC 1985). Water contaminated with 

benzo[g,h,i]perylene can be decontaminated by carbon adsorption (EPA 1981c). Anthracene in waste 

chemical streams may be subjected to ultimate disposal by controlled incineration (HSDB 1994; EPA 

1981a). 

Bioremediation is emerging as a practical alternative to traditional disposal techniques (Cemiglia 1993; 

Thomas and Lester 1993; Wilson and Jones 1993). It has only recently been considered as a viable 

treatment method for contaminated soils but is now being used or is under consideration by the EPA 

for clean-up in over 135 Super-fund and underground storage tank sites (Carraway and Doyle 1991; 

Sims 1990; EPA 1989d). In situ treatment involves addition of nutrients, an oxygen source, and, 

sometimes, specifically adapted microorganisms that enhance degradation. Current in situ treatments 

have been used with some success for removal of two- and three-ring PAHs but generally are 

considered ineffective for removal of most PAHs from soil (Wilson and Jones 1993). 

On-site methods such as landfarming also have been used successfully to degrade PAHs with three or 

fewer aromatic rings. The waste material is applied to the soil as a slurry and the area is fertilized, 

irrigated, limed, and tilled. The major disadvantage of landfarming is that contaminants can move 

from the treatment area. To enhance treatment and minimize movement of contaminants, prepared 

beds have been used. For this type of remediation, the contaminated soil is removed to a specially 
PAHs 228 

4. PRODUCTION, IMPORT/EXPORT, USE, AND DISPOSAL 

prepared area lined with a low permeability material and the bed is managed to optimize degradation 

(Wilson and Jones 1993). 

A third type of bioremediation involves the use of a bioreactor in a dedicated treatment area. The 

contaminated soil is excavated, slurried with water, and treated in the reactor. The horizontal drum 

and airlift-type reactors are usually operated in the batch mode but may also be operated in a 

continuous mode. Because there is considerable control over the operating conditions, treatment often 

is quick and effective. Contaminated groundwater and effluent also may be treated in either fixed-film’ 

or stirred-tank bioreactors. However, bioreactors are still in the developmental stages and further 

research is required to optimize their efficiency and cost effectiveness (Wilson and Jones 1993). 

A pilot-scale evaluation of the bioreactor method was carried out in a joint Superfund Innovative 

Technology Evaluation (SITE) project and a project to collect information for the Best Demonstrated 

Available Technologies (BDAT) database (Lewis 1993). Five 64-L bioslurry reactors were charged 

with a 30% (wt/vol) ratio of creosote-contaminated soil from a Superfund site, inoculated with PAH 

degraders, and inorganic nutrients were added. Total PAH -degradation averaged 93.4 ± 3.2% over all 

reactors during the 12-week study, with 97.4% degradation of the 2- and 3-ring PAHs and 90% 

degradation of the 4- to 6-ring PAHs. 

Huesemann et al. (1993) carried out a 16-week laboratory study to assess the biotreatability of PAHs 

in refinery American Petroleum Institute (API) oil separator sludge. Two biotic treatments were 

evaluated: (1) a nutrient-amended, inoculated, aerated slurry reactor, and (2) an oxygen-sparged 

reactor. A sterile, nitrogen-sparged reactor was used as a control. Naphthalene, anthracene, 

phenanthrene, and benzo[a]pyrene were completely biodegraded in 4 weeks in both biotic reactors. 

Chrysene biodegraded in 4 weeks in the aerated reactor and in 16 weeks in the oxygen-sparged 

reactor. No significant changes in pyrene concentration were observed in the oxygen-sparged reactor 

and only a 30% degradation was observed in the aerated reactor. The authors concluded that aerobic 

biotreatment was successful in removing most BDAT PAHs from refinery API oil separator sludge. 
PAHs 229 

5. POTENTIAL FOR HUMAN EXPOSURE 

5.1 OVERVIEW 

This chapter provides a discussion of the environmental fate and potential for human exposure to 

17 PAHs. For the purposes of describing environmental fate, these PAHs have been grouped into low, 

medium, and high molecular weight classes (see Section 5.3.1, Transport and Partitioning). In general, 

chemicals within each class have similar environmental fates. When available, data are provided for 

the individual PAHs that are the subject of the profile. When data on each compound are not 

available, data on members of the weight class are provided. Data regarding total PAHs or 

generalizations about PAHs are also used to provide insight into the behavior of the compounds 

cbvered in this profile. 

PAHs are released to the environment through natural and synthetic sources with emissions largely to 

the atmosphere. Natural sources include emissions from volcanoes and forest fires. Synthetic sources 

provide a much greater release volume than natural sources; the largest single source is the burning of 

wood in homes. Automobile and truck emissions are also major sources of PAHs. Environmental 

tobacco smoke, unvented radiant and convective kerosene space heaters, and gas cooking and heating 

appliances may be significant sources of PAHs in indoor air. Hazardous waste sites can be a 

concentrated sources of PAHs on a local scale. Examples of such sites are abandoned wood-treatment 

plants (sources of creosote) and former manufactured-gas sites (sources of coal tar). PAHs can enter 

surface water through atmospheric deposition and from discharges of industrial effluents (including 

wood-treatment plants), municipal waste water, and improper disposal of used motor oil. Several of 

the PAHs have been detected at hazardous waste sites at elevated levels. In air, PAHs are found 

sorbed to particulates and as gases. Particle-bound PAHs can be transported long distances and are 

removed from the atmosphere through precipitation and dry deposition. PAHs are transported from 

surface waters by volatilization and sorption to settling particles. The compounds are transformed in 

surface waters by photooxidation, chemical oxidation, and microbial metabolism. In soil and 

sediments, microbial metabolism is the major process for degradation of PAHs. Although PAHs are 

accumulated in terrestrial and aquatic plants, fish, and invertebrates, many animals are able to 

metabolize and eliminate these compounds. Bioconcentration factors (BCFs), which express the 

concentration in tissues compared to concentration in media, for fish and crustaceans are frequently in 
PAHs 230 

5. POTENTIAL FOR HUMAN EXPOSURE 

the 10-10,000 range. Food chain uptake does not appear to be a major source of exposure to PAHs 

for aquatic animals. 

The greatest sources of exposure to PAHs for most of the United States population are active or 

passive inhalation of the compounds in tobacco smoke, wood smoke, and contaminated air, and 

ingestion of the compounds in foodstuffs. The general population may also be exposed to PAHs in 

drinking water and through skin contact with soot and tars. Higher than background levels of PAHs 

are found in foods that are grilled or smoked. Estimates of human exposures to PAHs vary. The 

average total daily intake of PAHs by a member of the general population has been estimated to be 

0.207 μg from air, 0.027 μg from water, and 0.16-l .6 μg from food. The total potential exposure to 

carcinogenic PAHs for adult males in the United States was estimated to be 3 μg/day. Smokers of 

unfiltered cigarettes may experience exposures twice as high as these estimates. Persons living in the 

vicinity of hazardous waste sites where PAHs above background levels have been detected may also 

be exposed to higher levels. 

PAHs have been identified in at least 600 of the 1,408 hazardous waste sites that have been proposed 

for inclusion in the EPA National Priorities List (NPL) (HazDat 1994). However, the number of sites 

evaluated for PAHs is not known. The frequencies of these sites can be seen in Figure 5-1. 

5.2 RELEASES TO THE ENVIRONMENT 

5.2.1 Air 

Most of the direct releases of PAHs to the environment are to the atmosphere from both natural and 

anthropogenic sources, with emissions from human activities predominating. PAHs in the atmosphere 

are mostly associated with particulate matter; however, the compounds are also found in the gaseous 

phase (NRC 1983; Yang et al. 1991). The primary natural sources of airborne PAHs are forest fires 

and volcanoes (Baek et al. 1991; NRC 1983). The residential burning of wood is the largest source of 

atmospheric PAHs (Peters et al. 1991; Ramdahl et al. 1982); releases are primarily the result of 

inefficient combustion and uncontrolled emissions (Freeman and Cattell 1990; NRC 1983; Tan et al. 

1992). Other important stationary anthropogenic sources include industrial power generation, 

incineration (Shane et al. 1990; Wild et al. 1992); the production of coal tar, coke, and asphalt; and 

petroleum catalytic cracking (Baek et al. 1991; Guerin 1978; Perwak et al. 1982; Santodonato 1981). 

PAHs 232 

5. POTENTIAL FOR HUMAN EXPOSURE 

Environmental tobacco smoke, unvented radiant and convective kerosene space heaters, and gas 

cooking and heating appliances may be important sources of PAHs in indoor air (Chuang et al. 1991; 

Hoffmann and Hoffmann 1993; Mumford et al. 1991; NRC 1986; Traynor et al. 1990). Stationary 

sources account for about 80% of total annual PAH emissions; the rest are from mobile sources. The 

most important mobile sources of PAHs are vehicular exhaust from gasoline and diesel-powered 

engines (Baek et al. 1991; Johnson 1988; Yang et al. 1991). Mobile sources are often the major 

contributors to PAH releases to the atmosphere in urban or suburban areas (Baek et al. 1991). The 

amount of anthracene released to the atmosphere in 1992 by U.S. industrial facilities sorted by state is 

given in Table 5-l (TRI92 1994). The TRI data should be used with caution since only certain types 

of facilities are required to report. This is not an exhaustive list. TRI92 (1994) data were not 

available for other PAHs included in this profile. 

The U.S. annual emissions (from early to mid 1970s) of polycyclic organic matter (a term generally 

used to describe PAHs, their nitrogen-containing analogs, and their quinone degradation products 

[Santodonato et al. 1981]) were estimated by NRC (1983) as follows: open buming 4,024 metric 

tons (39%), residential heating-3,956 metric tons (38%), automobiles and trucks-2,266 metric tons 

(22%), and industrial boilers-74 metric tons (1%). NRC (1983) estimated that the total amount of 

benzo[a]pyrene produced in the United States is between 300 and 1,300 metric tons annually. Peters 

et al. (1981) estimated that a total of 11,031 metric tons of PAHs were released to the atmosphere in 

the United States on an annual basis, with 36% of the total coming from residential heating, 6% from 

industrial processes, 1% from incineration, 36% from open burning, 1% from power generation, and 

21% from mobile sources. This estimate can be compared to that of Ramdahl et al. (1982), who 

reported that a total of 8,598 tons of PAHs were emitted to the atmosphere annually from the 

following sources: (1) residential heating-16%; (2) industrial processes 41%; (3) incineration-l % 

(4) open burning-13%; (5) power generation-5%; and (6) mobile sources-25%. 

The composition of PAH emissions varies with the combustion source. For example, emissions from 

residential wood combustion contain more acenaphthylene than other PAHs (Perwak et al. 1982), 

whereas auto emissions contain more benzo[g,h,i]perylene and pyrene (Rogge et al. 1993a; 

Santodonato et al. 1981). PAHs in diesel exhaust particulates are dominated by three- and four-ring 

compounds, primarily fluoranthene, phenanthrene, and pyrene (KelIy et al. 1993; Rogge et al. 1993a; 

Westerholm and Li 1994). Diesel exhaust vapor emissions are dominated by phenanthrene and 

anthracene (Westerholm and Li 1994). Acenaphthene, fluorene, and phenanthrene have been found to 

PAHs 234 

5. POTENTIAL FOR HUMAN EXPOSURE 

be predominant in total (particle- and vapor-phase) diesel emissions (Lowenthal et al. 1994). 

Phenanthrene was the most abundant and frequently detected PAH in samples of fly ash and bottom 

ash collected from municipal refuse incinerators in the United States (Shane et al. 1990), whereas 

benzo[g,h,i]perylene was the most abundant and frequently detected PAH in fly ash samples collected 

from municipal solid waste incinerators in the United Kingdom (Wild et al. 1992). Fluoranthene, 

benzo[a]fluoranthene, benzo[g,h,i]perylene, indeno[ 1,2,3-c,d]pyrene, phenanthrene, and chrysene were 

predominant in emission particle samples collected from a municipal waste incinerator, whereas 

benzo[g,h,i]perylene and benz[a]anthracene were predominant in emission particle samples collected 

from a municipal and medical/pathological waste incinerator (Williams et al. 1994). Emission particle 

samples from a pilot scale rotary kiln incinerator charged with polyethylene contained predominantly 

benz[a]anthracene and phenanthrene when an afterburner was used, whereas pyrene, fluoranthene, and 

phenanthrene were predominant without an afterburner; total PAH concentrations were reduced by a 

factor greater than 100 by the use of an afterburner (Williams et al. 1994). In coal tar pitch emissions, 

concentrations of phenanthrene and pyrene have been reported to be 20-80 times greater than the 

concentrations of benzo[a]pyrene and benzo[g,h,i]perylene (Sawicki 1962). Chrysene/triphenylene, 

pyrene, and fluoranthene were dominant among the PAHs found in fine particle emissions from natural 

gas home appliances (Rogge et al. 1993b). Cigarette mainstream smoke contains a wide variety of 

PAHs with reported concentrations of benzo[a]pyrene ranging from approximately 5-80 ng/cigarette; 

sidestream smoke concentrations are significantly higher with sidestream/mainstream concentration 

ratios for benzo[a]pyrene ranging from 2.5 to 20 (Hoffmann and Hoffmann 1993; IARC 1983). 

5.2.2 Water 

Important sources of PAHs in surface waters include deposition of airborne PAHs (Jensen 1984), 

municipal waste water discharge (Barrick 1982), urban storm water runoff (MacKenzie and Hunter 

1979), runoff from coal storage areas (Stahl et al. 1984; Wachter and Blackwood 1979), effluents .from 

wood treatment plants and other industries (DeLeon et al. 1986; Snider and Manning 1982; USDA 

1980), oil spills (Giger and Blumer 1974), and petroleum pressing (Guerin 1978). Brown and Weiss 

(1978) estimated that l-2 tons of benzo[a]pyrene were released from municipal sewage effluents and 

0.1-0.4 tons of benzo[a]pyrene were released from petroleum refinery waste waters in the United 

States in 1977. 
PAHs 235 

5. POTENTIAL FOR HUMAN EXPOSURE 

Most of the PAHs in surface waters are believed to result from atmospheric deposition (Santodonato et 

al. 1981). However, for any given body of water, the major source of PAHs could vary. Jensen 

(1984) studied benzo[a]pyrene loading in a marine coastal area and determined that atmospheric 

deposition was indeed the major source of benzo[a]pyrene, with lesser amounts contributed by refinery 

effluent, municipal waste water, urban runoff, and rivers. Prahl et al. (1984) found that combustionderived PAHs adsorbed to suspended sediments in rivers accounted for the major portion of PAHs in 

the waters of a Washington coastal area, and other studies have identified industrial effluents, road 

runoff, and oil spills as the major contributors in specific bodies of water (DeLeon et al. 1986; 

Santodonato et al. 1981). 

The amount of anthracene released to surface water and publicly owned treatment works (POTWs) in 

1992 by U.S. industrial facilities sorted by state is shown in Table 5-l (TRI92 1994). The TRI data 

should be used with caution since only certain facilities are required to report. This is not an 

exhaustive list. TRI92 (1994) data were not available for other PAHs included in this profile. 

Because most of the PAHs released to aquatic environments tend to remain near the sites of 

deposition, lakes, rivers, estuaries, and coastal marine environments near centers of human populations 

and industrial activity tend to be the major repositories of aquatic PAHs (Neff 1979). 

5.2.3 Soil 

Most of the PAHs in soil are believed to result from atmospheric deposition after local and long-range 

transport. The presence of PAHs in the soil of regions remote from any industrial activity supports 

this contention (Thomas 1986). Other potential sources of PAHs in soil include sludge disposal from 

public sewage treatment plants, automotive exhaust, irrigation with coke oven effluent, leachate from 

bituminous coal storage sites, and use of soil compost and fertilizers (Perwak et al. 1982; Santodonato 

et al. 1981; Stahl et al. 1984; White and Lee 1980). The principal sources of PAHs in soils along 

highways and roads are vehicular exhausts and emissions from wearing of tires and asphalt. PAHs 

may also be released to soils at concentrations above background and landfill sites (Black et al. 1989) 

and industrial sites, including creosote production (Ellis et al. 1991), wood-preserving (Mueller et al. 

1991; Weissenfels et al. 1990), and coking plants (Weissenfels et al. 1990; Werner et al. 1988). Soils 

at the sites of former manufactured gas plants are also heavily contaminated with PAHs (Bewley et al. 

1989; Tumey and Goerlitz 1990). 
PAHs 236 

5. POTENTIAL FOR HUMAN EXPOSURE 

The amount of anthracene released to surface water and publicly owned treatment works (POTWs) in 

1992 by U.S. industrial facilities sorted by state is shown in Table 5-l (TRI92 1994). Based on data 

in Table 5-1, only relatively small amounts of anthracene were discharged in hazardous waste sites 

from U.S. industrial facilities in 1992. However, some of the anthracene wastes transferred off-site 

(see Table 5-l) ultimately may be disposed of on land. The TRI data should be used with caution 

since only certain facilities are required to report. This is not an exhaustive list. TRI92 (1994) data 

were not available for other PAHs included in this profile. 

5.3 ENVIRONMENTAL FATE 

5.3.1 Transport and Partitioning 

The global movement of PAHs can be summarized as follows: PAHs released to the atmosphere are 

subject to short- and long-range transport and are removed by wet and dry deposition onto soil, water, 

and vegetation. In surface water, PAHs can volatilize, photolyze, oxidize, biodegrade, bind to 

suspended particles or sediments, or accumulate in aquatic organisms (with bioconcentration factors 

often in the 10-10,000 range). In sediments, PAHs can biodegrade or accumulate in aquatic 

organisms. PAHs in soil can volatilize, undergo abiotic degradation (photolysis and oxidation), 

biodegrade, or accumulate in plants. PAHs in soil can also enter groundwater and be transported 

within an aquifer. 

Transport and partitioning of PAHs in the environment are determined to a large extent by 

physicochemical properties such as water solubility, vapor pressure, Henry’s law constant, 

octanol-water partition coefficient (Kow), and organic carbon partition coefficient (Koc). In general, 

PAHs have low water solubilities. The Henry’s law constant is the partition coefficient that expresses 

the ratio of the chemical’s concentrations in air and water at equilibrium and is used as an indicator of 

a chemical’s potential to volatilize. The Koc indicates the chemical’s potential to bind to organic 

carbon in soil and sediment. The Kow is used to estimate the potential for an organic chemical to 

move from water into lipid and has been correlated with bioconcentration in aquatic organisms. Some 

of the transport and partitioning characteristics (e.g., Henry’s law constant, Koc values, and Kow, 

values) of the 17 PAHs are roughly correlated to their molecular weights. These properties are 

discussed by grouping these PAHs as follows: 
PAHs 237 

5. POTENTIAL FOR HUMAN EXPOSURE 

•	 Low molecular weight compounds (152-l 78 g/mol)-acenaphthene, acenaphthylene,

 anthracene, fluorene, and phenanthrene; 

•	 Medium molecular weight compounds (202 g/mol)-fluoranthene and pyrene; and 

•	 High molecular weight compounds (228-278 g/mol)-benz[a]anthracene, 

benzo[b]fluoranthene, benzo[j]fluoranthene, benzo[k]fluoranthene, benzo[g,h,i]perylene, 

benzo[a]pyrene, benzo[e]pyrene, chrysene, dibenz[a,h]anthracene, and 

indeno[ 1,2,3-c,d]pyrene. 

As an example, Hattemer-Frey and Travis (1991) found that the low solubility, low vapor pressure and 

high Kow of benzo[a]pyrene result in its partitioning mainly between soil (82%) and sediment (17%), 

with ≈1% partitioning into water and <1% into air, suspended sediment and biota. 

PAHs are present in the atmosphere in the gaseous phase or sorbed to particulates. The phase 

distribution of PAHs in the atmosphere is important in determining their fate because of the difference 

in rates of chemical reactions and transport between the two phases. The phase distribution of any 

PAH depends on the vapor pressure of the PAH, the atmospheric temperature, the PAH concentration, 

the affinity of the PAH for the atmospheric suspended particles (Koc), and the nature and 

concentrations of the particles (Baek et al. 1991). In general, PAHs having two to three rings 

(naphthalene, acenaphthene, acenaphthylene, anthracene, fluorene, phenanthrene) are present in air 

predominantly in the vapor phase. PAHs that have four rings (fluoranthene, pyrene, chrysene, 

benz[a]anthracene) exist both in the vapor and particulate phase, and PAHs having five or more rings 

(benzo[a]pyrene, benzo[g,h,i]perylene) are found predominantly in the particle phase (Baek et al. 1991; 

Jones et al. 1992). The ratio of particulate to gaseous PAHs in air samples collected in Antwerp, 

Belgium, was 0.03 for anthracene, 0.49 for pyrene, 3.15 for summed benz[a]anthracene and chrysene, 

and 11.5 for summed benzo[a]fluoranthene and benzo[b]fluoranthene (NRC 1983). 

Using field data from Osaka, Japan, Pankow et al. (1993) examined the effects of relative humidity. 

(RH) on measured gas/particle partition coefficients over the range 42% ≥ 95%. They found that for 

seven PAHs or groups of PAHs (including phenanthrene + anthracene, fluoranthene, and pyrene) 

sorption decreased with increasing RH. 

Atmospheric residence time and transport distance depend on the size of the particles to which PAHs 

are sorbed and on climatic conditions which will determine rates of wet and dry deposition. About 

90-95% of particulate PAHs are associated with particle diameters <3.3 μm, and the peak distributions 
PAHs 238 

5. POTENTIAL FOR HUMAN EXPOSURE 

are localized between 0.4 and 1.1 μm (Baek et al: 1991). Both coarse particles with aerodynamic 

diameters >3-5 μm and nucleic particles with diameters <0.1 μm have limited atmospheric residence 

times. The coarse particles are removed from the atmosphere by wet and dry deposition, while the 

nucleic particles are removed mainly by coagulation with other nucleic particles or with larger 

particles, followed by wet and dry deposition. Particles with a diameter range of 0.1-3.0 μm, with 

which airborne PAHs are principally associated, remain airborne for a few days or longer, due to 

slower dry deposition and less efficient wet deposition (Baek et al. 1991). Therefore, airborne 

particulate PAHs in this size range can transport long distances (Lunde and Bjorseth 1977). Larger 

particles emitted from urban sources tend to settle onto streets and become part of urban runoff. 

However, PAHs in urban air are primarily associated with submicrometer-diameter soot particles that 

have residence times of weeks and are subject to long-range transport (Butler and Crossley 1981). 

Long-range transport of PAHs was examined by Lunde and Bjorseth (1977), Bjorseth et al. (1978a), 

and Bjorseth and Olufsen (1983) who found that PAHs originating in Great Britain had been 

transported as far as Norway and Sweden. 

The relative importance of wet and dry deposition in removing PAHs from the atmosphere varies with 

the individual PAH. For example, Perwak et al. (1982) estimated that a total of 23% of 

benzo[a]pyrene released to the atmosphere is deposited on soil and water surfaces. Dry deposition of 

benzo[a]pyrene adsorbed to atmospheric aerosols accounts for most of the removal; wet deposition is 

less significant by a factor of 3-5. In a mass balance study of the atmospheric deposition of PAHs to 

Siskiwit Lake, which is located on a wilderness island in northern Lake Superior, dry aerosol 

deposition of particulate phase PAHs was found to be the predominant form of input to surface waters 

by an average ratio of 9:l over wet deposition (McVeety and Hites 1988). 

PAH compounds tend to be removed from the water column by volatilization to the atmosphere, by 

binding to suspended particles or sediments, or by being accumulated by or sorbed onto aquatic biota. 

The transport of PAHs from water to the atmosphere via volatilization will depend on the Henry’s law 

constants (Hs) for these compounds. The low molecular weight PAHs have Henry’s law constants in 

the range of 10-3-10-5 atm-m3

/mol; medium molecular weight PAHs have constants in the 10-6 range; 

and high molecular weight PAHs have values in the range of 10-5-10-8. Compounds with values 

ranging from 10-3 to 10-5 are associated with significant volatilization, while compounds with values 

less than 10-5 volatilize from water only to a limited extent (Lyman et al. 1982). Half-lives for 

volatilization of benz[a]anthracene and benzo[a]pyrene (high molecular weight PAHs) from water have 
PAHs 239 

5. POTENTIAL FOR HUMAN EXPOSURE 

been estimated to be greater than 100 hours (Southworth 1979). Southworth et al. (1978) stated that 

lower molecular weight PAHs could be substantially removed by volatilization if suitable conditions 

(high temperature, low depth, high wind) were present. Southworth (1979) estimated half-lives for 

volatilization of anthracene (a low molecular weight PAH) of 18 hours in a stream with moderate 

current and wind, versus about 300 hours in a body of water with a depth of 1 meter and no current. 

Even for PAHs susceptible to volatilization, other processes, such as adsorption, photolysis or 

biodegradation (see Section 5.3.2.2) may become more important than volatilization in slow-moving, 

deep waters. 

Because of their low solubility and high affinity for organic carbon, PAHs in aquatic systems are 

primarily found sorbed to particles that either have settled to the bottom or are suspended in the water 

column. It has been estimated that two-thirds of PAHs in aquatic systems are associated with particles 

and only about one-third are present in dissolved form (Eisler 1987). In an estuary, volatilization and 

adsorption to suspended sediments with subsequent deposition are the primary removal processes for 

medium and high molecular weight PAHs, whereas volatilization and biodegradation (see 

Section 5.3.2.2) are the major removal processes for low molecular weight compounds (Readman et al. 

1982). In an enclosed marine ecosystem study, less than 1% of the original amount of radiolabeled 

benz[a]anthracene added to the system remained in the water column after 30 days; losses were 

attributed to adsorption to settling particles and to a lesser extent to photodegradation (Hinga and 

Pilson 1987). 

Baker et al. (1991) found that several PAHs were significantly recycled in the water column of Lake 

Superior. Fluorene and phenanthrene were rapidly removed from surface waters and settled through 

the water column to the sediment-water interface where a large fraction of the recently settled 

contaminants were released back into the water column. Higher molecular weight PAHs were found 

to have lower settling fluxes, but these compounds were efficiently buried in the surficial sediments 

with little recycling. Settling particles were found to be greatly enriched in hydrophobic organic 

chemicals. 

The Koc of a chemical is an indication of its potential to bind to organic carbon in soil and sediment. 

The low molecular weight PAHs have Koc values in the range of 103

-104, which indicates a moderate 

potential to be adsorbed to organic carbon in the soil and sediments. The medium molecular weight 

compounds have Koc values in the lo4 range. High molecular weight PAHs have Koc values in the 
PAHs 240 

5. POTENTIAL FOR HUMAN EXPOSURE 

range of 105-106, which indicates stronger tendencies to adsorb to organic carbon (Southworth 1979). 

PAHs from lands cleared by slash and bum methods have been found to be deposited in charred litter 

and to move into soils by partitioning and leaching (Sullivan and Mix 1985). Phenanthrene and 

fluoranthene (low and medium molecular weight PAHs, respectively) from these areas were 

incorporated into soil to a greater extent (i.e., less strongly adsorbed to organic carbon in the charred 

litter) than high molecular weight PAHs such as benzo[g,h,i]pyrene and indeno[ 1,2,3-c,d]pyrene. 

Because mobile colloids may enhance the mobility in porous medias of hydrophobic pollutants such as 

PAHs, Jenkins and Lion (1993) tested bacterial isolates from soil and subsurface environments for 

their ability to enhance transport of phenanthrene in aquifer sand. The most mobile isolates tested 

significantly enhanced the transport of phenanthrene, as a model PAH, in sand. 

Sorption of PAHs to soil and sediments increases with increasing organic carbon content and with 

increasing surface area of the sorbent particles. Karickhoff et al. (1979) reported adsorption 

coefficients for sorption of pyrene to sediments as follows: sand-9.4-68; silt-1,500-3,600; and 

clay-l ,400-3,800. Gardner et al. (1979) found that from three to four times more anthracene and 

about two times more fluoranthene, benz[a]anthracene, and benzo[a]pyrene were retained by marsh 

sediment than by sand. 

PAHs may also volatilize from soil. Volatilization of acenaphthene, acenaphthylene, anthracene, 

fluorene, and phenanthrene (low molecular weight PAHs) from soil may be substantial (Coover and 

Sims 1987; Southworth 1979; Wild and Jones 1993). However, of 14 PAHs studied in two soils, 

volatilization was found to account for about 20% of the loss of 1-methylnaphthalene and 30% of the 

loss of naphthalene; volatilization was not an important loss mechanism for anthracene, phenanthrene, 

fiuoranthene, pyrene, chrysene, benz[a]anthracene, benzo[b]fluoranthene, dibenz[a,h]anthracene, 

benzo[a]pyrene, and indeno[l,2,3-c,d]pyrene (Park et al. 1990). 

Physicochemical properties of several phenanthrene and anthracene metabolites [ 1-hydroxy-2-napthoic 

acid (lH2NA); 2,3-dihydroxy naphthalene (23DHN); 2-carboxy benzaldehyde (2CBA); and 3,4-

dihydroxy benzoic acid (34DHBA)] were experimentally measured and/or estimated and used with a 

Fugacity Level 1 model to estimate the distribution of the metabolites and their parent compounds in a 

contaminated soil (Ginn et al. 1994). The volumes of the air, water and soil phases were assumed to 

be 20%, 30%, and 48%, respectively. A volume of 2% was assumed for nonaqueous phase liquid 
PAHs 241 

5. POTENTIAL FOR HUMAN EXPOSURE 

(NAPL) phase. The parent compounds, anthracene and phenanthrene, had the greatest tendency to be 

associated with the NAPL and soil phases. The polar metabolites of phenanthrene, lH2NA and 

34DHBA, were associated more with the water phase of the subsurface. The metabolites 2CBA and 

23DHN had a stronger affinity for the NAPL phase than for the water phase. 

PAHs have been detected in groundwater either as a result of migration directly from contaminated 

surface waters or through the soil (Ehrlich et al. 1982; Wilson et al. 1986). Fluorene from an 

abandoned creosote pit was found to migrate through sand and clay into groundwater (Wilson et al. 

1986). PAHs have also been shown to be transported laterally within contaminated aquifers (Ehrlich 

et al. 1982). 

PAHs can be accumulated in aquatic organisms from water, sediments, and food. Bioconcentration 

factors (BCFs) for several species of aquatic organisms are listed in Table 5-2. In fish and crustaceans 

BCFs have generally been reported in the range of 10-10,000 (Eisler 1987). In general, 

bioconcentration was greater for the higher molecular weight compounds than for the lower molecular 

weight compounds. Bioconcentration experiments performed with radiolabeled compounds may 

overestimate the BCFs of some PAHs. For example, Spacie et al. (1983) estimated BCFs of 900 for 

anthracene and 4,900 for benzo[a]pyrene in bluegills (whole body) based on total radiolabeled carbon 

(

14C) activity. However, the estimated BCFs based only on the parent compounds were 675 and 490, 

respectively, indicating that biotransformation of the parent compounds occurred in addition to 

bioconcentration. Biotransformation by the mixed function oxidase (MFO) system in the fish liver can 

result in the formation of carcinogenic and mutagenic intermediates; exposure to PAHs has been linked 

to the development of tumors in fish (Eisler 1987). The ability of fish to metabolize PAHs may 

explain why benzo[a]pyrene frequently is not detected or found only at very low levels in fish from 

environments heavily contaminated with PAHs (Varanasi and Gmur 1980, 1981). The breakdown 

products (polyhydroxy compounds) are eliminated in feces (via bile) and urine. Although fish and 

most crustaceans evaluated to date have the MFO system required for biotransformation of PAHs, 

some molluscs and other aquatic invertebrates are unable to metabolize PAHs efficiently (Varanasi et 

al. 1985). Varanasi et al. (1985) ranked the extent of benzo[a]pyrene metabolism by aquatic 

organisms as follows: fish > shrimp > amphipod crustaceans > clams. Half-lives for elimination of 

PAHs in fish ranged from >2 days to 9 days (Niimi 1987). 


PAHs 244 

5. POTENTIAL FOR HUMAN EXPOSURE 

Mollusks also eliminate accumulated PAHs. Neff (1982) reported that oysters (Crassostrea gigas) 

eliminated the following percentages of accumulated PAHs during a 7-day elimination period: 

benzo[a]pyrene-0%; benz[a]anthracene-32%; fluoranthene-66%; and anthracene-79%. 

Fish and crustaceans readily assimilate PAHs from contaminated food, whereas mollusks and 

polychaete worms have limited assimilation (Eisler 1987). Biomagnification (a systematic increase in 

tissue concentrations moving up a food chain) has not been reported because of the tendency of many 

aquatic organisms to eliminate these compounds rapidly (Eisler 1987). In general, PAHs obtained 

from the diet contribute to total tissue concentrations only to a limited extent. For example, food 

chain uptake of anthracene by fathead minnows (Pimephales promelas) consuming water fleas 

(Daphnia pulex) was estimated to be about 15% of the amount accumulated from the water 

(Southworth 1979). In a simple aquatic food chain involving seston (i.e., organic and inorganic 

particulate matter >0.45 μm), blue mussels, and the common eider duck, significant changes were 

observed in the composition of 19 PAHs moving through the trophic levels. Decreasing PAH 

concentrations were found with increasing trophic level, probably as a result of the selective 

biotransformation capacity of the organisms for different PAHs. The high theoretical flux of PAHs 

through the food chain did not result in increasing concentrations with increasing trophic level (i.e., 

biomagnification was not observed), indicating rapid biotransformation of the compounds (Broman et 

al. 1990). 

Sediment-associated PAHs can be accumulated by bottom-dwelling invertebrates and fish (Eisler 

1987). For example, Great Lakes sediments containing elevated levels of PAHs were reported by 

Eadie et al. (1983) to be the source of the body burdens of the compounds in bottom-dwelling 

invertebrates. Varanasi et al. (1985) found that benzo[a]pyrene was accumulated in fish, amphipod 

crustaceans, shrimp, and clams when estuarine sediment was the source of the compound. 

Approximate tissue to sediment ratios were 0.6-1.2 for amphipods, 0.1 for clams, and 0.05 for fish 

and shrimp. 

Some terrestrial plants can take up PAHs from soil via the roots or from air via the foliage; uptake 

rates are dependent on the concentration, solubility, and molecular weight of the PAH and on the plant 

species (Edwards 1983). Mosses and lichens have been used to monitor atmospheric deposition of 

PAHs (Thomas et al. 1984). About 30-70% of atmospheric PAHs (indeno[l,2,3-c,d]pyrene, 

fluoranthene, and benzo[a]pyrene) deposited on a forest were sorbed onto tree foliage (i.e., leaves and 
PAHs 245 

5. POTENTIAL FOR HUMAN EXPOSURE 

needles) and then deposited as falling litter (Matzner 1984). Vaughan (1984) stated that atmospheric 

deposition on leaves often greatly exceeds uptake from soil by roots as a route of PAH accumulation. 

The uptake of PAHs from soil to plants and the subsequent biomagnification is generally quite low 

(Sims and Overcash 1983). Ratios of PAH concentrations in vegetation to those in soil have been 

reported to range from 0.001 to 0.18 for total PAHs and from 0.002 to 0.33 for benzo[a]pyrene 

(Edwards 1983). ,In a study of PAH uptake from cropland soils conducted in the United Kingdom, 

elevated concentrations of PAHs in soils were not correlated with concentrations in plant tissues (Wild 

et al. 1992). The cropland soils had received repeated applications of PAHs in sewage sludge that was 

applied to the soils over a number of years. PAH content of the soils substantially increased as a 

result of the sludge amendments, and residues of some PAHs persisted in the soils for years. Tissues 

from plants grown in the treated soils were relatively enriched with low molecular weight PAHs (e.g., 

acenaphthene, fluorene, phenanthrene), but increased PAH concentrations (relative to tissues from 

plants grown in control plots that did not receive sludge amendments) were not consistently detected. 

The PAH concentrations in aboveground plant parts were not strongly related to soil PAH levels but 

were probably the result of atmospheric deposition. The presence of PAHs in root crop tissues was 

probably due to adsorption of the compounds to root surfaces. In a similar study, Wild and Jones 

(1993) used carrots. (Daucus carotu) as a test crop to investigate the potential for PAHs to move from 

sewage sludge amended soil into the human food chain. Due to the over-riding influence of 

atmospheric delivery of PAHs, there was no evidence that sludge application increased the PAH 

concentration of the foliage. Low molecular weight PAHs such as fluoranthene and pyrene were 

relatively enriched in the peel, probably because of their greater bioavailability. Transfer of PAHs 

from the root peel to the core appeared to be minimal. This again suggests that simple adsorption 

onto the peel maybe an important process. 

Simonich and Hites (1994a) studied the partitioning of PAHs between vegetation and the atmosphere 

throughout the growing season and under natural conditions. They found the partitioning process to 

be dependent primarily upon the atmospheric gas-phase PAH concentration and the ambient 

temperature. During the spring and fall, when ambient temperatures are low, gas-phase PAHs partition 

into vegetation. In the summer, some PAHs volatilize and return to the atmosphere. They also 

developed a mass-balance model for PAHs in the northeastern United States and published values for 

PAH concentrations and fluxes in air, water, sediments, and soils (Simonich and Hites 1994b). Their 

model showed that 44 ± 18% of PAHs emitted into the atmosphere from sources in the region studied 
 

PAHs 246 

5. POTENTIAL FOR HUMAN EXPOSURE 

were removed by vegetation. They further hypothesized that most of the PAHs absorbed by 

vegetation at the end of the growing season are incorporated into the soil and permanently removed 

from the atmosphere. 

PAHs may accumulate in terrestrial animals through the food chain or by ingestion of soil. The 

environmental fate of creosote coal tar distillate (which contained 21% phenanthrene and 9% 

acenaphthene) was studied in a terrestrial microcosm containing soil, rye grass, insects, snails, 

mealworm larvae, and earthworms by Gile et al. (1982). Two gray-tailed voles (Microtus canicaudus) 

were added 54 days after the start of the experiment, which continued for 19-26 more days. Average 

surface soil concentrations (measured on an unspecified day) were 0.60 ppm (phenanthrene) and 

1.19 ppm (acenaphthene). During the last 3 days of the experiment, the following phenanthrene 

concentrations were measured: snail-3.27 ppm; pill bugs-l .72 ppm; and earthworm-l 8.30 ppm. 

The acenaphthene concentrations measured were as follows: snail-11.2 ppm (day 37); pill 

bugs-0.99 ppm (day 75); and earthworm-71.9 ppm (days 72-75). The whole body concentration in 

the vole analyzed for phenanthrene was 7.20 ppm; in the vole analyzed for acenaphthene it was 

37.00 ppm. The authors found that these compounds were not metabolized in this system. Whole 

body concentrations in the vole exceeded soil concentrations by a factor of 12 for phenanthrene and 

31 for acenaphthene; however, most of the radiolabeled acenaphthene was found as bound residues in 

the gastrointestinal tract of the animal and, therefore, was not accumulated. 

5.3.2 Transformation and Degradation 

5.3.2.1 Air 

The processes that transform and degrade PAHs in the atmosphere include photolysis and reaction with 

NOx, N2O5, OH, ozone, sulfur dioxide, and peroxyacetyl nitrate (Baek et al. 1991; NRC 1983). 

Possible atmospheric reaction products are oxy-, hydroxy-, nitro- and hydroxynitro-PAH derivatives 

(Baek et al. 1991). Photochemical oxidation of a number of PAHs has been reported with the 

formation of nitrated PAHs, quinones, phenols, and dihydrodiols (Holloway et al. 1987; Kamens et al. 

1986). Some of these breakdown products are mutagenic (Gibson et al. 1978). Reaction with ozone 

or peroxyacetylnitrate yields diones; nitrogen oxide reactions yield intro and dinitro PAHs. Sulfonic 

acids have also been formed from reaction with sulfur dioxide. 
PAHs 247 

5. POTENTIAL FOR HUMAN EXPOSURE 

The rates of homogeneous vapor phase chemical reactions are usually faster than heterogeneous 

chemical reactions of particulate PAHs with sunlight and oxidants in the atmosphere, particularly due 

to light shielding and stabilizing (toward both oxidation and photolysis) effects in the adsorbed state 

(Behymer and Hites 1988). 

PAHs have a wide range of volatilities and therefore are distributed in the atmosphere between the gas 

and particle phases. The 24 ring PAHs exist, at least partially, in the gas phase. Atkinson et al. 

(1991) calculated atmospheric lifetimes (1.44 times the half-life) of several gas-phase PAHs due to 

reactions with measured or estimated ambient concentrations of OH radicals, NO3 radicals, N2O5, and 

O3. Their laboratory studies showed that, for PAHs not containing cyclopenta-fused rings, the major 

gas-phase process resulting in atmospheric loss will be reaction with the OH radical. Calculated 

atmospheric lifetimes for acenaphthene, acenaphthylene, phenanthrene, and anthracene were on the 

order of a few hours. Nighttime reaction with N2O5 was estimated to be a minor source of 

atmospheric loss. The reactions of PAHs, including fluoranthene and pyrene, with the OH radical (in 

the presence of NOx) and with N2O5 led to the formation of nitroarenes that have been identified in 

the ambient air. As a class of compounds, the nitrated PAHs have been found to be much more 

mutagenic than their parent PAHs (Kamens et al. 1993). 

Most PAHs in the atmosphere .are associated with particulates (Baek et al.. 1991). Vu-Due and Huynh 

(1991) describe two types of chemical reactions that appear to be the predominant mode of 

transformation of these PAHs: (1) reactions between PAHs adsorbed on the particle surfaces and 

oxidant gases like NO2, O3, and SO3 that do not appear to be influenced by exposure to UV 

irradiation and (2) photooxidation of PAHs irradiated either under solar radiation or simulated sunlight 

which produces a variety of oxidized derivatives such as quinones, ketones, or acids. Kamens et al. 

(1990) estimate that, even in highly polluted air, photolysis is the most important factor in the decay 

of particle-sorbed PAHs in the atmosphere, followed by reaction with NO2, N2O5, and HNO3. 

The National Research Council (NRC 1983) noted that compounds adsorbed to soot are more resistant 

to photochemical reactions than pure compounds. Butler and Crossley (1981) estimated half-lives for 

degradation of the following PAHs adsorbed to soot particles and exposed to sunlight in air containing 

10 ppm nitrogen oxides: benzo[a]pyrene-7 days; benzo[g,h,i]perylene-8 days; 

benz[a]anthracene-11 days; pyrene-14 days; chrysene-26 days; fluoranthene-27 days; and 

phenanthrene-30 days. However, Thomas et al. (1968) reported that benzo[a]pyrene adsorbed on 
PAHs 248 

5. POTENTIAL FOR HUMAN EXPOSURE 

soot was readily photooxidized, with 60% of the compound destroyed within the first 40 minutes of 

exposure to sunlight. The effect of substrate on PAH photolytic half-lives was investigated by 

Behymer and Hites (1988). Photolysis of 18 PAHs adsorbed to low-carbon fly ash produced a wjde 

range of half-lives that indicated a relationship between structure and photochemical reactivity. 

Photolysis of the same compounds adsorbed to fly ash samples containing >5% carbon produced 

similar half-lives, indicating that for these fly ash samples, photolysis is dependent on the physical and 

chemical structure of the adsorbent and independent of PAH structure. The investigators postulated 

that dark (i.e., high carbon content) substrates stabilize PAHs to photolytic breakdown since they 

absorb more light, making less light available for photolysis. McDow et al. (1993) hypothesized that 

PAHs in atmospheric particles may be either dissolved in a liquid organic phase or adsorbed at an 

organic phase-solid elementai carbon interface. Therefore, the reactivity of PAHs might depend, not 

only on the surface characteristics of the particle’s solid core, but also on the chemical composition of 

the organic phase that surrounds the core. Experiments revealed that photodegradation of PAHs 

(including benz [a]anthracene, chrysene, benzo[a]pyrene, benzo[b]fluoranthene, and 

benzo[k]fluoranthene) in a mixture of methoxyphenols, based on relative amounts collected in actual 

samples from hardwood burning, was 10-30 times faster than in hexane. Their results demonstrated 

that variations in chemical composition of different types of particles such as diesel exhaust and wood 

smoke might strongly affect the reactivity of PAHs. Eisenberg and Cunningham (1985) found that the 

photochemical reaction products of PAHs (anthracene, phenanthrene, fluoranthene, benz[a]anthracene, 

chrysene, and benzo[a]pyrene) adsorbed on particulates include singlet oxygen, which may be 

implicated in the formation of mutagenic compounds. 

Some PAHs are degraded by oxidation reactions that have been measured in the dark (to eliminate the 

possibility of photodegradation). Korfmacher et al. (1980) found that, while fhrorene was completely 

oxidized, fluoranthene and phenanthrene were not oxidized, and benzo[a]pyrene and anthracene 

underwent minimal oxidation. These compounds were tested adsorbed to coal fly ash; the authors 

stated that the form of the compound (adsorbed or pure) and the nature of the adsorbent greatly 

affected the rate and extent of oxidation. 

Several studies have been carried out to investigate the reaction of PAHs with ozone at ambient 

concentrations (Baek et al. 1991). Alebic-Juretic et al. (1990) found degradation of PAHs on particle 

surfaces by ozone to be an important pathway for their removal from the atmosphere. Half-lives of 

PAHs obtained under laboratory conditions were used to predict lifetimes in an atmosphere containing 
PAHs 249 

5. POTENTIAL FOR HUMAN EXPOSURE 

a constant ozone concentration of 0.05 ppm. The predicted lifetimes were 3 hours for benzo[a]pyrene, 

12 hours for pyrene, and 6 days for fluoranthene. Lane and Katz (1977) investigated the kinetics of 

the dark reaction of several PAHs with ozone and found the reaction to be extremely fast under 

simulated atmospheric conditions, with a reported half-life of 0.62 hours for benzo[a]pyrene exposed 

to 190 ppb of ozone. 

In an attempt to determine the atmospheric oxidation processes that would result in an arene oxide 

functional group in PAHs, Murray and Kong (1994) studied the reaction of particle-bound PAHs with 

oxidants derived from the reactions of ozone with alkenes. Phenanthrene and pyrene were converted 

to arene oxides under these simulated atmospheric conditions. Control experiments indicated that the 

oxidant responsible for the transformation was not ozone, but a product of the reaction of ozone with 

tetramethylethylene (TME), probably the carbonyl oxide or the dioxirane derived from TME. 

5.3.2.2 Water 

The most important processes contributing to the degradation of PAHs in water are photooxidation, 

chemical oxidation, and biodegradation by aquatic microorganisms (Neff 1979). Hydrolysis is not 

considered to be an important degradation process for PAHs (Radding et al. 1976). The contribution 

of the individual processes,to the overall fate of a PAH will depend largely on the temperature, depth, 

pollution status, flow rate, and oxygen content of the water. As a result, a process that is a major 

loss/degradation process for a particular PAH in a certain surface water may not be so in another 

surface water with different water quality. 

The rate and extent of photodegradation vary widely among the PAHs (Neff 1979). Unfortunately, 

there is no easily defined trend in the rates of photolysis that could be correlated with the chemical 

structure of PAHs. For example, the rate of aquatic photolysis of naphthalene containing two benzene 

rings is much slower than anthracene which contains three benzene rings (Anderson et al. 1986). 

Based on half-life data, photolysis in water may be an important fate determining process for 

acenaphthene, acenaphthylene, anthracene, pyrene, benzo[a]pyrene, and benz[a]anthracene relative to 

the other PAHs discussed in this document (Behymer and Hites 1988; Anderson et al. 1986; Zepp and 

Schlotzhauer 1983). A study by Nagata and Kondo (1977) reported that anthracene, phenanthrene, and 

benz[a]anthracene were susceptible to photodegradation, and that benzo[a]pyrene, chrysene, fluorene, 

and pyrene were resistant to photodegradation. In the photooxidation of PAHs, the most common 
PAHs 250 

5. POTENTIAL FOR HUMAN EXPOSURE 

reactions result in the formation of peroxides, quinones, and diones (NAS 1972). The major 

photoproducts of anthracene, phenanthrene, and benz[a]anthracene are anthraquinone, 9,10­

phenanthrequinone, and 7,12-benz[a]anthraquinone, respectively (David and Boule 1993). 

The rate of photolysis is accelerated by the presence of certain sensitizers (Zepp and Schlotzhauer 

1983). Conversely, the rate of photolysis is decreased by the presence of certain quenchers in water 

(e.g., certain carbonyl compounds). The importance of photolysis will also decrease with the increase 

of depth in a body of water, particularly in turbid water, because of light attenuation and scattering 

(Zepp and Schlotzhauer 1979). 

Generally, oxidation with singlet oxygen and peroxy radicals are the two important oxidative processes 

for environmental pollutants in water. The rate constants for reactions of PAHs with singlet oxygen 

and peroxy radicals (Mabey et al. 1981) and the typical concentrations of the two oxidants in 

environmental waters (Mill and Mabey 1985) suggest that these reactions may not be important in 

controlling the overall fate of PAHs in water. 

PAHs in water can be chemically oxidized by chlorination and ozonation. A high efficiency of PAH 

degradation from chlorination has been reported by Harrison et al. (1976a, 1976b) for both laboratory 

and waste-water treatment plant conditions. Pyrene was the most rapidly degraded PAH. 

Benz[a]anthracene, benzo[a]pyrene, and perylene were also highly degraded. Indeno[ 1,2,3-c,d]pyrene 

and benzo[g,h,i]pyrene were intermediate with respect to relative degradation. Benzo[k]fluoranthene 

and fluoranthene were the most slowly degraded of the compounds tested. 

The PAH-related by-products resulting from chlorination are not fully known (Neff 1979). Oyler et al. 

(1978) identified the following products resulting from the chlorination of PAHs: anthraquinone, a 

chlorohydrin of fluoranthene; and monochloro derivatives of fluorene, phenanthrene, l-methylphenanthrene, and 1-methylnaphthalene. Mori et al. (1993) found that treatment of aqueous 

benz[a]anthracene (B[a]A) solution with chlorine in both the presence and absence of bromide ion 

produced a variety of halogenated compounds. The main product was the oxygenated compound, 

B[a]A-7, l2-dione. A variety of mutagenic halogen substituted and halogen additive (polar) 

compounds also were produced. The oxidation reaction with chlorine of B[a]A in water was 

accelerated in the presence of bromide ion. 
PAHs 251 

5. POTENTIAL FOR HUMAN EXPOSURE 

In water, ozonation is generally slower and less efficient than chlorination in degrading PAHs (Neff 

1979). Reaction pathways for ozonation of some PAHs include benz[a]anthracene to 7,12-quinone; 

benzo[a]pyrene to 3,6-, 1,6-, and 4,5diones; and fluorene to fluorenone (NAS 1972). 

In general, PAHs can be significantly metabolized by microbes under oxygenated conditions. 

However, under anoxic conditions, degradation will be extremely slow (Neff 1979). Concentrations of 

dissolved oxygen >0.7 mg/L is adequate for biotransformation and the presence of a minimal 

concentration of PAH is required for biodegradation to proceed (Borden et al. 1989). The minimum 

total PAH concentration below which biotransformation may be inhibited under ambient nutrient 

conditions may be 30-70 μg/L (Borden et al. 1989). Some other factors that increase the rates of PAH 

biodegradation are higher water temperature (summer versus winter) and the presence of adapted 

microorganisms (Aamand et al. 1989; Anderson et al. 1986; Lee and Ryan 1983). Some PAHs are 

partially or completely degraded by some species of aquatic bacteria and fungi. The bacterial 

degradation pathway includes an initial dioxygenase attack to form cis-dihydrodiols (via dioxetane 

intermediates) that are further oxidized to dihydroxy products. In fungi and mammalian systems 

(which, unlike bacteria, have cytochrome P-450 enzyme systems), truns-dihydrodiol is produced via an 

arene oxide intermediate (Anderson et al. 1986; Cemiglia and Heitkamp 1989; Neff 1979). This is 

significant since the arene oxides have been linked to the carcinogenicity of PAHs. Algae were found 

to transform benzo[a]pyrene to oxides, peroxides, and dihydrodiols (Kirso et al. 1983; Warshawsky et 

al. 1983). 

Microorganisms in stored groundwater samples completely degraded acenaphthene and acenaphthylene 

within 3 days (Ogawa et al. 1982). When these reactions occurred under aerobic conditions, there was 

no evidence of anaerobic degradation of PAHs within the aquifer from which the samples were 

obtained. 

Information on the biodegradation of PAHs by fungi is limited compared to the information that is 

available about bacteria. However, the fungus Cunninghamella elegans has been reported to be 

capable of metabolizing naphthalene (Cemiglia and Gibson 1979), anthracene, benzo[a]pyrene 

(Cemiglia and Heitkamp 1989), and fluorene (Pothuluri et al. 1993). 

No correlation between biodegradability and molecular weight is evident in three- to four-ring PAHs. 

For example, phenanthrene with three benzene rings biodegraded in an estuarine water from Savannah, 
 

PAHs 252 

5. POTENTIAL FOR HUMAN EXPOSURE 

Georgia, with a half-life of 19 days in August, but anthracene, containing the same number of benzene 

rings, did not biodegrade at all (Lee and Ryan 1983). Based on estimated reaction rates or half-lives, 

acenaphthylene, acenaphthylene, and fluorene, the three PAHs that have lower molecular weights than 

phenanthrene, may not readily biodegrade in water (Lee and Ryan 1983; Mabey et al. 1981). While 

both naphthalene and phenanthrene biodegraded in water, other PAHs, such as anthracene, 

benz[a]anthracene, chrysene, and fluorene, did not readily biodegrade in water, but degraded readily in 

sediment water slurries (Lee and Ryan 1983). On the other hand, PAHs with five or more benzene 

rings, such as benzo[a]pyrene, dibenz[a,h]anthracene, and benzo[g,h,i]perylene, may not biodegrade 

readily even in sediment-water slurries (Lee and Ryan 1983; Mabey et al. 1981). 

Based on theoretical modeling, photolysis would account for 5% and biodegradation 91% of the 

transformation/removal of anthracene from deep, slow moving, and somewhat turbid water. The 

corresponding values in a very shallow, fast-moving, clear water were 47 and 12%, respectively 

(Southworth 1979). 

5.3.2.3 Sediment and Soil 

Microbial metabolism is the major process for degradation of PAHs in soil environments. Photolysis, 

hydrolysis, and oxidation generally are not considered to be important processes for the degradation of 

PAHs in soils (Sims and Overcash 1983). However, in a study of PAH losses from four surface soils 

amended with PAHs in sewage sludge, losses due to volatilization and photolysis from sterilized soils 

were considered to be important for PAHs composed of less than four aromatic rings, whereas abiotic 

losses were insignificant for PAHs containing four or more aromatic rings (Wild and Jones 1993). 

Another study that assessed the fate of several PAHs, which included naphthalene, anthracene, 

phenanthrene, fluoranthene, pyrene, chrysene, benz[a]anthracene, benzo[b]fluoranthene, 

dibenz[a,h]anthracene, benzo[a]pyrene, dibenzo[a,i]pyrene and indeno[l,2,3-c,d]pyrene, in two soils 

concluded that abiotic degradation (photolysis and oxidation) accounted for mean losses of 13, 8.3, and 

15.8% loss in case of naphthalene, anthracene, and phenanthrene, respectively. No significant abiotic 

loss was observed for the other PAHs (Park et al. 1990). 

The rate and extent of biodegradation of PAHs in soil are affected by environmental factors; the 

organic content; structure and particle size of the soil; characteristics of the microbial population; the 

presence of contaminants such as metals and cyanides that are toxic to microorganisms; and the 
PAHs 253 

5. POTENTIAL FOR HUMAN EXPOSURE 

physical and chemical properties of the PAHs (Wilson and Jones 1993). Based on experimental 

results, the estimated half-lives (days) of the PAHs in soil were: naphthalene, 2.1-2.2; anthracene, 

50-l34; phenanthrene, 16-35; fluoranthene, 268-377; pyrene, 199-260; chrysene, 371-387; 

benz[a]anthracene, 162-261; benzo[b]fluoranthene, 211-294; benzo[a]pyrene, 229-309; 

dibenz[a,h]anthracene, 361-420; dibenzo(a,i)pyrene, 232-361; and indeno[l,2,3-c,d]pyrene, 288-289 

(Park et al. 1990). Although there are differences in the biodegradation half-life values estimated by 

different investigators (Park et al. 1990; Wild and Jones 1993; Symons et al. 1988), their results 

suggest that the biodegradation half-lives of PAH with more than three rings will be considerably 

longer (>20 days to hundreds of days) than the PAHs with three or fewer rings. Environmental factors 

that may influence the rate of PAH degradation in soil include temperature, Ph, oxygen concentration, 

PAH concentrations and contamination history of soil, soil type, moisture, nutrients, and other 

substances that may act as substrate co-metabolites (Sims and Overcash 1983). The size and 

composition of microbial populations in turn can be affected by these factors. For example, in low-Ph 

soils, fungi are dominant over bacteria, and thereby control microbial degradation in these 

environments. Sorption of PAHs to organic matter and soil particulates also influences bioavailability, 

and hence, biotransformation potential. Sorption of PAHs by soil organic matter may limit 

biodegradation of compounds that would otherwise rapidly undergo metabolism (Manila1 and 

Alexander 1991; Weissenfels et al. 1992). 

Although the pathways of microbial degradation are well known for anthracene, benzo[a]pyrene, and 

phenanthrene, degradation pathways for other PAHs are largely unknown (Sims and Overcash 1983). 

Metabolism of PAHs by bacteria includes the formation of cis-dihydrodiols through dioxetane 

intermediates, whereas in fungi (and mammalian systems) trans-dihydrodiols are produced through 

arene oxide intermediates (Sims and Overcash 1983). MacGillivray and Shiaris (1994) estimated the 

relative contribution of prokaryotic (bacteria) and eukaryotic (yeast, fungi) microorganisms to PAH 

biotransformation using phenanthrene as a model compound. They found that the relative contribution 

of eukaryotic microorganisms to phenanthrene transformation in inoculated sterile sediment was less 

than 3% of the total activity. 

In laboratory studies, Sims et al. (1988) demonstrated extensive degradation of two-ring PAHs in 

sandy soils, with half-lives of approximately 2 days. The three-ring PAHs, anthracene, and 

phenanthrene had half-lives of 16 and 134 days, respectively. Four- to six-ring PAHs generally had 

half-lives >200 days. Anthracene and fluoranthene showed slightly higher biodegradation rates than 
PAHs 254 

5. POTENTIAL FOR HUMAN EXPOSURE 

benz[a]anthracene or benzo[a]pyrene in a study with fine and medium sands and marsh sediments 

(Gardner et al. 1979). Degradation rates expressed as a percentage of the mass removed per week for 

the four compounds were anthracene-2.0-3.0%, fluoranthene-1.9-2.4%, benz[a]anthracene1.4-l .8%, and benzo[a]pyrene-0.84-1.4%. The ranges of half-lives of phenanthrene and 

benzo[g,h,i]perylene in four soils amended with PAHs in sewage sludge were 83-193 days and 

282-535 days, respectively. Mean half-lives were found to be positively correlated with log Kow and 

inversely correlated with log water solubility. Previous exposure of the test soils to PAHs enhanced 

the rate of biodegradation of low molecular weight PAHs but had little effect on the loss of higher 

molecular weight compounds (Wild and Jones 1993). 

Herbes and Schwall (1978) investigated the rates of microbial transformation of PAHs in freshwater 

sediments from both pristine and oil-contaminated streams. They found that turnover times (l/k) in 

the uncontaminated sediment were 10-400 times greater than in contaminated sediment. Absolute 

rates of PAH transformation (micrograms of PAH per gram of sediment per hour) were 

3,000-125,000 times greater in the contaminated sediment. Turnover times in the oil-contaminated 

sediment increased 30-l00-fold per additional ring from naphthalene through benz[a]anthracene; 

naphthalene was broken down in hours while the turnover times for benz[a]anthracene and 

benz[a]pyrene were ~400 days and >3.3 years, respectively. Therefore, four- and five-ring PAHs, 

including the carcinogenic benz[a]anthracene and benz[a]pyrene, may persist even in sediments that 

have received chronic PAH inputs. 

The rate of biodegradation may be altered by the degree of contamination. At hazardous waste sites, 

half-lives may be longer since other contaminants at the site may be toxic to degrading . 

microorganisms. Bossert and Bartha (1986) reported reduced biodegradation of PAHs in soil 

containing a chemical toxic to microorganisms. 

Efroymson and Alexander (1994) investigated the effects of nonaqueous phase-liquids (NAPLs) on the 

biodegradation of hydrophobic compounds, including phenanthrene, in soil and subsoil. Mineralization 

of phenanthrene in the subsoil was reduced if the compound was dissolved in a NAPL. However, the 

suppression of the mineralization of phenanthrene in soil by NAPLs was short-lived, suggesting 

growth of organisms capable of using phenanthrene. 
PAHs 255 

5. POTENTIAL FOR HUMAN EXPOSURE 

5.4 LEVELS MONITORED OR ESTIMATED IN THE ENVIRONMENT 

5.4.1 Air 

There is a relatively large body of data characterizing PAH air levels at a variety of U.S. sites. 

Caution must be used in interpreting and comparing results of different studies, however, because of 

the different sampling methods used. PAHs occur in the atmosphere in both the particle phase and the 

vapor phase, as discussed in Section 5.2.1. Three-ring PAH compounds are found in the atmosphere 

primarily in the gaseous phase, whereas, five- and six-ring PAHs are found mainly in the particle 

phase; four-ring PAH compounds are found in both phases. To fully characterize atmospheric PAH 

levels, both particle- and vapor-phase samples must be collected. Many of the earlier monitoring 

studies used filter sampling methods, which provided information on particle-phase PAH 

concentrations only, and which did not account for losses of some of the lower molecular weight 

PAHs by volatilization. As a result, the early use of particulate samples may have resulted in an 

underestimation of total PAH concentrations. More recent monitoring studies often use sampling 

methods that collect both particle- and vapor-phase PAHs and that prevent or minimize volatilization 

losses, thus providing more reliable characterization of total atmospheric PAH concentrations (Baek et 

al. 1991). 

Several monitoring studies indicate that there are higher concentrations of PAHs in urban air than in 

rural air. Pucknat (1981) summarized 1970 data from the U.S. National Air Surveillance Network and 

reported that benzo[a]pyrene concentrations in 120 U.S. cities were between 0.2 and 19.3 ng/m3

. 

Ambient benzo[a]pyrene concentrations in nonurban areas ranged between 0.1 and 1.2 ng/m3

. More 

recently, Greenberg et al. (1985) evaluated atmospheric concentrations of particulate phase PAHs at 

four New Jersey sites (three urban and one rural) over two summer and winter seasons during 

198l-82. Urban PAH concentrations were approximately 3-5 times higher than those at the rural site; 

in addition, winter concentrations were approximately 5-10 times higher than summer concentrations. 

Geometric mean concentrations of ten PAHs (benzo[a]pyrene, benzo[e]pyrene, benzo[b]fluoranthene, 

benzofi]fluoranthene, benzo[k]fluoranthene, benz[a]anthracene, indeno[l,2,3-c,d]pyrene, 

benzo[g,h,i]perylene, pyrene, and chrysene) ranged from 0.03 to 0.62 ng/m3

 in urban areas and from 

0.01 to 0.12 ng/m3

 in the rural area during the summer seasons. During the winter seasons, geometric 

mean concentrations of these PAHs ranged from 0.40 to 11.15 ng/m3

 in urban areas and from 0.08 to 

1.32 ng/m3

 in the rural area. Geometric mean concentrations of benzo[a]pyrene ranged from 0.11 to 
PAHs 256 

5. POTENTIAL FOR HUMAN EXPOSURE 

0.23 ng/m3 (urban) and 0.04 to 0.06 ng/m3 (rural) during the summer seasons, and from 0.69 to 

1.63 ng/m3 (urban) and 0.17 to 0.32 (rural) during the winter seasons. A more extensive study by 

Harkov and Greenberg (1985) of atmospheric benzo[a]pyrene concentrations at 27 New Jersey sites 

indicated similar differences in mean urban (0.6 ng/m3

) and rural (0.3 ng/m3) concentrations. 

Significant seasonal trends were also observed, with mean benzo[a]pyrene concentrations during the 

winter more than an order of magnitude greater than during the summer. 

Several other studies provide evidence that atmospheric concentrations of particle-phase PAHs are 

higher in winter than in summer. In a 1981-82 study conducted in the Los Angeles area; atmospheric 

concentrations of 10 PAHs (anthracene, fluoranthene, pyrene, chrysene, benz[a]anthracene, combined 

benzo[e]pyrene and perylene, benzo[b]fluoranthene, benzo[k]fluoranthene, benzo[a]pyrene, and 

combined benzo[g,h,i]perylene, and indeno[l,2,3-c,d]pyrene) ranged from 0.14 to 1.45 ng/m3 (with an 

average of 0.43 ng/m3) during the summer (August-September), and from 0.40 to 4.46 ng/m3 (with an 

average of 1.28 ng/m3

) during the winter (February-March) (Grosjean 1983). A similar seasonal 

variation in particle-phase PAH concentrations in the Los Angeles atmosphere was seen in an earlier 

1974-75 study (Gordon 1976). Quarterly geometric mean concentrations of 11 PAHs (pyrene, 

fluoranthene, benz[a]anthracene, chrysene, benzo[a]pyrene, benzo[e]pyrene, benzo[b]fluoranthene, 

benzo[j]fluoranthene, benzo[k]fluoranthene, benzo[g,h,i]perylene, and indeno[ 1,2,3-c,d]pyrene) ranged 

from 0.06 to 2.71 ng/m3 (with an average of 0.45 ng/m3) during the May-October period, and from 

0.26 to 8.25 ng/m3 (with an average of 1.46 ng/m3) during the November-April period. The highest 

and lowest concentrations were observed during the fourth (November-January) and second 

(May-July) quarters, respectively. Ratios of fourth quarterly and second quarterly geometric mean 

concentrations ranged from 3.9 for indeno[l,2,3-c,d]pyrene to 7.5 for benzo[a]pyrene and 9.8 for 

benz[a]anthracene. Possible factors contributing to these seasonal variations in PAH levels include the 

following: changes in emission patterns; changes in meteorological conditions (i.e., daylight hours and 

temperature); and changes in space heating emissions, volatilization, and photochemical activity. 

Certain monitoring data suggest that ambient levels of some PAHs may be decreasing. Faoro and 

Manning (1981) analyzed a limited sample of U.S. National Air Surveillance Network data updated 

through 1977, which indicated that benzo[a]pyrene concentrations have shown consistent, sizable 

declines during the period from 1967 to 1977 at 26 urban sites and 3 background sites studied (data 

not provided). 
PAHs 257 

5. POTENTIAL FOR HUMAN EXPOSURE 

Over the past two decades, the ambient air levels of PAHs in a number of major cities have been 

characterized. Although data from studies in different areas cannot be used to indicate definitive 

temporal trends in PAH air levels, a comparison of the results of these studies yields no strong 

suggestion that the ambient air levels of PAHs may be decreasing, except in traffic tunnels. 

In a 1981-82 study that characterized air levels of 13 PAHs in Los Angeles, Grosjean (1983) reported 

mean ambient particle-phase PAH concentrations ranging from 0.32 ng/m3 for benzo[k]fluoranthene to 

3.04 ng/m3 for combined benzo[g,h,i]perylene and indeno[ 1,2,3-c,d]pyrene. Mean concentrations of 

anthracene, fluoranthene, pyrene, chrysene, benz[a]anthracene, combined perylene and benzo[e]pyrene, 

benzo[b]fluoranthene, and benzo[a]pyrene were 0.54; 0.94, 1.62, 0.97, 0.48, 0.43, 0.94, and 

0.64 ng/m3, respectively. Similar results were obtained in an earlier (1974-1975) study of atmospheric 

particle-phase PAHs in the Los Angeles area, where ambient annual geometric mean concentrations 

ranged from 0.17 ng/m3 for benzo[j]fluoranthene to 3.27 ng/m3 for benzo[g,h,i]perylene (Gordon 

1976). The annual geometric mean concentration of benzo[a]pyrene was 0.46 ng/m3; most individual 

PAHs had annual geometric mean concentrations of <0.6 ng/m3

. The relatively high levels of 

benzo[g,h,i]perylene found in these studies.have been attributed to high levels of automobile 

emissions, which are known to contain high levels of benzo[g,h,i]perylene relative to other PAHs 

(Santodonato et al. 1981). During the same time period, Fox and Staley (1976) reported somewhat 

higher ambient average concentrations of particle-phase PAHs in College Park, Maryland, ranging 

from 3.2 ng/m3 for benzo[a]pyrene to 5.2 ng/m3 for pyrene. 

In a 1985-86 study, reported average ambient concentrations (combined particle- and vapor-phase) of 

eight PAHs in Denver ranged between 0.83 ng/m3

 for benzo[k]fluoranthene and 39 ng/m3 for 

phenanthrene (Foreman and Bidleman 1990). In a study conducted in Hamilton, Ontario, between 

May 1990 and June 1991, the concentrations of PAHs in respirable air particulate samples were found 

to range from 0.6 ng/m3

 for phenanthrene to 4.3 ng/m3 for benzo[g,h,i]perylene, and 5.1 ng/m3 for 

combined benzo[b,j,k]fluoranthenes (Legzdins et al. 1994). In a recent limited study, mean 

concentrations of particle-phase PAHs in New York City air were reported to range from 0.11 ng/m3 

for anthracene to 4.05 ng/m3 for benzo[g,h,i]perylene (Tan and Ku 1994). 

Atmospheric PAH concentrations have been found to be significantly elevated in areas of enclosed 

traffic tunnels. In a 1985-86 study in the Baltimore Harbor Tunnel the average concentrations of 

particle-phase PAHs ranged from 2.9 ng/m3

 for anthracene to 27 ng/m3 for pyrene (Benner and 
PAHs 258 

5. POTENTIAL FOR HUMAN EXPOSURE 

Gordon, 1989). These values are up to an order of magnitude lower than those obtained in 1975 by 

Fox and Staley (1976), which ranged from 66 ng/m3 for benzo[a]pyrene to 120 ng/m3 for pyrene. 

Benner and Gordon (1989) postulated that the observed decrease in PAH concentrations over the 

1975-85 decade resulted from the increasing use of catalytic converters in U.S. automobiles over that 

period, These authors also reported concentrations of PAHs in a typical vapor-phase sample from the 

Boston Harbor Tunnel for four PAHs included in this profile: anthracene (32.3 ng/m3), fluoranthene 

(25.6 ng/m3), phenanthrene (184 ng/m3), and pyrene (28.3 ng/m3). They emphasized that the vaporphase samples included PAHs inherently present in the vapor phase as well as the more volatile 3- and 

4-ring PAHs that may be desorbed from particles during sampling. These results underscore the need 

to evaluate both particle- and vapor-phase samples to obtain more reliable estimates of total 

atmospheric PAH concentrations. 

5.4.2 Water 

PAHs have been detected in surface waters of the United States. In an assessment of STORET data 

covering the period 1980-82, Staples et al. (1985) reported median concentrations in ambient water of 

<10 μg/L for 15 PAHs (acenaphthene, acenaphthylene, anthracene, benz[a]anthracene, 

benzo[b]fluoranthene, benzo[k]fluoranthene, benzo[g,h,i]perylene, benzo[a]pyrene, chrysene, 

fluoranthene, fluorene, indeno[ 1,2,3-c,d]pyrene, naphthalene, phenanthrene,. and pyrene). The number 

of samples ranged from 630 (naphthalene) to 926 (fluoranthene); the percentage of samples in which 

these PAHs were detected ranged from 1.0 (benzo[g,h,i]perylene) to 5.0 (phenanthrene) and 

7.0 (naphthalene). 

Basu and Saxena (1978a) reported concentrations of selected PAHs in surface waters used as drinking 

water sources in four U.S. cities (Huntington, West Virginia; Buffalo, New York; and Pittsburgh and 

Philadelphia, Pennsylvania). Total concentrations of PAHs ranged from 4.7 ng/L in Buffalo to 

600 ng/L in Pittsburgh. Mean concentrations of benzo[a]pyrene in the Great Lakes have been detected 

at levels between 0.03 and 0.7 ppt (ng/L) (Environment Canada 1991). 

DeLeon et al. (1986) analyzed surface water from 11 locations in the Mississippi River. Seventeen 

PAHs were identified in the samples at levels ranging from 1 ng/L for 6 compounds to a high of 

34 ng/L for phenanthrene. The highest concentration of phenanthrene was detected in a sample 
PAHs 259 

5. POTENTIAL FOR HUMAN EXPOSURE 

collected near New Orleans, Louisiana, near an industrial area, implicating industrial effluent or 

surface runoff from this area as a possible source. 

During April and May 1990, Hall et al. (1993) analyzed 48-hour composite samples from three 

locations in the Potomac River and three locations in the upper Chesapeake Bay for eight PAHs: 

perylene, fluorene, phenanthrene, anthracene, fluoranthene, pyrene, benz[a]anthracene, and chrysene. 

Pyrene was the.only PAH found (0.42 μg/L) in these samples; it was detected in only one of nine 

Chesapeake Bay samples and not detected in any of the Potomac River samples (detection limit, 

0.04 μg/L). 

In a more recent study by Pham et al. (1993), raw water samples from 5 areas in the St. Lawrence 

River and its tributaries were analyzed for 12 PAHs. The highest mean total PAH concentrations were 

observed in samples collected in the spring (27.3 ng/L) and autumn (21.03 ng/L), which was attributed 

to snow melt and increased runoff during these respective seasons. The lowest mean total PAH 

concentration was observed in summer (14.63 ng/L). High molecular weight PAHs were detected 

more frequently in the spring and autumn samples. Phenanthrene, benzo[b]fluoranthene, fluoranthene, 

and pyrene were predominant, comprising on average 33.8%, 17.4%, 17.1%, and 12.8% of the total 

PAHs, respectively: With the exception of anthracene and benzo[b]fluoranthene, a general decrease in 

concentration with increasing molecular weight was observed. 

PAHs have been detected in urban runoff generally at concentrations much higher than those reported 

for surface water. Data collected as part of the Nationwide Urban Runoff Program indicate 

concentrations of individual PAHs in the range of, 300-10,000 ng/L, with the concentrations of most 

PAHs above 1,000 ng/L (Cole et al. 1984). In a recent study by Pitt et al. (1993) which involved the 

collection and analysis of approximately 140 urban runoff samples from a number of different source 

areas in Birmingham, Alabama, and under various rain conditions, fluoranthene was one of two 

organic compounds detected most frequently (23% of samples). The highest frequencies of detection 

occurred in roof runoff, urban creeks, and combined sewer overflow samples. The maximum reported 

concentration of fluoranthene in these samples was 130 μg/L. 

Industrial effluents also have elevated PAH levels. Morselli and Zappoli (1988) reported elevated 

PAH levels in refinery waste waters, with concentrations for most PAHs in the range of 400 ng/L 

(benzo[b]fluoranthene) to 16,000 ng/L (phenanthrene). In an analysis of STORET data covering the 
PAHs 260 

5. POTENTIAL FOR HUMAN EXPOSURE 

period 1980-88, Staples et al. (1985) reported median concentrations in industrial effluents of 

<10 μg/L (10,000 ng/L) for 15 PAHs. The number of samples ranged from 1,182 (benzo­

[blfluoranthene) to 1,288 (phenanthrene); the percentage of samples in which PAHs were detected 

ranged from 1.5 (benzo[g,h,i]perylene) to 7.0 (fluoranthene). 

Few data are available on the concentrations of PAHs in U.S. groundwater. Basu and Saxena (1978b) 

reported total PAH concentrations in groundwater from three sites in Illinois, Indiana, and Ohio to be 

in the range of 3-20 ng/L. Groundwater levels of PAHs near a coal and oil gasification plant and 

U.S. wood treatment facilities have been found to be elevated. Groundwater samples from the site of 

a Seattle coal and oil gasification plant which ceased operation in 1956 were found to contain 

acenaphthylene, acenaphthene, fluorene, phenanthrene, fluoranthene, pyrene, and chrysene at 

concentrations ranging from not detected (detection limit 0.005 mg/L) to 0.25, 0.18, 0.14, 0.13, 0.05, 

0.08, and 0.01 mg/L, respectively (Tumey and Goerlitz 1990). Individual PAHs in the groundwater 

from 5 U.S. wood treatment facilities were reported at average concentrations of 57 ppb (0.057 mg/L) 

for benzo[a]pyrene to 1,825 ppb (1.8 mg/L) for phenanthrene (Rosenfeld and Plumb 1991). 

An evaluation of the analytical data from 358 hazardous waste sites with over 5,000 wells indicated 

that anthracene, fluoranthene, and naphthalene were detected (practical quantitation limit, 10-200 μg/L) 

in groundwater from at least 0.1% of the sites in three of the ten EPA Regions into which the United 

States is divided (Garman et al. 1987). A review of groundwater monitoring data from 479 waste 

disposal sites (178 CERCLA or Super-fund sites, 173 RCRA sites, and 128 sanitary/municipal landfill 

sites) located throughout the United States indicated that 14 of the PAHs included in this profile were 

detected at frequencies ranging from 2 detections at one site in one EPA Region for 

indeno[l,2,3-c,d]pyrene, to 85 detections at 16 sites in 4 EPA Regions for fluorene (Plumb 1991). 

Benzo[a]pyrene was detected 13 times at 6 sites in 6 EPA Regions. Concentrations were not reported. 

Data summarized by Sorrel et al. (1980) indicate low levels of PAHs in finished drinking waters of the 

United States. Reported maximum concentrations for total PAHs (based on measurement of 15 PAHs) 

in the drinking water of 10 cities ranged from 4 to 24 ng/L; concentrations in untreated water ranged 

from 6 to 125 ng/L. The low concentrations of PAHs in finished drinking water were attributed to 

efficient water treatment processes. Shiraishi et al. (1985) found PAHs in tap water at concentrations 

of 0.1-1.0 ng/L, primarily as chlorinated derivatives of naphthalene, phenanthrene, fluorene, and 

fluoranthene. The significance to human health of these compounds is not known (Eisler 1987). 
PAHs 261 

5. POTENTIAL FOR HUMAN EXPOSURE 

5.4.3 Sediment and Soil 

PAHs are ubiquitous in soil. Because anthropogenic combustion processes are a major source of 

PAHs in soils, soil concentrations have tended to increase over the last 100-150 years, especially in 

urban areas (Jones et al. 1989a, 1989b). Background concentrations for rural, agricultural, and urban 

soils (from the United States and other countries) are given in Table 5-3. In general, concentrations 

ranked as follows: urban > agricultural > rural. Evidence of the global distribution of PAHs was 

given by Thomas (1986) who detected benzo[g,h,i]perylene and fluoranthene at concentrations above 

150 μg/kg in arctic soils. Soil samples collected from remote wooded areas of Wyoming contained 

total PAH concentrations of up to 210 μg/kg. 

Recent data on PAH concentrations in soil at contaminated sites are summarized in Table 5-4. 

Because of the different sampling methods and locations at each site, this tabulation does not provide a 

reliable inter-site comparison. Additional studies indicate significantly elevated concentrations of 

PAHs at contaminated sites. Soil samples collected from the Fountain Avenue Landfill in New York 

City contained PAH concentrations ranging from 400 to 10,000 μg/kg (Black et al. 1989). In a 1988 

study at a hazardous waste land treatment site for refinery process wastes, which had been operative 

since 1958, average PAH concentrations in surface soils (0-30 cm) ranged from not detected 

(detection limits 0.1-2.0 mg/kg dry weight) for acenaphthylene, acenaphthene, anthracene, 

benz[a]anthracene, and benzo[k]fluoranthene to 340 mg/kg dry weight for dibenz[a,h]anthracene 

(Loehr et al. 1993). In addition to dibenz[a,h]anthracene, the three most prevalent compounds at this 

depth were benzo[a]pyrene (204 mg/kg), benzo[b]fluoranthene (130 mg/kg), and chrysene 

(100 mg/kg). PAH concentrations decreased with increasing depth and the majority of PAHs were not 

detected at depths below 60 cm. At 90-135 cm, only phenanthrene (1.4 mg/kg), pyrene (4.0 mg/kg), 

chrysene (0.9 mg/kg), and dibenz[a,h]anthracene (0.8 mg/kg) were found. 

Sediments are major sinks for PAHs, primarily because of the low solubility of these compounds and 

their strong affinity for organic carbon in particulate matter. PAH concentrations in sediment are 

generally much higher than those detected in surface water, i.e., in the range of μg/kg (ppb) rather than 

ng/kg (ppt) . 

In an assessment of STORET data covering the period 1980-1982, Staples et al. (1985) reported 

median concentrations in sediment of ≤500 μg/kg dry weight for 15 PAHs (acenaphthene, 


PAHs 264 

5. POTENTIAL FOR HUMAN EXPOSURE 

acenaphthylene, anthracene, benz[a]anthracene, benzo[b]fluoranthene, benzo[k]fluoranthene, 

benzo[g,h,i]perylene, benzo[a]pyrene, chrysene, fluoranthene, fluorene, indenopyrene, naphthalene, 

phenanthrene, and pyrene). The number of samples ranged from 236 (anthracene) to 360 (benzo­

[alpyrene, fluoranthene); the percentage of samples in which these PAHs were detected ranged from 

6.0 (acenaphthene, benzo[b]fluoranthene, benzo[k]fluoranthene, indeno[ 1,2,3-c,d]pyrene) to 

22.0 (fluoranthene, pyrene). 

Eadie et al. (1982) analyzed surficial sediments in southwestern Lake Erie near a large coal-fired 

power plant. Sediment concentrations for total PAHs were generally in the range of 530-700 μg/kg, 

although concentrations in river and near-shore sediments reached nearly 4,000 μg/kg (4 ppm). Heit et 

al. (1981) reported total concentrations of PAHs (3-7 ring PAHs) from two lakes in the Adirondack 

acid lake region of 2,660 μg/kg and 770 μg/kg (calculated from data presented). Average 

concentrations of total PAHs in sediments from three coastal South Carolina marinas were reported to 

range from 35.6 to 352.3 μg/kg (Marcus et al. 1988). Benzo[a]pyrene levels in bottom sediments of 

the Great Lakes have been reported to range from 34 to 490 ppb (μg/kg) (Environment Canada 1991). 

Concentrations of PAHs in sediments from Cape Cod and Buzzards Bay in Massachusetts and the Gulf 

of Maine have been reported to be in the range of 540-1,300 μg/kg (Hites et al. 1980). 

Concentrations of low molecular weight PAHs (naphthalene, acenaphthylene, fluorene, phenanthrene, 

anthracene, and 2-methylnaphthalene) and high molecular weight PAHs (fluoranthene, pyrene, 

benz[a]anthracene, chrysene, benzofluoranthenes, benzo[a]pyrene, indeno[l,2,3-c,d]pyrene, 

dibenz[a,h]anthracene, and benzo[g,h,i]perylene) in sediment from the highly polluted Boston Harbor 

have been reported to range from approximately 100 to 11,000 μg/kg dry wt, and 800 to 23,000 μg/kg 

dry wt, respectively (Demuth et al. 1993). 

Total PAH concentrations in bottom sediments from the main stem of the Chesapeake Bay were 

reported to range from 45 to 8,920 μg/kg for samples collected from 16 stations in 1986 (Huggett et 

al. 1988). At least 14 PAHs were found to be dominant among pollutants of surface sediments from 

the Elizabeth River, a subestuary of the James River in Virginia, with a maximum total PAH 

concentration of 170,000 μg/g (ppm) observed in one sample from a site of two large wood 

preservative spills (Bieri et al. 1986). In a more recent study, surface sediment samples from the 

highly contaminated Elizabeth River were found to contain total concentrations of 14 PAHs ranging 

from 1.5 to 4,230 μg/g (ppm) dry weight (Halbrook et al. 1992). Significantly lower concentrations, 
PAHs 265 

5. POTENTIAL FOR HUMAN EXPOSURE 

ranging from 0.34 to 0.95 μg/g (340-950 μg/kg) dry weight, were found in sediment samples from the 

nearby Nansemond River which served as a clean reference site. 

Two-thirds of 105 sediment samples collected throughout Florida during the summers of 1989 and 

1990 from sites known or suspected to be contaminated with priority pollutants were found to contain 

at least one of 15 PAH target analytes (Jacobs et al. 1993). Pyrene was detected most frequently (61% 

of samples); dibenz[a,h]anthracene and naphthalene were detected least frequently (4% of samples). 

Total PAH concentrations ranged from below the detection limit to 1,090 mg/kg. Mean concentrations 

for individual PAHs ranged from 0.87 mg/kg (dibenz[a,h]anthracene and naphthalene) to 30.8 mg/kg 

(acenaphthene). 

Drainage stream sediments from a wood-preserving facility near Pensacola, Florida, were found to be 

highly contaminated with creosote-derived PAHs, with maximum concentrations from two sampling 

sites ranging from 300 μg/kg for naphthalene to 12,000 μg/kg for phenanthrene and 140,000 μg/kg for 

anthracene (Elder and Dresler 1988). Fluoranthene, pyrene, benz[a]anthracene, chrysene, 

acenaphthene, and fluorene were other dominant PAHs. PAHs were not detected in water samples 

from the drainage stream. Furthermore, no significant PAH contamination was found in surface 

sediments from estuarine sites adjacent to the drainage stream; PAHs were detected in sediment 

samples from only one of seven estuarine sites at concentrations ranging from 75 μg/kg for 

benz[a]anthracene to 190 μg/kg for fluoranthene. 

In 1991, Kennicutt et al. (1994) found that sediment samples from Casco Bay in Maine contained total 

PAH concentrations ranging from 16 to 20,800 μg/kg dry weight. PAHs were found at all 

65 locations sampled. PAHs with four or more rings accounted for more than 60% of Casco Bay 

sedimentary PAHs. The predominance of PAHs with highly condensed ring structures with few 

alkylations indicated a pyrogenic or combustion source as the major contributor. 

Mean total PAH concentrations of sediments collected in 1985-87 from Moss Landing Harbor, 

Elkhom Slough, and nearshore Monterey Bay, California, were found to range from 1,470 to 3,080, 

157 to 375, and 24 to 114 μg/kg dry weight, respectively (Rice et al. 1993). The Moss Landing 

Harbor and nearshore Monterey Bay ecosystems are subject to PAH contamination from various local 

industries, harbor-related activities, power generation, municipal waste treatment, and agricultural 

runoff. The largest Pacific Coast fossil-fueled power plant is located at Moss Landing. Elkhom 
PAHs 266 

5. POTENTIAL FOR HUMAN EXPOSURE 

Slough is a seasonal estuary which receives freshwater runoff. Combustion PAHs (i.e., 

benz[a]anthracene, benzo[b]fluoranthene, benzo[k]fluoranthene, benzo[g,h,i]perylene, benzo[a]pyrene, 

benzo[e]pyrene, chrysene, fluoranthene, and pyrene) were predominant, with mean total concentration 

at these 3 sites ranging from 1,250 to 2,710, to 335, and 11 to 59 ug/kg dry weight, respectively. 

Median concentrations of PAHs in sediment coves collected in 1991 from three northern New Jersey 

waterways (Arthur Kill, Hackensack River, and Passaic River) highly contaminated with petroleum 

hydrocarbons ranged from 0.47 mg (470 ug/kg) (acenaphthylene) to 5.10 mg (5,100 μg/kg) (pyrene) 

(Huntley et al. 1993). In addition to pyrene, fluoranthene, chrysene, and benzo[a]pyrene were the 

most frequently detected PAHs, with median concentrations at the three sites ranging from 2.40 to 

4.10, 1.35 to 2.85, and 0.86 to 2.30 mg/kg, respectively. Mean total PAH concentrations at the 

29 sampling stations ranged from 0 to 161 mg/kg. A mean total PAH concentration of 139 mg/kg was 

found at a sampling station downstream from a chemical control Superftmd site. At most sampling 

stations, PAH concentrations increased with sample depth up to approximately 45-50 cm, indicating a 

decline in recent loadings relative to historic inputs. 

5.4.4 Other Environmental Media 

PAHs have been detected in many food products including cereal, potatoes, grain, flour, bread, 

vegetables, fruits, oils, and smoked or broiled meat and fish. The concentrations in uncooked foods 

largely depend on the source of the food. For example, vegetables and fruits obtained from a polluted 

environment may contain higher PAH concentrations than those obtained from nonpolluted 

environments. Benzo[a]pyrene, dibenz[a,h]anthracene, and chrysene have been detected in vegetables 

grown near a heavily traveled road (Wang and Meresz 1982). The method of cooking can also 

influence the PAH content of food; the time of cooking, the distance from the heat source, and the 

drainage of fat during cooking (e.g., cooking in a pan versus on a grill) all influence PAH content. 

For example, charcoal broiling increases the amounts of PAHs in meat. In a composite sample 

characterized to be typical of the U.S. diet, Howard (1979) found that PAH concentrations in all food 

groups were less than 2 ppb (μg/g). The following ranges of benzo[a]pyrene concentrations (wet or 

dry weight not specified) were summarized by Santodonato et al. (1981) from studies conducted in 

many countries: 

• cooking oils: 0.5-S ppb (μg/g) 

• margarine: 0.2-6.8 ppb 
PAHs 267 

5. POTENTIAL FOR HUMAN EXPOSURE 

• smoked fish: trace-6.6 ppb 

• smoked or broiled meats: trace-105 ppb 

• grains and cereals: not detected-60 ppb 

• fruits: not detected-29.7 ppb 

• vegetables: not detected-24.3 ppb 

These data include samples from areas identified as “polluted.” 

Gomaa et al. (1993) recently reported the results of a study to screen smoked foods, including turkey, 

pork, chicken, beef, and fish products, for carcinogenic and noncarcinogenic PAHs. Eighteen 

commercially available liquid smoke seasonings and flavorings were also evaluated. All smoked meat 

products and liquid smoke seasonings were purchased from local supermarkets in Michigan. Total 

PAH concentrations in smoked red meat products ranged from 2.6 μg/kg in cooked ham to 29.8 μg/kg 

in.grilled pork chops, while those in smoked poultry products ranged from 2.8 μg/kg in smoked turkey 

breast to 22.4 μg/kg in barbecued chicken wings. Total PAH concentrations in smoked fish products 

ranged from 9.3 μg/kg in smoked shrimp to 86.6 μg/kg in smoked salmon. Total concentrations of 

carcinogenic PAHs (benz[a]anthracene, benzo[b]fluoranthene, benzo[a]pyrene, dibenz[a,h]anthracene, 

and indeno[l,2,3,-c,d]pyrene) ranged from not detected in several red meat products to 7.4 μg/kg in 

grilled pork chops; from not detected in several poultry products to 5.5 μg/kg in barbecued chicken 

wings; and from 0.2 μg/kg in smoked trout and shrimp to 14.9 and 16.1 μg/kg in smoked oysters and 

salmon, respectively. Total PAH concentrations in liquid smoke flavorings and seasonings ranged 

from 6.3 to 43.7 μg/kg, while total carcinogenic PAH concentrations ranged from 0.3 to 10.2 μg/kg. 

Smoked meat products processed with natural wood smoke had higher total PAH and total 

carcinogenic PAH concentrations than those processed with liquid smoke flavorings. Carcinogenic 

PAHs were not detected in 10% of the smoked food samples and 24% of the samples had 

concentrations of carcinogenic PAHs <I μg/kg. Benzo[a]pyrene was not detected in 31% of the 

samples; 45% of the samples had concentrations <l μg/kg. Benzo[a]pyrene was found at 

concentrations >l μg/kg in 24% of the samples, which included pork sausage (1.8-2.3 μg/kg), grilled 

pork chops (2.5 μg/kg), whole ham (1.1 μg/kg), beef sausage (1.1 μg/kg), salmon (3.9 μg/kg), and 

oysters (3.0 μg/kg). Benzo[a]pyrene was detected in 92% (12/13) of liquid smoke flavorings and 

seasoning samples, with concentrations ranging from 0.1 to 3.4 μg/kg. 

Similar results have been obtained in recent investigations of benzo[a]pyrene concentrations in smoked 

foods in other countries. In Brazil, benzo[a]pyrene was detected in 52% (23/44) of smoked meat 
PAHs 268 

5. POTENTIAL FOR HUMAN EXPOSURE 

samples; concentrations ranged from 0.1 to 5.9 μg/kg and were generally <1.0 μg/kg (Yabiku et al. 

1993). In France, benzo[a]pyrene concentrations in smoked fish, poultry, and pork products were 

found to range from <0.2 to 1.9, 0.3 to 1.9, and <0.2 to 7.2 μg/kg, respectively; 36% (26/71) of the 

samples analyzed had benzo[a]pyrene concentrations >1 μg/kg (Moll et al. 1993). Because many 

imported food products are included in the U.S. food supply, these data may be relevant to estimating 

dietary PAH exposures of the general U.S. population. 

In data summarized by Edwards (1983) the maximum total concentration of PAHs in vegetation near 

a source was 25,000 ppb (25 μg/g) (dry weight), while concentrations in nonsource areas ranged from 

20 to 1,000 ppb (0.02-l .0 μg/g). In general, concentrations in leaves, stems, and fruits were higher 

than those in roots. Fluoranthene, pyrene, and chrysene/triphenylene were found in concentrations of 

1.2, 2.0, and 2.9 μg/g, respectively, in composite samples of green leaves from 62 plant species in the 

Los Angeles area; corresponding values for dried leaf samples were 0.47, 1.l, and 1.9 μg/g (Rogge et 

al. 1993d). Edwards (1983) reported that washing removed a maximum of 25% of PAHs on the 

leaves of plants. 

PAHs have been found in the tissues of aquatic organisms. In an assessment of STORET data 

covering the period 1980-1982, Staples et al. (1985) reported median concentrations in biota of 

<2.0 mg/kg (ppm) wet weight for 8 PAHs (acenaphthene, acenaphthylene, benz[a]anthracene, 

benzo[a]pyrene, chrysene, fluoranthene, fluorene, and pyrene) and <2.5 mg/kg wet weight for seven 

PAHs (anthracene, benzo[b]fluoranthene, benzo[k]fluoranthene, benzo[g,h,i]perylene, indenopyrene, 

naphthalene, and phenanthrene). The number of samples ranged from 83 (naphthalene) to 

140 (acenaphthylene); only benzo[g,h,i]perylene (1 sample, 0.8%) and indenopyrene (1 sample, 0.8%) 

were found in detectable concentrations. 

In summary of data on tissue contamination in mussels and oysters from the first 3 years (1986-1988) 

of the National Oceanic and Atmospheric Administration (NOAA) Mussel Watch Project, which 

involved the analysis of samples from 177 coastal and estuarine U.S. sites, overall mean concentrations 

of low molecular weight PAHs ranged from not detected (detection limits 3.3-67 ng/g dry weight) to 

4,200 ng/g dry weight (NOAA 1989). Mean concentrations of low molecular weight PAHs for 

individual years 1986, 1987, and 1988 ranged up to 9,600, 3,200, and 4,300 ng/g dry weight, 

respectively. Overall mean concentrations of high molecular weight PAHs ranged from not detected 

(detection limits 3.9-47 ng/g dry weight) to 11,000 ng/g dry weight. Mean concentrations of high 
PAHs 269 

5. POTENTIAL FOR HUMAN EXPOSURE 

molecular weight PAHs for 1986, 1987, and 1988 ranged up to 15,000, 10,000, and 11,000 ng/g dry 

weight, respectively. States with sites ranked among the highest five for concentrations of low 

molecular weight PAHs in 1986, 1987, 1988, and overall in 1986-1988, were California, Florida, 

Hawaii, Massachusetts, Mississippi, New York, Oregon, and Washington; for high molecular weight 

PAHs these states were California, Florida, Hawaii, Massachusetts, New York, and Washington. No 

consistent trends over the 1986-88 period were observed in the data; significant increases and 

decreases in concentrations of low and high molecular weight PAHs were observed with almost equal 

frequency. Low molecular weight PAH concentrations showed significant increasing trends at a single 

site each in New York, New Jersey, Florida, Texas, California, and Oregon; significant decreasing 

trends were observed at a single site each in Massachusetts, Maryland, and Mississippi, and at two 

sites in Texas. High molecular weight PAH concentrations showed significant increasing trends at a 

single site each in New York and Washington, and at two sites in Florida; significant decreasing trends 

were observed at a single site each in Connecticut, Maryland, and Florida. 

Concentrations of phenanthrene and total PAHs ranged from 2 to 296 and 63 to 2,328 μg/kg (ng/g) 

wet weight, respectively, in caged mussels (Elliptio complanata) after 3 weeks’ exposure at various 

locations in St. Mary’s River, which is heavily contaminated from industrial and municipal discharges 

in the Sault Ste. Marie, Ontario, area (Kauss 1991). PAH concentrations ranging from approximately 

50 ng/g wet weight for acenaphthylene to 4,660 nglg wet weight for fluoranthene were found in the 

digestive glands of the American lobster (Homerus americancus) collected in the proximity of a coalcoking plant that had been closed for a decade (King et al. 1993). Benzo[a]pyrene concentration was 

reported to be 720 nglg wet weight. 

In a study to evaluate the concentrations of PAHs in various fish and shellfish species from Prince 

William Sound, Alaska, following the 1989 Exxon Valdez spill of more than 10 million gallons of 

crude oil, PAHs were not detected in 18% (72/402) of the samples; trace levels were found in 78% 

(312/402) of the samples; and individual PAH concentrations ranging from 5 to 12 μg/kg (wet or dry 

weight not specified) were found in 4% (18/402) of the samples. There was no apparent difference 

between PAH concentrations in salmon collected from impacted areas and those collected from control 

areas; however, there was a suggestion that contamination may be increasing with time. No PAHs 

were detected in 14% (31/221) of samples collected in 1989, trace levels were found in 85% of these 

samples, and only 1% (3 samples) had individual PAH concentration >5 μg/kg; whereas in the 1990 

 

 

PAHs 271 

5. POTENTIAL FOR HUMAN EXPOSURE 

per cigarette (Hoffmann and Hecht 1990). Concentrations of PAHs in sidestream smoke are 

significantly higher than in mainstream smoke with sidestreammainstream concentration ratios for 

benzo[a]pyrene ranging from 2.5 to 20 (Adams et al. 1987; Evans et al. 1993; Grimmer et al. 1987; 

Hoffmann and Hoffmann 1993; IARC 1983). Benzo[a]pyrene concentrations of 0.42-63 ppb (ng/g) 

have been reported in snuff (Brunnemann et al. 1986). 

PAH concentrations in a variety of other media have been evaluated. PAH concentrations in fly ash 

and bottom ash samples from domestic municipal incinerators ranged from not detected to 

7,400 μg/kg, with phenanthrene the most abundant and frequently detected compound (Shane et al. 

1990). Machado et al. (1993) reported the total concentrations of 16 PAHs (all PAHs in this profile 

except benzo[e]pyrene) in asphalt and coal tar pitch to be 50, 122 and 294, 300 ppm μg/g), 

respectively; benzo[a]pyrene concentrations were <6 and 18,100 ppm, respectively. The concentrations 

of benzo[a]pyrene (250-480 ppm) and several other PAHs in coal tar fumes were higher than those in 

asphalt fumes by two to three orders of magnitude. The PAH content of asphalt and coal tar pitch 

fumes increased with increasing generation temperature. 

Tire wear particles, brake lining particles, and paved road dust from a residential area had total PAH 

concentrations of 226.1, 16.2, and 58.7 μg/g, with maximum concentrations of individual PAHs of 

54.1 μg/g (pyrene), 2.6 μg/g (benzo[g,h,i]perylene), and 9.4 μg/g (pyrene), respectively (Rogge et al. 

1993c). Benzo[a]pyrene concentrations in these media were 3.9, 0.74, and 2.3 μg/g, respectively. 

Combined particle- and vapor-phase emissions from scrap tire fires have been reported to contain 

average total PAH concentrations of 3.2 mg/m3

, with average benzo[a]pyrene concentrations ranging 

from 0.07 to 0.08 mg/m3

 (Lemieux and Ryan 1993). Tire pyrolysis oil, which may be used as a fuel, 

contains high levels of PAHs, with average total PAH concentrations ranging from 14,540 ppm (μg/g) 

to over 100,000 ppm (10%); benzo[a]pyrene concentrations ranged from <10 to 600 ppm (Williams 

and Taylor 1993). 

5.5 GENERAL POPULATION AND OCCUPATIONAL EXPOSURE 

The greatest sources of exposure to PAHs for most of the U.S. population are active or passive 

inhalation of the compounds in tobacco smoke, wood smoke, and contaminated air, and ingestion of 

these compounds in foodstuffs. Smoking one pack of cigarettes a day has been estimated to result in 

exposure to carcinogenic PAHs of up to 5 μg/day (Menzie et al. 1992) and in exposure to 


PAHs 274 

5. POTENTIAL FOR HUMAN EXPOSURE 

ingestion of unprocessed grains and cooked meats. This dietary intake estimate was increased to 

6-9 μg/day for individuals consuming diets with a large meat content as a result of the additional 

contribution from charcoal-cooked or smoked meats and fish. Exposure via inhalation of ambient air 

was estimated to be 0.16 μg/day (median), with a range of 0.02-3 μg/day, assuming an inhalation rate 

of 20 m3/day. Smoking one pack of unfiltered cigarettes per day increases this estimate by an 

additional 2-5 μg/day; chain smokers consuming three packs per day increase their exposure by an 

estimated 6-15 μg/day. Exposure to carcinogenic PAHs for the typical adult male from ingestion of 

drinking water and incidental ingestion of soil is minor compared to other potential routes of exposure. 

Drinking water exposure was estimated to be 0.006 μg/day (median), with a range of 

0.0002-0.12 μg/day, assuming a consumption rate of 2 L/day. Assuming incidental ingestion of 

50 mg soil/day, which may be more typical for small children than for most adults, the estimated 

median soil intake of carcinogenic PAHs was 0.06 μg/day (range, 0.003-0.3 μg/day). Therefore, the 

total potential exposure of carcinogenic PAHs for adult males was estimated to be 3 μg/day (median), 

with a maximum value of 15 μg/day. Smokers of nonfiltered cigarettes may experience exposures 

twice as high as these estimates. Ingestion of food appears to be the main source of exposure to 

PAHs for nonsmokers, although inhalation of ambient air is also an important route. 

In a Dutch market-basket survey conducted from 1984 to 1986, the mean daily dietary intake of PAHs 

by 18-year-old males in composites of 221 different foods from 23 commodity groups was estimated 

to range from 5 to 17 μg/day. The most frequently detected PAHs were benzo[b]fluoranthene (59% of 

samples), fluoranthene (48%), and benzo[k]fluoranthene (46%). The largest contribution of PAHs to 

the total diet came from the sugar and sweets, cereal products, and oils, fats, and nuts commodity 

groups (de Vos et al. 1990). 

Consumption of Great Lakes fish is not expected to contribute significantly to dietary intake unless the 

fish are smoked (Environment Canada 1991). The estimated exposure from consuming 114 g fish 

containing 50 ppt (ng/g) benzo[a]pyrene once a week would be 5.7 ng/person/week, or 11.6 pg/kg of 

body weight per day for a 70-kg individual. However, in some areas of the United States, fish 

consumption advisories have been issued based on elevated concentrations of PAHs found in locally 

caught fish or shell fish (see Section 5.6) (RTI 1993). 

The average intake of benzo[g,h,i]perylene by adults from drinking water sources has been estimated 

to be 2 ng/day (assuming a drinking water ingestion rate of 2 L/day), inhalation exposure to the 
PAHs 275 

5. POTENTIAL FOR HUMAN EXPOSURE 

compound has been estimated to be 10 ng/day (assuming an air intake rate of 20 m3/day) (EPA 

1989a). 

Indoor air can be an important source of human exposure to PAHs. Potential indoor combustion 

sources of PAHs include tobacco smoke, unvented space heaters, and food preparation (Lioy and 

Greenberg et al. 1990). PAHs are among the major carcinogenic agents in environmental tobacco 

smoke (ETS), which is comprised primarily of diluted sidestream smoke, with a much smaller 

contribution from exhaled mainstream smoke (Hoffmann and Hoffmann 1993). Exposure to ETS is of 

particular concern because it has recently been declared a human lung carcinogen by the U.S. EPA. 

Concentrations of some PAHs in cigarette smoke-polluted environments are listed in Table 5-5. 

Chuang et al. (1991) monitored the concentrations of PAHs in the indoor and outdoor air of eight 

homes in Columbus, Ohio, in the winter of 1986-87 and obtained the indoor air results shown in 

Table 5-6. 

Environmental tobacco smoke was the most significant influence on indoor air PAH levels; homes 

occupied by smokers had higher average concentrations of most PAHs than homes occupied by 

nonsmokers. In homes occupied by nonsmokers, the.highest average concentrations of most PAHs 

were found in homes that had gas cooking and heating appliances, followed by homes with gas 

heating and electric cooking appliances. Homes equipped with electric cooking and heating had the 

lowest average concentrations of most PAHs. 

The Total Human Environmental Exposure Study (THEES), a multimedia study of human exposure to 

benzo[a]pyrene, was conducted in a rural town, Phillipsburg, New Jersey, where the major industry 

was a grey-iron pipe manufacturing plant that contributed to high levels of benzo[a]pyrene in the 

ambient atmosphere (Butler et al. 1993; Lioy 1990; Lioy et al. 1988; Waldeman et al. 1991). 

Benzo[a]pyrene concentrations in respirable particulate personal samples from 10 homes in areas near 

the foundry were measured in the range of 0.1 to 8.1 ng/m3

, depending on personal habits (whether 

the windows were kept open, how frequently the doors were opened, cooking methods, hobbies, 

whether home improvements were being made) and sources of home heating. The mean outdoor air 

concentration of benzo[a]pyrene was 0.9 ng/m3

. In samples of food collected from family meals over 

a 2-week period, the concentration level of benzo[a]pyrene ranged from 0.004 to 1.2 ng/g (wet 

weight)., No detectable amounts of benzo[a]pyrene were observed in the drinking water supply 

(detection limit, 0.1 ng/L). In comparing the inhalation and ingestion pathways in each home, Lioy et 

PAHs 277 

5. POTENTIAL FOR HUMAN EXPOSURE 

al. (1988) found that potential intake could be similar in each medium. Of the 20 weeks of exposure 

(10 homes over a 2-week exposure period), 10 had higher food benzo[a]pyrene exposures and 10 had 

higher inhalation benzo[a]pyrene exposures. The range of estimated food exposures 

(l0-4,005 ng/week) was much greater than the range of estimated air exposures (78-385 ng/week). 

The dominance of one pathway or the other seemed to depend on indoor combustion sources (e.g., 

cigarette smoke or coal-burning stoves) and personal eating habits. For smokers, inhalation of tobacco 

smoke was the main source of benzo[a]pyrene exposure; intake from this source was much higher than 

inhalation of ambient and indoor air or ingestion of food. Smokers also had higher exposure through 

food intake than nonsmokers. For the average nonsmoker, ingestion of food was the most important 

route of exposure (Lioy 1990). 

Occupational exposures to PAHs can result from processes such as petroleum refining, metalworking, 

the production of coke, the manufacture of anodes, and the production of aluminum. In reviewing the 

available data on occupational exposures, it is important to understand that it cannot be implied that 

the results of a study at one industrial site would be valid for another site or for the same site at 

another time. 

Occupational exposures to PAHs are possible in all operations involved in extraction and processing of 

crude oil, including drilling, pumping and treating, transport, storage, and refinement (Suess et al. 

1985). The main route of exposure is inhalation, although there is also potential for significant dermal 

exposure (IARC 1989). Workers in petroleum refineries are exposed to PAHs from a variety of 

sources, including atmospheric distillation, catalytic cracking, residual fuel oil, lubricant oil processing, 

bitumen processing and loading, coking, and waste-water treatment (IARC 1989). In a study of nine 

US refineries, total PAH concentrations of 10 μg/m3

 were reported in personal samples taken in the 

fluid catalytic cracking and delayed coker units (Futagaki 1983). Total PAH concentrations ranging 

from approximately 1 to 40 μg/m3

 were observed in area samples from bitumen processing units. 

Total PAH concentrations in personal samples from the de-asphalting unit in one refinery were found 

to range from 2.5 to 49.8 μg/m3

. At least 85% of the total PAH in these samples was comprised of 

two-ring compounds (i.e., naphthalene and its derivatives) and 94% of two- or three-ring compounds. 

PAHs with five or more rings were found to contribute from <0.1% at the catalytic cracker unit to 

2.5% at the delayed coker unit. The highest combined concentration of benzo[a]pyrene and 

benzo[e]pyrene was 9.3 μg/g3

 in a personal sample from a coker cutter; however, these two PAHs 

were not detected in most samples (detection limit, 0.01 μg/m3). Exposures to four- to six-ring PAHs 
PAHs 278 

5. POTENTIAL FOR HUMAN EXPOSURE 

of <0.1 μg/m3 have been associated with loading road tankers with bitumen in refineries (Brandt and 

Molyneux 1985). In an evaluation of turn-around operations on reaction and fractionator towers, 

concentrations of anthracene, benzo[a]pyrene, chrysene, and pyrene in personal samples were either 

too low to be detected or ≤1 μg/m3; naphthalene and its methyl derivatives accounted for >99% of the 

total PAH measured in personal samples (Dynamac Corp. 1985). Area samples taken at various sites 

during shut-down, leak testing and start-up operations after turn-arounds showed the same distribution 

pattern of individual PAHs with total PAH concentrations generally ≤100 μg/m3 (maximum 

400 μg/m3). 

Metalworkers may also be exposed to PAHs from refined mineral oils used in machining operations, 

with the level of exposure depending on the type of oil refinement procedure used (IARC 1984). Acid 

refined mineral oils have a significant PAH content and have been shown to cause skin cancer in 

workers exposed to them (Jarvholm and Easton, 1990). Solvent refining procedures almost completely 

remove PAHs from mineral oils and, therefore, should almost completely eliminate the risks of 

exposure to carcinogenic PAHs (Bingham et al. 1965; Doak et al. 1983; IARC 1984). There is 

evidence, however, that the concentration of carcinogenic PAHs in solvent-refined cutting oils may 

increase during use, particularly during operations such as quenching where the oil is severely heated 

(Agarwal et al. 1986; Apostoli et al. 1993; IARC 1984; La Fontaine 1978; Thony et al. 1976). Total 

PAH concentrations in air samples from work areas related to the use of cutting, hardening, and 

extruding oilshave been reported to be 66, 90, and 106 ng/m3, respectively (Apostoli et al. 1993). 

In a summary of data on industrial exposures in IARC (1984), concentrations of airborne 

benzo[a]pyrene in aluminum production facilities from 1959 to 1982 in several countries ranged from 

not detected to 975 μg/m3

; concentrations in a U.S. aluminum reduction plant ranged from 0.03 to 

53.0 μg/m3, depending on the work site. Concentrations of airborne benzo[a]pyrene in coke oven 

operations ranged from 0 to 383 μg/m3

 depending on the work area; average concentrations in a U.S. 

plant were reported to range from 0.15 to 6.72 μg/m3. More recently, inhalation exposures to 

phenanthrene, pyrene, and benzo[a]pyrene of 485, 108, and 48 μg/8 hours, respectively, have been 

reported for workers in a German coke plant (Grimmer et al. 1994). However, Van Rooij et al. 

(1993b) has recently concluded that among coke oven workers dermal absorption is a major route of 

exposure to PAHs, accounting for an average of 75% and 51% of total absorbed pyrene and 

benzo[a]pyrene, respectively; the mean dermal and respiratory uptakes of pyrene in 12 workers were 

reported to range from 4 to 34 and 0.5 to 32 μg/day, respectively. 
PAHs 279 

5. POTENTIAL FOR HUMAN EXPOSURE 

The use of coal tar, pitch, asphalt, creosote, soot, and anthracene oil is widespread in the manufacture 

of fuel, dyes, plastics, paints, insulating materials, impregnating materials, building materials, 

road-building materials, embedding material, rubber, inks, and brushes (Hueper 1949). Faulds et al. 

(1981) faund that PAHs in diesel engine exhaust attach to respirable dust particles and travel long 

distances in underground mines, resulting in exposure of mine workers far removed from engine sites. 

Mechanics may also be exposed to PAHs resulting from the pyrolytic decomposition of the organic 

fractions of abraded particles from clutch and brake linings (Knecht et al. 1987). Potential exposure to 

PAHs in road sealing work involving coal tar and bitumen was discussed by Darby et al. (1986). In a 

study to evaluate inhalation and dermal exposures of 10 roofers removing an old coal tar pitch roof 

and applying a new asphalt roof, the PAH content of forehead skin wipes taken at the end of the 

workshift (0.097 μg/cm2

; equivalent to an estimated daily skin exposure of 19.4 μg/day) was found to 

correlate with the PAH concentrations in personal air samples (10.2 μg/m3

) (Wolff et al. 1989c). 

Relative concentrations of PAHs in air and wipe samples were: fluoranthene > pyrene > 

benz[a]anthracene > benzo[a]pyrene > benzo[b]fluoranthene > benzo[g,h,i]perylene > 

benzo[k]fluoranthene. Anthracene was found in the air samples but was not detected in the wipe 

samples. 

Data on PAH exposures in the United States for many other occupations are limited. Most of the 

recent studies have been conducted in other countries. PAH concentrations (combined particle- and 

vapor-phase) in two work areas in a silicon carbide plant ranged from not detected (detection limit, 

0.01 mg/m3

) for benzo[a]pyrene and benzo[e]pyrene, to 0.99 mg/m3 for fluoranthene and 3.46 mg/m3 

for naphthalene (Dufresne et al. 1987). Higher ambient concentrations were observed in the furnace 

area, ranging from 0.04 mg/m3

 for benzo[a]pyrene and benzo[e]pyrene to 3.85 mg/m3 for fluoranthene 

and 58.0 mg/m3

 for naphthalene. Total PAH concentrations in areas near cooking fume sources in the 

food and catering industries in Finland have been reported to range from 0.2 to 31.8 μg/m3 (Vainiotalo 

and Matveinen 1993). Benzo[a]pyrene concentrations of <l-44 ng/m3

 in the breathing zone of urban 

bus drivers in France have been reported (Limasset et al. 1993). The mean particulate total PAH 

exposure level in a Swedish electrode paste plant was found to vary from 4.3 to 84.6 μg/m3

 over 

various work operations, with an overall mean particulate PAH exposure level in the plant of 

14.4 μg/m3

 (Ovrebo et al. 1994). The mean PAH exposure level in a Belgian graphite electrode plant 

was reported to be 19.7 μg/m3

 (Van Hummelen et al. 1993). Drivers of large diesel-powered trucks in 

Switzerland were not found to have exposures to total PAHs or benzo[a]pyrene that were significantly 

different from controls (Guilleman et al. 1992). 

PAHs 281 

5. POTENTIAL FOR HUMAN EXPOSURE 

therapeutically (i.e., psoriatic patients) exposed to coal tar, as compared to unexposed subjects. 

Weston et al. (1994) reported similar results for coal-tar treated psoriasis patients. Tolos et al. (1991) 

and Van Rooij et al. (1994) reported significant increases in the average 1-hydroxypyrene 

concentrations in the urine of smokers over nonsmokers. Significant increases in urinary 

1-hydroxypyrene levels have also been observed in children living in areas of high density automobile 

traffic over those of children in suburban areas (Kanoh et al. 1993). Several researchers have reported 

substantial increases in the urinary concentrations of 1-Hydroxypyrene and other PAH metabolites 

among workers exposed to PAHs in a variety of occupational settings, including coke plants (Grimmer 

et al. 1994; Jongeneelen et al. 1990; Van Hummelen et al. 1993), graphite electrode plants (Van 

Hummelen et al. 1993), foundries (Santella et al. 1993), and creosote wood treatment plants (Viau et 

al. 1993) and during clean-up of dump sites contaminated with coal tars (Viau et al. 1993), handling of 

petroleum coke (Jongeneelen et al. 1989), and road surfacing operations (Jongeneelen et al. 1988). 

Most of these increases were statistically significant over controls. There is conflicting evidence 

regarding an exposure-response relationship between PAH exposures levels and urinary PAH 

metabolite concentrations. For example, Grimmer et al. (1994) reported a good correlation between 

PAH inhalation exposures and levels of urinary metabolites of benzo[a]pyrene, phenanthrene, and 

pyrene, whereas Jongeneelen et al. (1990) did not find a strong relationship between air monitoring 

data and urinary levels of 1-hydroxypyrene. In a study of PAH inhalation exposures of aluminum 

plant workers, Becher and Bjorseth (1983) found that the high concentrations in the occupational 

setting did not correspond to the measured concentrations of urinary PAH metabolites. The authors 

suggested that PAHs adsorbed to airborne particulate matter may not be bioavailable and that the 

exposure-uptake relationship may not be linear over the entire PAH concentration range. When 

urinary 1-hydroxypyrene excretion is used in the assessment of PAH exposure, the contributions of 

alternative routes of exposure (i.e., inhalation and dermal) and the variability in the baseline excretion 

among individual PAH metabolites due to tobacco smoking and dietary PAH intake should be taken 

into account (Van Rooij et al. 1993b, 1994). Assays for other biomarkers of PAH exposure are 

currently being developed in animal models (Singh and Weyard 1994), but have not been evaluated in 

humans. 

The detection of PAH-DNA adducts in urine, blood and other tissues by immunoassay and 

32P-postlabelling has also been used as an indicator of exposure (Harris et al. 1985; Herikstad et al. 

1993; Ovrebo et al. 1994; Perera et al. 1993; Santella et al. 1993). 
PAHs 282 

5. POTENTIAL FOR HUMAN EXPOSURE 

5.6 POPULATIONS WITH POTENTIALLY HIGH EXPOSURES 

Human exposure to PAHs is expected to be highest among certain occupational groups (e.g., 

individuals working with coal tar and its products, foundry workers, miners, chimney sweeps), 

smokers and nonsmokers living or working in close proximity to smokers, members of the general 

population who heat their homes with wood-burning stoves, individuals living in the vicinity of 

emission sources or using products containing PAHs, and people living in the vicinity of NPL sites 

where PAHs have been detected above background levels. People who consume grilled or smoked 

food may ingest high levels of these compounds. Anyone who works extensively with products such 

as roofing materials, asphalt, and other PAH-containing substances may be exposed through inhalation 

or skin contact. 

Recreational and subsistence fishers that consume appreciably higher amounts of locally caught fish 

from contaminated waterbodies may be exposed to higher levels of PAHs associated with dietary 

intake (EPA 1993b). PAH contamination has triggered the issuance of several human health 

advisories. As of September 30, 1993, PAHs were identified as the causative pollutants in five fish 

consumption advisories in three different states. This information is summarized in Table 5-7 (RTI 

1993). EPA is considering including PAHs as target analytes and has recommended that these 

chemicals be monitored in fish and shellfish tissue samples collected as part of state toxics monitoring 

programs. EPA recommends that residue data obtained from these monitoring programs be used by 

states to conduct risk assessments to determine the need for issuing fish and shellfish consumption 

advisories for the protection of the general public as well as recreational and subsistence fishers (EPA 

1993b). 

5.7 ADEQUACY OF THE DATABASE 

Section 104(i)(5) of CERCLA, as amended, directs the Administrator of ATSDR (in consultation with 

the Administrator of EPA and agencies and programs of the Public Health Service) to assess whether 

adequate information on the health effects of PAHs is available. Where adequate information is not 

available, ATSDR, in conjunction with the NTP, is required to assure the initiation of a program of 

research designed to determine the health effects (and techniques for developing methods to determine 

such health effects) of PAHs. 

PAHs 284 

5. POTENTIAL FOR HUMAN EXPOSURE 

The following categories of possible data needs have been identified by a joint team of scientists from 

ATSDR, NTP, and EPA. They are defined as substance-specific informational needs that if met would 

reduce the uncertainties of human health assessment. This definition should not be interpreted to mean 

that all data needs discussed in this section must be filled. In the future, the identified data needs will 

be evaluated and prioritized, and a substance-specific research agenda will be proposed. 

5.7.1 Identification of Data Needs 

Physical and Chemical Properties. The physical and chemical properties (Kow, Koc, vapor 

pressure, Henry’s law constant, etc.) have been sufficiently characterized for most of the 17 PAHs and 

allow prediction of their environmental fate. 

Production, Import/Export, Use, Release, and Disposal. According to the Emergency 

Planning and Community Right-to-Know Act of 1986, 42 U.S.C. Section 11023, industries are 

required to submit chemical release and off-site transfer information to the EPA. The Toxics Release 

Inventory (TRI), which contains this information for 1992, .became available in May of 1994. This 

database will be updated yearly and should provide a list of industrial production facilities and 

emissions. 

PAHs are produced primarily as a result of combustion processes both from anthropogenic and natural 

sources (HSDB 1994; IARC 1982). Of the 17 PAH compounds discussed in this profile, only 

acenaphthene, acenaphthylene, and anthracene are produced commercially. However, several other 

PAH compounds were imported into the United States in 1984 and 1985 (see Section 4.2). 

There is no known commercial use for most of the 17 PAHs discussed in this profile. Anthracene, 

acenaphthene, fluorene, and phenanthrene are chemical intermediates used in the manufacture of dyes, 

plastics, pesticides, explosives, and chemotherapeutic agents (Hawley 1987; HSDB 1992; Windholz 

1983). Fluoranthene is used as a lining material to protect the interior of steel and ductile iron 

drinking water pipes and storage tanks (NRC 1983). 

PAHs are most likely to be released directly in the atmosphere. Other contaminated media of 

relevance to human exposure include foods and drinking water. 
PAHs 285 

5. POTENTIAL FOR HUMAN EXPOSURE 

Rules governing the disposal of PAHs have been promulgated by EPA. Although information 

regarding recommended remedial techniques is available for most of the 17 PAHs discussed in this 

profile (EPA 1980, 1981a; HSDB 1992; IARC 1985), additional information about the amounts of 

these compounds disposed of by these remediation methods would be helpful in determining important 

routes of human exposure. 

Environmental Fate. The environmental fate of PAHs is well characterized. No further studies 

are needed. PAHs are transported in and partitioned to the air, water, and soil. Transformation and 

degradation processes of PAHs in the air, water, and soil have been well studied. Atmospheric halflives 

of PAHs are generally less than 30 days. Photochemical oxidation of a number of PAHs has 

been reported (EPA 1988a). The National Research Council (NRC 1983) noted that compounds 

adsorbed to soot are more resistant to photochemical reactions than pure compounds. In surface water, 

PAHs can volatilize, photodegrade, oxidize, biodegrade, bind to particulates, or accumulate in aquatic 

organisms (with bioconcentration factors often in the 100-2,000 range). Half-lives for volatilization of 

benz[a]anthracene and benzo[a]pyrene (high molecular weight PAHs) from water have been estimated 

to be greater than 100 hours (Southworth 1979), and the half-life for volatilization of anthracene (a 

low molecular weight PAH) was estimated to be 18 hours (Southworth et al. 1978). Hydrolysis is not 

considered to be an important degradation process for PAHs (Radding et al. 1976). The rate and 

extent of photodegradation varies widely among the PAHs (Neff 1979). PAHs in soil can biodegrade 

or accumulate in plants. Microbial metabolism is the major process for degradation of PAHs in soil 

environments. Photolysis, hydrolysis, and chemical oxidation are not considered important processes 

for the degradation of PAHs in soils (Sims and Overcash 1983). 

Bioavailability from Environmental Media. Limited information is available regarding the 

bioavailability of PAHs from plants grown in contaminated soils. PAHs can be absorbed following 

inhalation, oral, or dermal exposure. All of these routes are of concern to humans because PAHs have 

been shown to contaminate the air, drinking water, soil, and food. There is a need to conduct 

additional studies on the bioavailability of PAHs from plants grown in contaminated soils and from 

contaminated soils. However, bioavailability of PAHs from contaminated air, water, and food is of 

primary concern, and some information is available concerning bioavailability following exposure by 

these routes. Indirect evidence indicates that PAHs are absorbed by humans following inhalation 

exposure (Becher and Bjorseth 1983). Furthermore, indirect evidence suggests that benzo[a]pyrene 

may be absorbed following oral exposure in humans. The concentration of benzo[a]pyrene in human 
PAHs 286 

5. POTENTIAL FOR HUMAN EXPOSURE 

feces was studied after eight volunteers ingested broiled meat that contained approximately 9 μg of 

benzo[a]pyrene (Hecht et al. 1979). Less than 0.1 μg/person of benzo[a]pyrene was measured in the 

feces of these individuals, suggesting absorption and perhaps metabolism of the compound. More 

direct data are needed on the extent of bioavailability of PAHs, particularly those that are particlebound, 

following the three major routes of exposure. 

Food Chain Bioaccumulation. PAHs can bioaccumulate in plants, aquatic organisms, and 

animals from intake of contaminated water, soil, and food. Extensive metabolism of the compounds 

by high-trophic-level consumers, including humans, has been demonstrated; therefore, food chain 

biomagnification of the compounds does not appear to be significant (Edwards 1983; Eisler 1983; Gile 

et al. 1982; Wild et al. 1992). However, in some areas of the United States, fish consumption 

advisories have been issued based on elevated concentrations of PAHs found in locally caught fish or 

shellfish (see Section 5.6) (RTI 1993). Additional information is needed on levels of PAHs in aquatic 

organisms that are of concern for human health. 

Exposure Levels in Environmental Media. PAHs have been produced and used in large 

volumes in the environment, home, and industry and. are widely distributed in the environment. They 

have been detected in air, water, sediment, soil, and food. Although some studies of background 

levels in different media have been conducted, additional site-specific concentration data in the vicinity 

of hazardous waste sites are needed. Studies should focus particularly on ambient air, in order to 

estimate exposure of the general population through inhalation of contaminated air as well as ingestion 

of or dermal contact with contaminated water or soil. Levels of PAHs tend to be higher in urban air 

than in rural air (Greenberg et al. 1985; Pucknat 1981). One study reported benzo[a]pyrene air levels 

of 0.2-19.3 ng/m3

 for urban air and 0.1-0.2 ng/m3 in rural air (Pucknat 1981). Higher levels of other 

PAHs have been measured in urban areas. Basu and Saxena (1978a) reported concentrations of 

selected PAHs in surface waters used as drinking water sources in four U.S. cities (Huntington, West 

Virginia; Buffalo, New York; and Pittsburgh and Philadelphia, Pennsylvania) as ranging from 4.7 ng/L 

in Buffalo to 600 ng/L in Pittsburgh. Data collected as part of the Nationwide Urban Runoff Program 

indicate concentrations of individual PAHs in the range of 300-10,000 ng/L, with the concentrations 

of most PAHs above 1,000 ng/L (Cole et al. 1984). Few data are available on the concentrations of 

PAHs in U.S. groundwater. Basu and Saxena (1978b) reported total PAH concentrations in 

groundwater from three sites in Illinois, Indiana, and Ohio to be in the range of 3-20 ng/L. Data 

summarized by Sorrel et al. (1980) indicate low levels of PAHs in finished drinking waters of the 
PAHs 287 

5. POTENTIAL FOR HUMAN EXPOSURE 

United States. Reported maximum concentrations for total PAHs (based on measurement of 1.5 PAHs) 

in the drinking water of 10 cities ranged from 4 to 24 ng/L. PAHs have been detected in unprocessed 

cereal, potatoes , grain, flour, bread, vegetables, fruits, and refined fats and oils. The concentrations in 

uncooked foods largely depend on the source of the food. The amount of PAHs found in food 

products depends as much on the method of preparation (especially grilling, smoking, or pickling) as 

on the origin of the food. 

Reliable monitoring data for the levels of PAHs in contaminated media at hazardous waste sites and 

associated background sites are needed so that the information obtained on levels of PAHs in the 

environment can be used in combination with the known body burden of PAHs to assess the potential 

risk of adverse health effects in populations living in the vicinity of hazardous waste sites. 

Exposure Levels in Humans. No data are available regarding the levels of PAHs in body tissues 

or fluids for populations living near hazardous waste sites. PAHs and their metabolites can be 

measured in the urine of exposed individuals. In workers exposed to PAHs, the PAH metabolite 

1-hydroxypyrene has been detected in the urine at concentrations of 0-40 μg/g creatinine (Jongeneelen 

et al. 1985). No correlation was found between occupational exposure levels and urine levels, so it is 

not known whether-urine metabolites could be detected following exposure to low levels of PAHs (as 

might be expected to occur in individuals living in the vicinity of hazardous waste sites). 

PAHs have generally not been detected in surveys of human tissue, presumably because the 

compounds are fairly rapidly metabolized. Phenanthrene was the only PAH detected in the 1982 

National Human Adipose Tissue Survey; it was found in trace concentrations in 13% of the samples 

(EPA 1986). Acenaphthylene, acenaphthene, fluorene, and chrysene were not found at levels below 

the detection limit (0.010 μg/g; 10 ppt). However, autopsies performed on cancer-free corpses found 

PAH levels of 11-2,700 ppt (ng/g) in fat samples (Obana et al. 1981). Several PAHs were detected, 

including anthracene, pyrene, benzo[e]pyrene, benzo[k]fluoranthene, benzo[a]pyrene, and 

benzo[g,h,i]perylene, with pyrene being detected in the highest concentrations. A similar study done 

on livers from cancer-free patients found levels of 6-500 ppt of all of the same PAHs except 

benzo[e]pyrene, which was not detected (Obana et al. 1981). As in the fat sample studies, pyrene 

appeared in the highest concentrations in the liver, but the overall levels were less than in fat. 
PAHs 288 

5. POTENTIAL FOR HUMAN EXPOSURE 

A few exposure estimates for the general population have been made from inhalation of ambient air 

and ingestion of contaminated drinking water and food, but sparse monitoring data limit the reliability 

of these estimates. Relatively recent estimates of the size of the workforce exposed to a few of the 

PAHs (such as benzo[a]pyrene) are available from NIOSH. However, monitoring data on workplace 

exposure levels are generally inadequate, partially because of the complexity of air emissions in terms 

of number of compounds detected. Information on exposure levels in humans is needed to better 

define exposure estimates in the general population and workforce, and to examine the relationship 

between levels of PAHs in the environment, human tissue levels, and the subsequent development of 

health effects. These data should be collected simultaneously with data on levels of PAHs air, water, 

and soil. For a sound database to serve as a foundation for higher level environmental or toxicological 

research, it should contain information on human exposure levels to PAHs, particularly for individuals 

living near hazardous waste sites. 

This information is necessary for assessing the need to conduct health studies on these populations. 

Exposure Registries. No exposure registries for PAHs were located. These substances are not 

currently compounds for which a subregistry has been established in the National Exposure Registry. 

The substances will be considered in the future when chemical selection is made for subregistries to be 

established. The information that is amassed in the National Exposure Registry facilitates the 

epidemiological research needed to assess adverse health outcomes that may be related to exposure to 

this substance. 

5.7.2 Ongoing Studies 

The National Institute of Environmental Health Sciences is funding research at Miami University, 

Oxford, Ohio, to evaluate the transfer of benzo[a]pyrene from sediments directly to a sediment-feeding 

fish. Specific objectives are to (1) to examine the effects of seasonal parameters (i.e., temperature, 

body lipids, gonadal development) on the rate and pattern of metabolism of benzo[a]pyrene in a range 

of size, age, and sexual maturity classes; and (2) evaluate the relative importance of sediment and 

water as vectors of uptake of benzo[a]pyrene. 

The National Institute of Environmental Health Sciences is funding research at the State University of 

New York at Albany to determine the aquatic bioavailability of PAHs from sediments collected near 
PAHs 289 

5. POTENTIAL FOR HUMAN EXPOSURE 

discharge points from two aluminum manufacturing plants in the Massena area of the St. Lawrence 

River. The results of this study should provide a better understanding of the relative contributions of 

individual industries to pollution of fish and wildlife consumed by area residents. 

The National Institute of Environmental Health Sciences is funding research at the State University of 

New York at Albany to conduct an epidemiologic study of Mohawk women and infants to test the 

hypotheses that exposure to PAHs, polychlorinated biphenyls (PCBs), polychlorinated dibenzo-pdioxins 

and dibenzofurans (PCDD/Fs) from a nearby Superfund hazardous waste site elevates body 

burdens and affects the cytochrome P-450-dependent mixed function monooxygenase system. 

Determination of 15 PAHs in breast milk will be among the analyses included in this study. The 

results of the study should enhance our understanding of how these important classes of chemicals 

arising from hazardous waste bioaccumulate in human adults and infants and whether P-450IA2 

induction is a sensitive biomarker of their early biologic effect. 

The National Institute of Environmental Health Sciences is funding research at Johns Hopkins 

University to investigate the molecular dosimetry of ingested PAHs from cooked meats in humans, and 

identify susceptibility factors that modulate the formation of DNA and protein adducts with these 

dietary carcinogens. Ultimately, molecular biomonitoring may allow quantitation of biological dose 

from ingested PAHs, thus accounting for variation in exposure, cooking processes, and metabolism. 

The National Institute of Environmental Health Sciences is funding the Massachusetts Institute of 

Technology to continue studies with fluoranthene by using a modification of a 32P-postlabelling assay 

to detect and quantify DNA adducts in mice and human lymphoblast cell lines from eventual 

application as a dosimeter to monitor DNA damage in tissues from occupationally or environmentally 

exposed human populations. 

PAHs 291 

6. ANALYTICAL METHODS
 

The purpose of this chapter is to describe the analytical methods that are available for detecting, and/or 

measuring, and/or monitoring PAHs, its metabolites, and other biomarkers of exposure and effect to 

PAHs. The intent is not to provide an exhaustive list of analytical methods. Rather, the intention is to 

identify well-established methods that are used as the standard methods of analysis. Many of the 

analytical methods used for environmental samples are the methods approved by federal agencies and 

organizations such as EPA and the National Institute for Occupational Safety and Health (NIOSH). 

Other methods presented in this chapter are those that are approved by groups such as the Association 

of Official Analytical Chemists (AOAC) and the American Public Health Association (APHA). 

Additionally, analytical methods are included that modify previously used methods to obtain lower 

detection limits, and/or to improve accuracy and precision. 

6.1 BIOLOGICAL SAMPLES 

Several analytical techniques have been used to determine trace levels of PAHs in biological tissues 

and fluids including adipose tissue, lungs, liver, skin, hair, blood, urine, and feces (Table 6-l). These 

include gas chromatography coupled with flame ionization detection (GC/FID), gas chromatography 

coupled with a mass spectrometry (GC/MS), high-performance liquid chromatography (HPLC) coupled 

with an ultraviolet (UV) or fluorescence detector, and thin-layer chromatography (TLC) with 

fluorescence detection. 

Recently, Liao et al. (1988) developed a relatively simple and rapid procedure for purifying human 

and bovine adipose tissue extracts so that trace levels of complex mixture of target analytes (including 

PAHs) could be detected and quantified by capillary GC/MS. By employing an activated Florisil 

column, Liao and co-workers showed that lipid contaminants bind effectively (more than 99.75%) with 

Florisil, thereby producing a relatively clean sample extract. A detection limit at a low ng/g level and 

an average sample recovery of 85% were achieved (Gay et al. 1980; Liao et al. 1988; Modica et al. 

1982). 










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6. ANALYTICAL METHODS 

Obana et al. (1981) reported the identification and quantification of six PAHs on EPA’s priority 

pollutant list: anthracene, pyrene, benzo[b]fluoranthene, benzo[k]fluoranthene, benzo[a]pyrene, and 

benzo[g,h,i]perylene using the HPLC/fluorescence detector technique. Levels measured in human. 

tissue ranged from 0.006 to 0.460 ng/g. Following extraction of the PAHs from the sample matrices 

by saponification with KOH, the extract was cleaned on alumina and silica gel columns, prior to 

quantitation. The known carcinogens, benz[a]anthracene and dibenz[a,h]anthracene, were not detected 

(detection limit <0.005 ng/g). The HPLC/UV detection technique has also been used to 

simultaneously determine fluoranthene, benz[a]anthracene, and pyrene in blood and lung tissues 

(Brandys et al. 1989). A detection limit of ppb (ng/g or ng/mL), satisfactory recoveries (65-109%), 

and adequate precision (119% relative standard deviation [RSD]) were achieved (Brandys et al. 1989). 

In addition to direct measurement of PAHs in biological tissues, it is also possible to determine the 

concentration of metabolites in biological fluids. Pyrene is predominantly excreted as a 1-hydroxypyrene 

conjugate (glucoronate and sulfate), although 1,2-dihydroxy- 1,2-dihydropyrene conjugates are 

.also excreted in urine (Grimmer et al. 1993). Phenanthrene, on the other hand, is mainly excreted as 

dihydrodiol conjugates. The metabolites of phenanthrene that have been detected in human urine are 

1-hydroxyphenanthrene, 2-hydroxyphenanthrene, 3-hydroxyphenanthrene, 4-hydroxyphenanthrene, 

9-hydroxyphenanthrene, 1,2-dihydroxy-1,2-dihydrophenanthrene, 3,4-dihydroxy-3,4-dihydrophenanthrene, and 9,10-dihydroxy-9,10-dihydrophenanthrene (Grimmer et al. 1993). There are 

apparently individual variations in the phenanthrol (hydroxyphenanthrene) and phenanthrene 

dihydrodiol conjugates excreted in the 24-hour urine sample (Grimmer et al. 1993). The major 

metabolite of benzo[a]pyrene in human tissue and body fluid is 7,8,9,10-tetrahydroxy-7,8,9,10­

tetrahydrobenzo[a]pyrene (Weston et al. 1993a, 1993b). 

Becher and Bjorseth (1983, 1985) and Becher (1986) developed an HPLC method for biological 

monitoring of PAHs and PAH metabolites in the urine of humans following occupational exposure to 

PAHs. Using the HPLC/fluorescence detector technique, recoveries of the individual PAH compounds 

varied between 10 and 85% with the more volatile 3-ring PAHs having the lowest recoveries. A 

detection limit of less than 1 μg of PAHs per mmol of creatinine was obtained. HPLC equipped with 

a fluorescence detector has also been used to measure 1-pyrenol (1-hydroxypyrene, a pyrene 

metabolite) in urine of workers exposed to PAHs in coal tar pitch with a detection limit of 
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6. ANALYTICAL METHODS 

0.45 nmol/L (Tolos et al. 1990). Recovery and precision data were not reported. A strong correlation 

was observed between the concentrations of urinary 1-hydroxypyrene in workers and environmental 

PAHs, indicating that pyrene may be used as a biomarker of exposure for assessing worker exposure 

to coal tar pitch containing pyrene (Tolos et al. 1990). Since 1-Hydroxypyrene glucuronide is 

approximately 5 times more fluorescent than 1-hydroxypyrene, the former may be a more sensitive 

biomarker for PAH exposure (Strickland et al. 1994). A sensitive HPLC/synchronous fluorescence 

spectroscopic method is available for the determination of 1-hydroxypyrene glucuronide (Strickland et 

al. 1994). Hecht et al. (1979) employed an HPLC analytical technique for determining the 

concentrations of benzo[a]pyrene and its metabolites in the feces of humans and rats following 

consumption of charcoal-broiled beef. A detection limit of 0.05 μg of benzo[a]pyrene metabolites per 

gram of sample was noted with HPLC/UV detection. 

There is considerable evidence that PAHs are enzymatically converted to highly reactive metabolites 

that bind covalently to macromolecules such as DNA, thereby causing mutagenesis and carcinogenesis 

in experimental animals. Thus, benzo[ajpyrene, a prototype of the carcinogenic PAHs and the most 

thoroughly studied PAH, is activated by microsomal enzymes to 7β, 8α-dihydroxy-(9α,10α)-epoxy7,8,9,10-tetrahydrobenzo[a]pyrene (BPDE) and binds covalently to DNA, resulting in formation of 

BPDE-DNA adducts (Harris et al. 1985; Haugen et al. 1986; Uziel et al. 1987). Sensitive methods are 

available to detect PAH-DNA adducts in the blood and tissues of humans and animals. These include 

immunoassays, i.e., enzyme-linked immunosorbent assays (ELISA), radioimmunoassay (RIA), 

dissociation-enhanced lanthanide fluoroimmunoassay (DELFIA), and ultrasensitive enzyme 

radioimmunoassay (USERIA); 32P- and 35S-postlabelling with radioactivity counting; surface-enhanced 

Raman spectroscopy; and synchronous luminescence spectroscopy (SLS) (Gorelick and Wogan 1989; 

Gorelick and Reeder 1993; Harris et al. 1985; Haugen et al. 1986; Helmenstine et al. 1993; Herikstad 

et al. 1993; Lau and Baird 1991; Perera et al. 1988; Phillips et al. 1987; Schoket et al. 1993). 

The ELISA technique is used for detection of antibodies in serum bound to BPDE-DNA adducts. The 

USERIA method involves measuring the immunological response of BPDE-DNA in the presence of 

rabbit anti-serum. Several researchers have employed the immunoassay techniques for detecting 

PAH-DNA adducts at 10-l5 mol levels in the blood and tissues of humans occupationally exposed to 
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6. ANALYTICAL METHODS 

PAHs (Harris et al. 1985; Haugen et al. 1986; Newman et al. i988; Perera et al. 1988; Shamsuddin et 

al. 1985; Weston et al. 1988). 

32P-postlabelling is a highly sensitive and specific method for detecting PAH-DNA adducts in the 

blood and tissues of humans and animals (Gorelick and Wogan 1989; Phillips et al. 1988, 1987; 

Willems et al. 1991). Detection limits ranging from 0.3x10-15 mol of PAH adduct per μg of DNA 

(<1 adduct in 107 nucleotides) to <10-18 mol of adduct per ug of DNA have been achieved (Phillips et 

al. 1988, 1987; Willems et al. 1991). Further advantages of the 32P-postlabelling technique are that 

adducts do not need to be fully characterized in order to be detected, and that the method is 

particularly suited to occupational exposure to a complex mixture of PAHs. Coupling 

32P-postlabelling methodology with HPLC analysis has improved the resolution of the labeled 

nucleotides and can be used to identify and quantify specific PAH-DNA adducts such as fluorantheneDNA adducts (Gorelick and Wogan 1989). A detection limit of 0.1 femtomole (fmol) of adduct 

(3 adducts per 108 nucleotides in 1 μg DNA) has been achieved. The advantage of this method is that 

it is not limited with respect to the amount of DNA that can be analyzed; therefore, sensitivity can be 

enhanced by analyzing larger quantities of DNA. Average recovery was 10-15% at 3 adducts per 

106

 nucleotides. Recovery was greater (30-40%) from DNA containing higher levels of adducts 

(Gorelick and Wogan 1989). The 32P-postlabelling assay and a combination of thin-layer and reversephase HPLC was also used to separate DNA adducts of 6 nitrated PAHs (King et al. 1994). 

PAH-DNA adducts have also been detected and identified using [35S]phosphorothioate postlabelling 

combined with HPLC analysis (Lau and Baird 1991). The sensitivity of this assay is 1 adduct per 

108 nucleotides for a 60-μg DNA sample with an overall adduct recovery of 20%. An advantage of 

35S-postlabelling over 32P-postlabelling is that 35S has a longer half-life (87 days) than 32P (14 days). 

This allows longer storage times between labeling and adduct analysis with minimal loss in sensitivity. 

35S also has a lower radioactive decay energy than 32P, which reduces the risk of human radiation 

exposure and eliminates the need for the radioisotope-shielding equipment that is required for studies 

with high specific radioactivity. On the other hand, 35S is also less sensitive than the 32P-postlabelling 

analysis because of the lower specific activity of [35S]adenosine triphosphatase (ATP) compared to 

[

32P]ATP and because of the requirement for more radioactivity per adduct for accurate HPLC 

analysis. However, if large samples of DNA are available, the sensitivity of 35S-postlabelling/HPLC 

can be increased substantially (Lau and Baird 1991). 
PAHs 305 

6. ANALYTICAL METHODS 

HPLC/UV detection and HPLC/fluorescence detection have been used for determining concentrations 

of PAH-DNA adduct and hydrolyzed PAH-DNA adducts in biological tissues and fluids (Alexandrov 

et al. 1990; Autrup and Seremet 1986; Jongeneelen et al. 1986; Rice et al. 1985b; Rogan et al. 1990; 

Salhab et al. 1987; Shugart 1986; Shugart et al. 1983; Weston and Bowman 1991; Weston et al. 

1988). A detection limit of l0-l5 mol of tetrols per sample was achieved (Haugen et al. 1986; Shugart 

et al. 1983; Weyand and Bevan 1987a). HPLC with a fluorescence detector has been used to measure 

the stereospecific formation of benzo[a]pyrene tetrols from cytochrome P-450-dependent metabolism 

of (-)-7,8-dihydroxy-7,8-dihydrobenzo[a]pyrene to BPDE in human hair (Alexandrov et al. 1990). 

This assay is simple, requiring only three human hair follicles and a low (0.5-2 μmol) substrate 

concentration. The limit of detection is ≈0.3 fmol of tetrols (Alexandrov et al. 1990). This is a rapid 

and noninvasive method that could be used to determine an individual’s capacity to activate 

carcinogens to DNA-binding intermediates (Alexandrov et al. 1990). HPLC with fluorescence 

detection has also been used to detect BPDE-DNA adducts in human lung tissues (Weston and 

Bowman 1991). A detection limit of 6 pg benzo[a]pyrene-tetrol/mL (1 adduct in 108

 nucleotides) was 

achieved (Weston and Bowman 1991). Recoveries ranged from 26 to 66% for the procedure. 

HPLC/UV has been used to identify and quantify a benzo[a]pyrene-DNA adduct, specifically 

7-(benzo[a]pyrene-6-yl)guanine (BP-N7Gua) in urine and feces in the femtomole range (Rogan et al. 

1990). The structure of the adduct was established by fluorescence line narrowing spectrometry 

(FLNS). Recovery and precision data were not reported (Rogan et al. 1990). 

Using benzo[a]pyrene as a model carcinogen, Vahakangas et al. (1985), Haugen et al. (1986), and 

Harris et al. (1985) have developed an synchronous luminescence spectroscopy (SLS) technique for 

detecting trace levels of PAH-DNA adducts in the blood of humans occupationally exposed to high 

levels of PAHs. Vahakangas et al. (1985) detected less than 1 benzo[a]pyrene moiety per 107

 DNA 

molecules by SLS technique following in vitro acid hydrolysis of BPDE-DNA adduct. Fiber-optic 

antibody-based fluoroimmunosensor (FIS) has been used to measure DNA adducts of benzo[a]pyrene 

in biological samples such as human placenta (Tromberg et al. 1988; Vo-Dinh et al. 1991). The FIS is 

used to detect the highly fluorescent benzo[a]pyrene 7,8,9,10-tetrol (BPT) after release from the 

weakly fluorescent BPDE-DNA by mild hydrolysis. The FIS is highly specific because of the 

antigen-antibody reaction. This assay is highly sensitive, achieving a detection limit of 14x10-l8 mol 
PAHs 306 

6. ANALYTICAL METHODS 

of BPT (Vo-Dinh et al. 1991). FIS precision is adequate (6.2-15% RSD) (Tromberg et al. 1988). 

Recovery data were not reported. 

6.2 ENVIRONMENTAL SAMPLES 

One of the difficulties associated with determination of PAHs in environmental samples is the 

complexity of PAH mixture in these samples. Even after extensive and rigorous clean-up, the PAH 

fraction may contain hundreds of compounds. Analytical methods that offer combinations of good 

chromatographic resolving power and detector selectivity are usually required to quantify selected 

compounds in such mixtures. There is essentially a three-step procedure for the analysis and 

determination of PAHs in environmental samples: (1) extraction and isolation of PAHs from the 

sample matrix; (2) clean-up of the PAH mixtures from impurities and fractionation of PAH into 

subgroups; and (3) identification and quantitative determination of the individual components in each 

of these subgroups. 

The collection of PAHs from air for quantification requires special considerations. Some of the PAHs, 

especially those with lower molecular weights, exist primarily in the vapor phase while PAHs with 

higher molecular weights exist primarily in the particulate phase (Santodonato et al. 1981). Therefore, 

a combination of a particulate filter (usually glass-fiber filter) and an adsorbent cartridge (usually 

XAD-2 or polyurethane foam) is used for the collection of PAHs (Andersson et al. 1983; Harvath 

1983; Hawthorne et al. 1993). Therefore, collection methods that use either a filtration system or an 

adsorbent alone may be incapable of collecting both particulate and vapor phase PAHs. In addition, a 

few PAHs are known to be susceptible to oxidation by ozone and other oxidants present in the air 

during the collection process (Santodonato et al. 1981). 

The commonly used methods for the extraction of PAHs from sample matrices are Soxhlet extraction, 

sonication, or partitioning with a suitable solvent or a solvent mixture. Dichloromethane, cyclohexane, 

benzene, and methanol have been widely used as solvents (see Table 6-2). Supercritical fluid 

extraction (SFE) of heterogeneous environmental samples with carbon dioxide in the presence of a 

modifier, such as 5-10%. methanol or dichloromethane is preferable to the conventional extraction 

method because SFE is much less time consuming and has comparable or better PAH extraction 












PAHs 319 

6. ANALYTICAL METHODS 

recovery than the conventional methods (Burford et al. 1993; Dankers et al. 1993; Hawthorne et al. 

1993; Hill and Hill 1993). 

Column chromatography on silica, alumina, Sephadex or Florisil has been used most often for the 

clean-up and fractionation of PAHs in the sample extract (Desiderie et al. 1984; Desilets et al. 1984; 

Oehme 1983; Quilliam and Sim 1988). HPLC can also be used for the clean-up and fractionation of 

PAHs in sample extract (Readman et al. 1986). A disposable Sep-Pak cartridge with an amino 

stationary phase was used for the clean-up of benzo[a]pyrene in cigarette smoke condensate (Dumont 

et al. 1993). Some soil and sediment samples containing high amounts of sulfur may require clean-up 

on an activated copper column (Giger and Schaffner 1978). 

A variety of analytical methods has been used for determining trace concentrations of PAHs in 

environmental samples (Table 6-2). These include GC with various detectors, HPLC with various 

detectors, and TLC with fluorimetric detectors. Various detection devices used for GC quantification 

include FID, MS, Fourier transform infrared spectrometer (FT-IR), laser induced molecular 

fluorescence detector (LIMF), diode array detector (DAD), and gas phase fluorescence detector 

(GPFDA). GC/MS and HPLC with UV or spectrofluorimetric detectors are perhaps the most prevalent 

analytical methods for determining concentrations of PAHs in environmental samples. 

Oehme (1983) and Low et al. (1986) employed capillary GC coupled with negative ion chemical 

ionization MS for detecting and differentiating isomeric PAHs (including PAHs on EPA’s priority 

pollutant list). This procedure was successfully used to differentiate the isomers benzolj]fluoranthene 

and benzo[b]fluoranthene at low ppb levels in complex matrices, such as air particulate matter and 

coal fly ash. An alternative method for the elucidation of PAH isomers is GC coupled with a 

charge-exchange and chemical ionization MS (Simonsick and Hites 1985). Simonsick and Hites 

(1985) demonstrated that the structural isomers pyrene, fluoranthene, aceanthrylene and 

acephenanthrylene can be identified on the basis of their first ionization potential and (M+l)+

/M+ mass 

ion ratio. 

HPLC has been one of the most widely used analytical methods for determining PAHs in complex 

environmental samples. The development of a chemically nonpolar stationary phase for HPLC has 
 

 

PAHs 320 

6. ANALYTICAL METHODS 

provided a unique selectivity for separation of PAH isomers that are often difficult to separate by GC 

columns. For example, chrysene, benz[a]anthracene, and triphenylene are baseline resolved with a 

C-18 reverse phase column packing. A detection limit of subpicogram to picogram levels of PAHs 

per sample has been achieved by HPLC with fluorescence detector (Fox and Staley 1976; Furuta and 

Otsuki 1983; Futoma et al. 1981; Golden and Sawicki 1978; Lawrence and Weber 1984; Marcomini et 

al. 1987; Miguel and De Andrade 1989; Nielsen 1979; Risner 1988; Ton&ins et al. 1982). HPLC 

equipped with a fluorescence detector has selectively measured 10 PAHs (phenanthrene, fluoranthene, 

pyrene, benz[a]anthracene, chrysene, benzo[b]fluoranthene, benzo[k]fluoranthene, benzo[a]pyrene, 

benzo[g,h,i]perylene, and indeno[l,2,3-c,d]pyrene) in ambient air (Miguel and De Andrade 1989). 

Detection limits for the 10 PAHs were in the range of 10-50 pg and RSD was <10%. Recovery data 

were not reported. PAH levels in the ng/L range have also been successfully determined in water 

using flotation enrichment and HPLC/fluorescence detection (Xu and Fang 1988). Good recoveries 

(86-107%) were achieved, and RSD was 2.7-13.6% RSD. A quenchofluorometric detection system 

provides an inexpensive method to achieve selective detection for the fluoranthenic PAHs as a group 

(Konash et al. 1981). UV detectors have been used to measure PAHs in fats and oil samples; 

however, these detectors lacked the sensitivity and specificity of the fluorescence detectors for 

determining PAHs at low levels (ppb and lower) (Van Heddeghem et al. 1980). 

A number of less commonly used analytical techniques are available for determining PAHs. These 

include synchronous luminescence spectroscopy (SLS), resonant (R)/nonresonant (NR)-synchronous 

scan luminescence (SSL) spectrometry, room temperature phosphorescence (RTP), ultravioletresonance 

Raman spectroscopy (UV-RRS), x-ray excited optical luminescence spectroscopy (XEOL), 

laser-induced molecular fluorescence (LIMF), supersonic jet/laser induced fluorescence (SSJ/LlF), 

low-temperature fluorescence spectroscopy (LTFS), high-resolution low-temperature 

spectrofluorometry, low-temperature molecular luminescence spectrometry (LT-MLS), and supersonic 

jet spectroscopy/capillary supercritical fluid chromatography (SJS/SFC) (Asher 1984; Garrigues and 

Ewald 1987; Goates et al. 1989; Jones et al. 1988; Lai et al. 1990; Lamotte et al. 1985; Lin et al. 

1991; Popl et al. 1975; Richardson and Ando 1977; Saber et al. 1991; Vo-Dinh et al. 1984; Vo-Dinh 

and Abbott 1984; Vo-Dinh 1981; Woo et al. 1980). More recent methods for the determination of 

PAHs in environmental samples include GC-MS with stable isotope dilution calibration (Bushby et al. 

1993), capillary electrophoresis with UV-laser excited fluorescence detection (Nie et al. 1993), and 
PAHs 321 

6. ANALYTICAL METHODS 

laser desorption laser photoionization time-of-flight mass spectrometry of direct determination of PAH 

in solid waste matrices (Dale et al. 1993). 

Among the less commonly used spectroscopic methods, SLS and room temperature phosphorescence 

(RTP) are used for determining trace levels of PAHs in environmental media. Vo-Dinh (1981), 

Vo-Dinh and Abbott (1984), and Vo-Dinh et al. (1984) reported a cost-effective and relatively simple 

SLS and RTP technique for determining trace amounts of PAHs (less than 1x10-9 g per sample) in air 

particulate extracts collected at a wood-burning area. Improved selectivity is the main advantage of 

SLS and RTP over conventional luminescence or fluorescence spectroscopy. Additionally, R/N-SSL 

spectrometry has been applied to determine trace amounts of anthracene and its derivatives in 

solvent-refined coal (Lin et al. 1991). The sensitivity of N-SSL (7x10-7 M) is about two orders of 

magnitude better than that of R-SSL spectrometry (7x10-5 M). The detection limit for N-SSL is 

several times better than that of conventional fluorescence spectrometry (3x10-6 M). The better 

sensitivity comes from a higher efficiency in fluorescence collection (Lin et al. 1991). The 

combination of R- and N-SSL spectrometries provides a sensitive and selective analytical method 

because of the spectral simplicity of R-SSL and the high sensitivity of N-SSL spectrometry (Lin et al. 

1991). This spectrometric method is also applicable to other PAHs in the environment, such as 

benzo[a]pyrene in airborne particulates. 

Low temperature-molecular luminescence spectrometry (LT-MLS), SLS, and HPLC/fluorescence 

detection have been used to measure pyrene in broiled hamburger (Jones et al. 1988). A comparison 

of the three methods showed that sensitivity for all three methods was in the low-ppb range and that 

all methods were comparably reproducible (6-9% RSD). Adequate recovery (75-85%) was obtained 

from the extraction procedure for all three methods. While HPLC is the least expensive and easiest to 

operate, it has the longest analysis time (30 minutes), and it provides the least resolution of 

components. LT-MLS is the fastest technique (5 minutes), and it gives mores spectral information 

than the other two methods. SLS, with an analysis time of 15 minutes, offers no real advantages over 

LT-MLS other than cost of equipment. 

Methods 8100, 8250, and 8310 are the test methods recommended by EPA (1986) for determining 

PAHs in a variety of matrices at solid waste sites. EPA Methods 610 and 625, recommended for 
PAHs 322 

6. ANALYTICAL METHODS 

municipal and industrial waste water have been used to measure PAHs in groundwater contaminated 

by petroleum hydrocarbons at detection limits in the low-ppb range (Thomas and Delfino 1991). 

Recovery and precision data were not reported. NIOSH (1985) has recommended methods 5506 and 

5515 as the analytical methods for determining PAHs in air samples at concentrations below ppm 

level. 

6.3 ADEQUACY OF THE DATABASE 

Section 104(i)(5) of CERCLA, as amended, directs the Administrator of ATSDR (in consultation with 

the Administrator of EPA and agencies and programs of the Public Health Service) to assess whether 

adequate information on the health effects of PAHs is available. Where adequate information is not 

available, ATSDR, in conjunction with the NTP, is required to assure the initiation of a program of 

research designed to determine the health effects (and techniques for developing methods to determine 

such health effects) of PAHs. 

The following categories of possible data needs have been identified by a joint team of scientists from 

ATSDR, NTP, and EPA. They are defined as substance-specific informational needs that if met would 

reduce the uncertainties of human health assessment. This definition should not be interpreted to mean 

that all data needs discussed in this section must be filled. In the future, the identified data needs will 

be evaluated and prioritized, and a substance-specific research agenda will be proposed. 

6.3.1 Identification of Data Needs 

Methods for Determining Biomarkers of Exposure and Effect. Adequate methods are
 

available to separate and quantify PAHs in biological materials such as adipose tissue (Gay et al.
 

1980; Liao et al. 1988; Modica et al. 1982), lungs (Brandys et al. 1989; Mitchell 1979; Tomingas et
 

al. 1976; Weston and Bowman 1991; Weyand and Bevan 1987a), liver (Amin et al. 1982; Obana et al.
 

1981; Rice et al. 1985b), skin (Phillips et al. 1987; Shugart et al. 1983), hair (Alexandrov et al. 1990),
 

blood (Brandys et al. 1989; Harris et al. 1985; Haugen et al. 1986; Perera et al. 1988; Phillips et al.
 

1988; Shamsuddin et al. 1985; Shugart 1986), urine (Au&up and Seremet 1986; Becher and Bjorseth
 

1985; Rogan et al. 1990; Tolos et al. 1990; Uziel et al. 1987), and feces (Hecht et al. 1979; Rogan et
 
PAHs 323 

6. ANALYTICAL METHODS 

al. 1990). These methods include GC/FID, GCMS, HPLC, TLC, and spectrofluorometry (SF). The 

difficulties involved in recovering bound benzo[a]pyrene from feces hinder studies on absorption and 

bioavailability in humans after exposure to benzo[a]pyrene. Therefore, there is a need to develop a 

satisfactory analytical method for the determination of benzo[a]pyrene in feces. Immunoassays (i.e., 

ELISA and USERIA, 32P-arid 35S-postlabelling, SLS, and FIS) are methods currently being developed 

to detect the presence of carcinogenic PAH adducts bound covalently to macromolecules (e.g., DNA). 

The parent compound is generally measured in biological tissues, but both the parent compound and 

its metabolites can be measured in biological fluids, particularly urine. However, improved methods 

for identifying and characterizing conjugated PAH metabolites from various biological fluids would be 

useful. PAH-DNA adducts can be measured in blood, serum, and other tissues. These methods are 

accurate, precise, and sensitive enough to measure background levels in the population and levels at 

which biological effects occur. Additional quantitative information regarding the relationships between 

body and environmental levels of PAHs for both short- and long-term exposures might allow 

investigators to predict environmental exposure levels from measured body levels. 

The urinary level of 1-Hydroxypyrene has the potential to be used as a biomarker for exposure to 

PAHs, and analytical methods for the detection of the hydroxy metabolite in urine of exposed and 

non-exposed control persons are available (Ariese et al. 1993a; Jongeneelen et al. 1988; Kanoh et al. 

1993; Mercado Calderon 1993; Van Hummelen et al. 1993). The correlation coefficient between total 

PAHs in air of a coke production plant and hydroxypyrene in urine of workers was 0.77 (p<0.0001) 

(Mercado Calderon 1993). A study attempted to use benzo[a]pyrene metabolite 3-hydroxybenzo[a]pyrene in urine as a biomarker for occupational exposure to PAH (Ariese et al. 1993a). Since 

the level of 3-hydroxybenzo[a]pyrene is about 3 orders of magnitude lower than 1-hydroxypyrene, a 

sensitive method was developed to estimate levels of 3-hydroxybenzo[a]pyrene in occupational groups 

(Ariese et al. 1993a). However, no significant correlation between the metabolite and levels of 

airborne benzo[a]pyrene was found. 

The available biomarkers of effect for PAHs are not specific for effects induced by PAHs other than 

cancer or genotoxicity. PAHs form DNA adducts that can be measured in body tissues or blood 

following exposure to PAHs and mixtures that contain PAHs. The formation of benzo[a]pyrene-DNA 

adducts has been demonstrated, and this may serve as a biomarker of PAH-induced carcinogenicity. 
PAHs 324 

6. ANALYTICAL METHODS 

HPLC and immunoassays, (i.e., ELISA and USERIA, 32P-and 35S-postlabelling, SLS, and FIS) are 

sensitive, selective, and reproducible methods being developed to detect the presence of carcinogenic 

PAH adducts bound covalently to macromolecules (e.g., DNA) (Gorelick and Wogan 1989; Haugen et 

al. 1986; Lau and Baird 1991; Phillips et al. 1988; Weston and Bowman 1991). Chromosomal 

aberration and sister chromatid exchange methods were used to show that several types of cultured 

human tissue cells demonstrated positive results for benzo[a]pyrene-induced genotoxicity (Abe et al. 

1983a, 1983b; Huh et al. 1982; Lo Jacono et al. 1992; Van Hummelen et al. 1993; Weinstein et al. 

1977; Wienke et al. 1990). However, statistically significant correlation between the cytogenetic 

markers and airborne occupational PAH levels was not found (Van Hummelen et al. 1993). 

Methods for Determining Parent Compounds and Degradation Products in 

Environmental Media. Standardized methods are available that are reliable, reproducible, and 

sensitive enough to separate and quantify PAHs in air (Andersson et al. 1983; Fox and Staley 1976; 

Golden and Sawicki 1978; Harvath 1983; Karlesky et al. 1987; Majer et al. 1970; Miguel and De 

Andrade 1989; Naikwadi et al. 1987; NIOSH 1984; Oehme 1983; Tomkins et al. 1982; Matsumoto 

and Kashimoto 1985), water (Beniot et al. 1979; Desiderie et al. 1984; EPA 1986; Furuta and Otsuki 

1983; Thomas and.Delfino 1991; Thruston 1978; Xu and Fang 1988), soil and sediment (Hawthorne 

and Miller 1987a, 1987b; John and Nickless 1977; Saber et al. 1991; Szepesy et al. 1981; Tanaka and 

Saito 1988), and other media, such as food (Hecht et al. 1979; Joe et al. 1984; Jones et al. 1988; 

Krahn and Malins 1982; Krahn et al. 1988; Lawrence and Das 1986; Lawrence and Weber 1984; 

Poole et al. 1987; Toussaint and Walker 1979; Tuominen et al. 1988; Van Heddeghem et al. 1980; 

Vassilaros et al. 1982; Vreuls et al. 1991), cigarette smoke (Risner 1988; Severson et al. 1976), coal 

tar (Alben 1980; Goates et al. 1989; Low et al. 1986), and fuels (Lin et al. 1991; Monarca and Fagioli 

1981; Symons and Crick 1983.; Ton&ins and Griest 1987; Vo-Dinh et al. 1984; Woo et al. 1980). 

These methods include GC, HPLC, TLC, and others. Various detection devices used for GC 

quantification include FID, MS, FT-IR, LIMF, DAD, or GPFDA. GC/MS and HPLC are perhaps the 

most prevalent analytical methods for determining concentrations of PAHs in environmental samples. 

These methods are adequate to measure environmental levels that may be associated with adverse 

human effects. All of the available analytical methods for PAHs in soil and food items are sensitive 

down to levels of <1 ppb. 
PAHs 325 

6. ANALYTICAL METHODS 

6.3.2 Ongoing Studies 

The Environmental Health Laboratory Sciences Division of the National Center for Environmental 

Health and Injury Control, Centers for Disease Control and Prevention, is developing methods for the 

analysis of PAHs and other volatile organic compounds in blood. These methods use purge and trap 

methodology, high resolution gas chromatography, and magnetic sector mass spectrometry, which 

gives detection limits in the low parts per trillion (ppt) range. 

PAHs 327 

7. REGULATIONS AND ADVISORIES
 

The international, national, and state regulations and guidelines regarding polycyclic aromatic 

hydrocarbons (PAHs) in air, water, and other media are summarized in Table 7-1. 

An MRL of 0.6 mg/kg/day has been derived for intermediate-duration oral exposure (15-364 days) to 

acenaphthene, based on a minimal LOAEL of 175 mg/kg/day for liver weight in mice (EPA 1989c). 

An MRL of 0.4 mg/kg/day has been derived for intermediate-duration oral exposure (15-364 days) to 

fluoranthene, based on a minimal LOAEL of 125 mg/kg/day for increased liver weight in mice (EPA 

1988e). 

An MRL of 0.4 mg/kg/day has been derived for intermediate-duration oral exposure (15-364 days) to 

fluorene, based on a minimal LOAEL of 125 mg/kg/day for relative liver weight in mice (EPA 

1989e). 

Reference doses have been developed by EPA for anthracene (0.3 mg/kg/day), acenaphthene 

(0.06 mg/kg/day), fluoranthene (0.04 mg/kg/day), fluorene (0.04 mg/kg/day), and pyrene 

(0.03 mg/kg/day). No reference concentrations exists for any of the PAHs. 

PAHs are regulated under The Emergency Planning and Community Right-to-Know (EPCRA) 

standards of 40 CFR Subpart J. EPCRA requires owners and operators of certain facilities that 

manufacture, import, process, or otherwise use these chemicals to report annually their release of those 

chemicals to any environmental media. 

OSHA regulates the benzene soluble fraction of coal tar pitch volatiles and mineral oil mist, which 

contain several of the PAH compounds. Employers of workers who are occupationally exposed must 

institute engineering controls and work practices to reduce and maintain employee exposure at or 

below permissible exposure limits (PEL). The employer must use engineering and work controls, if 

feasible, to reduce exposure to or below an 8-hour time-weighted average (TWA) of 0.2 mg/m3 for 

coal tar pitch volatiles and 5 mg/m3

 for mineral oil mist. 
PAHs 328 

7. REGULATIONS AND ADVISORIES 

PAHs are regulated by the Clean Water Effluent Guidelines in Title 40, Sections 400-475, of the Code 

of Federal Regulations. For each point source category, PAHs may be regulated as a group of 

chemicals controlled as Total Toxic Organics, may have a specific regulatory limitation, or may have a 

zero discharge limitation. The point source categories for which the PAHs are controlled as a Total 

Toxic Organic include electroplating, and metal molding and casting. The point source categories for 

which the PAHs have specific regulatory limitations include organic chemicals, plastics, and synthetic 

fibers; cokemaking; and nonferrous metals manufacturing. 

Under the Resource Conservation and Recovery Act (RCRA), several PAHs are listed as hazardous 

wastes when they are discarded commercial chemical products, off-specification species, container 

residues, and spill residues (40 CFR 261.33). 










































 

______________________ 

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PAHs 455 

9. GLOSSARY
 

Acute Exposure-Exposure to a chemical for a duration of 14 days or less, as specified in the 

Toxicological Profiles. 

Adsorption Coefficient (Koc)-The ratio of the amount of a chemical adsorbed per unit weight of 

organic carbon in the soil or sediment to the concentration of the chemical in solution at equilibrium. 

Adsorption Ratio (Kd)-The amount of a chemical adsorbed by a sediment or soil (i.e., the solid 

phase) divided by the amount of chemical in the solution phase, which is in equilibrium with the solid 

phase, at a fixed solid/solution ratio. It is generally expressed in micrograms of chemical sorbed per 

gram of soil or sediment. 

Bioconcentration Factor (BCF)-The quotient of the concentration of a chemical in aquatic 

organisms at a specific time or during a discrete time period of exposure divided by the concentration 

in the surrounding water at the same time or during the same period. 

Cancer Effect Level (CEL)-The lowest dose of chemical in a study, or group of studies, that 

produces significant increases in the incidence of cancer (or tumors) between the exposed population 

and its appropriate control. 

Carcinogen-A chemical capable of inducing cancer. 

Ceiling Value-A concentration of a substance that should not be exceeded, even instantaneously. 

Chronic Exposure- Exposure to a chemical for 365 days or more, as specified in the Toxicological 

Profiles. 

Developmental Toxicity- The occurrence of adverse effects on the developing organism that may 

result from exposure to a chemical prior to conception (either parent), during prenatal development, or 

postnatally to the time of sexual maturation. Adverse developmental effects may be detected at any 

point in the life span of the organism. 

Embryotoxicity and Fetotoxicity- Any toxic effect on the conceptus as a result of prenatal exposure 

to a chemical; the distinguishing feature between the two terms is the stage of development during 

which the insult occurred. The terms, as used here, include malformations and variations, altered 

growth, and in utero death. 

EPA Health Advisory-An estimate of acceptable drinking water levels for a chemical substance 

based on health effects information. A health advisory is not a legally enforceable federal standard, 

but serves as technical guidance to assist federal, state, and local officials. 

Immediately Dangerous to Life or Health (IDLH)-The maximum environmental concentration of a 

contaminant from which one could escape within 30 min without any escape-impairing symptoms or 

irreversible health effects. 
PAHs 456 

9. GLOSSARY 

Intermediate Exposure-Exposure to a chemical for a duration of 15-364 days, as specified in the 

Toxicological Profiles. 

Immunologic Toxicity-The occurrence of adverse effects on the immune system that may result 

from exposure to environmental agents such as chemicals. 

In vitro-Isolated from the living organism and artificially maintained, as in a test tube. 

In vivo-Occurring within the living organism. 

Lethal Concentration(LO) (LCLO)-The lowest concentration of a chemical in air which has been 

reported to have caused death in humans or animals. 

Lethal Concentration(50) (LC50)-A calculated concentration of a chemical in air to which exposure 

for a specific length of time is expected to cause death in 50% of a defined experimental animal 

population. 

Lethal Dose(LO) (LDLO)-The lowest dose of a chemical introduced by a route other than inhalation 

that is expected to have caused death in humans or animals. 

Lethal Dose(50) (LD50)-The dose of a chemical which has been calculated to cause death in 50% of 

a defined experimental animal population. 

Lethal Time(50) (LT50)-A calculated period of time within which a specific concentration of a 

chemical is expected to cause death in 50% of a defined experimental animal population. 

Lowest-Observed-Adverse-Effect Level (LOAEL)-The lowest dose of chemical in a study, or 

group of studies, that produces statistically or biologically significant increases in frequency or severity 

of adverse effects between the exposed population and its appropriate control. 

Malformations-Permanent structural changes that may adversely affect survival, development, or 

function. 

Minimal Risk Level-An estimate of daily human exposure to a dose of a chemical that is likely to 

be without an appreciable risk of adverse noncancerous effects over a specified duration of exposure. 

Mutagen-A substance that causes mutations. A mutation is a change in the genetic material in a 

body cell. Mutations can lead to birth defects, miscarriages, or cancer. 

Neurotoxicity-The occurrence of adverse effects on the nervous system following exposure to 

chemical. 

No-Observed-Adverse-Effect Level (NOAEL)-The dose of chemical at which there were no 

statistically or biologically significant increases in frequency or severity of adverse effects seen 

between the exposed population and its appropriate control. Effects may be produced at this dose, but 

they are not considered to be adverse. 

Octanol-Water Partition Coefficient (Kow)-The equilibrium ratio of the concentrations of a 

chemical in n-octanol and water, in dilute solution. 
PAHs 457 

9. GLOSSARY 

Permissible Exposure Limit (PEL)-An allowable exposure level in workplace air averaged over an 

8-hour shift. 

q1*-The upper-bound estimate of the low-dose slope of the dose-response curve as determined by the 

multistage procedure. The ql* can be used to calculate an estimate of carcinogenic potency, the 

incremental excess cancer risk per unit of exposure (usually μg/L for water, mg/kg/day for food, and 

μg/m3 for air). 

Reference Dose (RfD)-An estimate (with uncertainty spanning perhaps an order of magnitude) of the 

daily exposure of the human population to a potential hazard that is likely to be without risk of 

deleterious effects during a lifetime. The RfD is operationally derived from the NOAEL (from animal 

and human studies) by a consistent application of uncertainty factors that reflect various types of data 

used to estimate RfDs and an additional modifying factor, which is based on a professional judgment 

of the entire database on the chemical. The RfDs are not applicable to nonthreshold effects such as 

cancer. 

Reportable Quantity (RQ)-The quantity of a hazardous substance that is considered reportable 

under CERCLA. Reportable quantities are (1) 1 pound or greater or (2) for selected substances, an 

amount established by regulation either under CERCLA or under Sect. 311 of the Clean Water Act. 

Quantities are measured over a 24-hour period. 

Reproductive Toxicity-The occurrence of adverse effects on the reproductive system that may result 

from exposure to a chemical. The toxicity may be directed to the reproductive organs and/or the 

related endocrine system. The manifestation of such toxicity may be noted as alterations in sexual 

behavior, fertility, pregnancy outcomes, or modifications in other functions that are dependent on the 

integrity of this system. 

Short-Term Exposure Limit (STEL)-The maximum concentration to which workers can be exposed 

for up to 15 min continually. No more than four excursions are allowed per day, and there must be at 

least 60 min between exposure periods. The daily TLV-TWA may not be exceeded. 

Target Organ Toxicity-This term covers a broad range of adverse effects on target organs or 

physiological systems (e.g., renal, cardiovascular) extending from those arising through a single limited 

exposure to those assumed over a lifetime of exposure to a chemical. 

Teratogen-A chemical that causes structural defects that affect the development of an organism. 

Threshold Limit Value (TLV)-A concentration of a substance to which most workers can be 

exposed without adverse effect. The TLV may be expressed as a TWA, as a STEL, or as a CL. 

Time-Weighted Average (TWA)-An allowable exposure concentration averaged over a normal 8-hour 

workday or 40-hour workweek. 

Toxic Dose (TD50)-A calculated dose of a chemical, introduced by a route other than inhalation, 

which is expected to cause a specific toxic effect in 50% of a defined experimental animal population. 
PAHs 458 

9. GLOSSARY 

Uncertainty Factor (UF)-A factor used in operationally deriving the RfD from experimental data. 

UFs are intended to account for (1) the variation in sensitivity among the members of the human 

population, (2) the uncertainty in extrapolating animal data to the case of human, (3) the uncertainty in 

extrapolating from data obtained in a study that is of less than lifetime exposure, and (4) the 

uncertainty in using LOAEL data rather than NOAEL data. Usually each of these factors is set equal 

to 10. 
PAHs A-l 

APPENDIX A 

USER’S GUIDE 

Chapter 1 

Public Health Statement 

This chapter of the profile is a health effects summary written in non-technical language. Its intended 

audience is the general public especially people living in the vicinity of a hazardous waste site or 

chemical release. If the Public Health Statement were removed from the rest of the document, it 

would still communicate to the lay public essential information about the chemical. 

The major headings in the Public Health Statement are useful to find specific topics of concern. The 

topics are written in a question and answer format. The answer to each question includes a sentence 

that will direct the reader to chapters in the profile that will provide more information on the given 

topic. 

Chapter 2 

Tables and Figures for Levels of Significant Exposure (LSE) 

Tables (2-l) 2-2, and 2-4) and figures (2-1, 2-2, and 2-3) are used to summarize health effects and 

illustrate graphically levels of exposure associated with those effects. These levels cover health effects 

observed at increasing dose concentrations and durations, differences in response by species, minimal 

risk levels (MRLs) to humans for noncancer end points, and EPA’s estimated range associated with an 

upper-bound individual lifetime cancer risk of 1 in 10,000 to 1 in 10,000,000. Use the LSE tables and 

figures for a quick review of the health effects and to locate data for a specific exposure scenario. The 

LSE tables and figures should always be used in conjunction with the text. All entries in these tables 

and figures represent studies that provide reliable, quantitative estimates of No-Observed-AdverseEffect Levels (NOAELs), Lowest-Observed-Adverse-Effect Levels (LOAELs), or Cancer Effect Levels 

(CELs). 

The legends presented below demonstrate the application of these tables and figures. Representative 

examples of LSE Table 2-l and Figure 2-l are shown. The numbers in the left column of the legends 

correspond to the numbers in the example table and figure. 

LEGEND 

See LSE Table 2-1 

(1) Route of Exposure One of the first considerations when reviewing the toxicity of a substance 

using these tables and figures should be the relevant and appropriate route of exposure. When 

sufficient data exists, three LSE tables and two LSE figures are presented in the document. The 

three LSE tables present data on the three principal routes of exposure, i.e., inhalation, oral, and 

dermal (LSE Table 2-1, 2-2, and 2-3, respectively). LSE figures are limited to the inhalation 

(LSE Figure 2-l) and oral (LSE Figure 2-2) routes. Not all substances will have data on each 

route of exposure and will not therefore have all five of the tables and figures. 
PAHs A-2 

APPENDIX A 

(2)	 Exposure Period Three exposure periods - acute (less than 15 days), intermediate (15 to 364 

days), and chronic (365 days or more) are presented within each relevant route of exposure. In 

this example, an inhalation study of intermediate exposure duration is reported. For quick 

reference to health effects occurring from a known length of exposure, locate the applicable 

exposure period within the LSE table and figure. 

(3)	 Health Effect The major categories of health effects included in LSE tables and figures are 

death, systemic, immunological, neurological, developmental, reproductive, and cancer. 

NOAELs and LOAELs can be reported in the tables and figures for all effects but cancer. 

Systemic effects are further defined in the “System” column of the LSE table (see key number 

18). 

(4)	 Key to Figure Each key number in the LSE table links study information to one or more data 

points using the same key number in the corresponding LSE figure. In this example, the study 

represented by key number 18 has been used to derive a NOAEL and a Less Serious LOAEL 

(also see the 2 ” 18r” data points in Figure 2-l). 

(5)	 Species The test species, whether animal or human, are identified in this column. Section 2.4, 

“Relevance to Public Health,” covers the relevance of animal data to human toxicity and Section 

2.3, “Toxicokinetics,” contains any available information on comparative toxicokinetics. 

Although NOAELs and LOAELs are species specific, the levels are extrapolated to equivalent 

human doses to derive an MRL. 

(6)	 Exposure Frequency/Duration The duration of the study and the weekly and daily exposure 

regimen are provided in this column. This permits comparison of NOAELs and LOAELs from 

different studies. In this case (key number 18), rats were exposed to [chemical x] via inhalation 

for 6 hours per day, 5 days per week, for 3 weeks. For a more complete review of the dosing 

regimen refer to the appropriate sections of the text or the original reference paper, i.e., Nitschke 

et al. 1981. 

(7)	 System This column further defines the systemic effects. These systems include: respiratory, 

cardiovascular, gastrointestinal, hematological, musculoskeletal, hepatic, renal, and 

dermal/ocular. “Other” refers to any systemic effect (e.g., a decrease in body weight) not covered 

in these systems. In the example of key number 18, 1 systemic effect (respiratory) was 

investigated. 

(8)	 NOAEL A No-Observed-Adverse-Effect Level (NOAEL) is the highest exposure level at which 

no harmful effects were seen in the organ system studied. Key number 18 reports a NOAEL of 3 

ppm for the respiratory system which was used to derive an intermediate exposure, inhalation 

MRL of 0.005 ppm (see footnote “b”). 

(9)	 LOAEL A Lowest-Observed-Adverse-Effect Level (LOAEL) is the lowest dose used in the 

study that caused a harmful health effect. LOAELs have been classified into “Less Serious” and 

“Serious” effects. These distinctions help readers identify the levels of exposure at which 

adverse health effects first appear and the gradation of effects with increasing dose. A brief 

description of the specific endpoint used to quantify the adverse effect accompanies the LOAEL. 

The respiratory effect reported in key number 18 (hyperplasia) is a less serious LOAEL of 10 

ppm. MRLs are not derived from Serious LOAELs. 

(10) Reference The complete reference citation is given in chapter 8 of the profile. 
PAHs A-3 

APPENDIX A 

(11) CEL A Cancer Effect Level (CEL) is the lowest exposure level associated with the onset of 

carcinogenesis in experimental or epidemiologic studies. CELs are always considered serious 

effects. The LSE tables and figures do not contain NOAELs for cancer, but the text may report 

doses not causing measurable cancer increases. 

(I 2) Footnotes Explanations of abbreviations or reference notes for data in the LSE tables are found 

in the footnotes. Footnote “b” indicates the NOAEL of 3 ppm in key number 18 was used to 

derive an MRL of 0.005 ppm. 

LEGEND 

See Figure 2-1 

LSE figures graphical!y illustrate the data presented in the corresponding LSE tables. Figures help the 

reader quickly compare health effects according to exposure concentrations for particular exposure 

periods. 

(13)	 Exposure Period The same exposure periods appear as in the LSE table. In this example, health 

effects observed within the intermediate and chronic exposure periods are illustrated. 

(14)	 Health Effect These are the categories of health effects for which reliable quantitative data 

exists. The same health effects appear in the LSE table. 

(15)	 Levels of Exposure Exposure concentrations or doses for each health effect in the LSE tables 

are graphically displayed in the LSE figures. Exposure concentration or dose is measured on the 

log scale “y” axis. Inhalation exposure is reported in mg/m3 or ppm and oral exposure is 

reported in mg/kg/day. 

(16)	 NOAEL In this example, 18r NOAEL is the critical endpoint for which an intermediate 

inhalation exposure MRL is based. As you can see from the LSE figure key, the open-circle 

symbol indicates to a NOAEL for the test species-rat. The key number 18 corresponds to the 

entry in the LSE table. The dashed descending arrow indicates the extrapolation from the 

exposure level of 3 ppm (see entry 18 in the Table) to the MRL of 0.005 ppm (see footnote “b” 

in the LSE table). 

(17)	 CEL Key number 38r is 1 of 3 studies for which Cancer Effect Levels were derived. The 

diamond symbol refers to a Cancer Effect Level for the test species-mouse. The number 38 

corresponds to the’ entry in the LSE table. 

(18)	 Estimated Upper-Bound Human Cancer Risk Levels This is the range associated with the 

upper-bound for lifetime cancer risk of 1 in 10,000 to 1 in 10,000,000. These risk levels are 

derived from the EPA’s Human Health Assessment Group’s upper-bound estimates of the slope 

of the cancer dose response curve at low dose levels (ql*). 

(19) Key to LSE Figure The Key explains the abbreviations and symbols used in the figure. 


PAHs A-6 

APPENDIX A 

Chapter 2 (Section 2.4) 

Relevance to Public Health 

The Relevance to Public Health section provides a health effects summary based on evaluations of 

existing toxicologic, epidemiologic, and toxicokinetic information. This summary is designed to 

present interpretive, weight-of-evidence discussions for human health endpoints by addressing the 

following questions. 

1. What effects are known to occur in humans? 

2. What effects observed in animals are likely to be of concern to humans? 

3. What exposure conditions are likely to be of concern to humans, especially around hazardous

 waste sites? 

The section covers endpoints in the same order they appear within the Discussion of Health Effects by 

Route of Exposure section, by route (inhalation, oral, dermal) and within route by effect. Human data 

are presented first, then animal data. Both are organized by duration (acute, intermediate, chronic). In 

vitro data and data from parenteral routes (intramuscular, intravenous, subcutaneous, etc.) are also 

considered in this section. If data are located in the scientific literature, a table of genotoxicity 

information is included. 

The carcinogenic potential of the profiled substance is qualitatively evaluated, when appropriate, using 

existing toxicokinetic, genotoxic, and carcinogenic data. ATSDR does not currently assess cancer 

potency or perform cancer risk assessments. Minimal risk levels (MRLs) for noncancer endpoints (if 

derived) and the endpoints from which they were derived are indicated and discussed. 

Limitations to existing scientific literature that prevent a satisfactory evaluation of the relevance to 

public health are identified in the Data Needs section. 

Interpretation of Minimal Risk Levels 

Where sufficient toxicologic information is available, we have derived minimal risk levels (MRLs) for 

inhalation and oral routes of entry at each duration of exposure (acute, intermediate, and chronic). 

These MRLs are not meant to support regulatory action; but to acquaint health professionals with 

exposure levels at which adverse health effects are not expected to occur in humans. They should 

help physicians and public health officials determine the safety of a community living near a chemical 

emission, given the concentration of a contaminant in air or the estimated daily dose in water. MRLs 

are based largely on toxicological studies in animals and on reports of human occupational exposure. 

MRL users should be familiar with the toxicologic information on which the number is based. 

Chapter 2.4, “Relevance to Public Health,” contains basic information known about the substance. 

Other sections such as 2.7, “Interactions with Other Chemicals”, and 2.8, “Populations that are 

Unusually Susceptible” provide important supplemental information. 
PAHs A-7 

APPENDIX A 

To derive an MRL, ATSDR generally selects the most sensitive endpoint which, in its best judgement, 

represents the most sensitive human health effect for a given exposure route and duration. ATSDR 

cannot make this judgement or derive an MRL unless information (quantitative or qualitative) is 

available for all potential systemic, neurological, and developmental effects. If this information and 

reliable quantitative data on the chosen endpoint are available, ATSDR derives an MRL using the 

most sensitive species (when information from multiple species is available) with the highest NOAEL 

that does not exceed any adverse effect levels. When a NOAEL is not available, a 

lowest-observed-adverse-effect level (LOAEL) can be used to derive an MRL, and an uncertainty 

factor (UF) of 10 must be employed. Additional uncertainty factors of 10 must be used both for 

human variability to protect sensitive subpopulations (people who are most susceptible to the health 

effects caused by the substance) and for interspecies variability (extrapolation from animals to 

humans). In deriving an MRL, these individual uncertainty factors are multiplied together. The 

product is then divided into the inhalation concentration or oral dosage selected from the study. 

Uncertainty factors used in developing a substance-specific MRL are provided in the footnotes of the 

LSE Tables.