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EPA/635/R-10/006C

www.epa.gov/iris

Toxicological Review of Benzo[a]pyrene

(CASRN 50-32-8)

In Support of Summary Information on the

Integrated Risk Information System (IRIS)

June 2012

NOTICE

This document is an Interagency Science Consultation draft. This information is distributed

solely for the purpose of pre-dissemination peer review under applicable information quality 

guidelines. It has not been formally disseminated by EPA. It does not represent and should not be 

construed to represent any Agency determination or policy. It is being circulated for review of its 

technical accuracy and science policy implications.

National Center for Environmental Assessment

Office of Research and Development

U.S. Environmental Protection Agency

Washington, DC 
Toxicological Review of Benzo[a]pyrene

This document is a draft for review purposes only and does not constitute Agency policy.

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DISCLAIMER

This document is a preliminary draft for review purposes only. This information is distributed 

solely for the purpose of pre-dissemination peer review under applicable information quality 

guidelines. It has not been formally disseminated by EPA. It does not represent and should not be 

construed to represent any Agency determination or policy. Mention of trade names or commercial 

products does not constitute endorsement of recommendation for use.
Toxicological Review of Benzo[a]pyrene

This document is a draft for review purposes only and does not constitute Agency policy.

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CONTENTS

AUTHORS | CONTRIBUTORS | REVIEWERS.................................................................................................viii

PREFACE........................................................................................................................................................xi

PREAMBLE TO IRIS TOXICOLOGICAL REVIEWS............................................................................................xiii

EXECUTIVE SUMMARY ............................................................................................................................. xxvii

LITERATURE SEARCH STRATEGY|STUDY SELECTION .............................................................................. xxxiii

1. HAZARD IDENTIFICATION....................................................................................................................1-1

1.1.Synthesis of Evidence....................................................................................................................1-1

1.1.1. Developmental Toxicity.....................................................................................................1-1

1.1.2. Reproductive Toxicity......................................................................................................1-16

1.1.3. Immunotoxicity ...............................................................................................................1-30

1.1.4. Other Toxicological Effects..............................................................................................1-37

1.1.5. Carcinogenicity ................................................................................................................1-46

1.2.Summary and Evaluation ............................................................................................................1-69

1.2.1. Weight of Evidence for Effects Other Than Cancer.........................................................1-69

1.2.2. Weight of Evidence for Carcinogenicity ..........................................................................1-70

2. DOSE-RESPONSE ANALYSIS .................................................................................................................2-1

2.1.Oral Reference Dose for Effects Other Than Cancer ....................................................................2-1

2.1.1. Identification of Studies and Effects for Dose-Response Analysis....................................2-1

2.1.2. Methods of Analysis..........................................................................................................2-5

2.1.3. Derivation of Candidate Reference Doses.........................................................................2-8

2.1.4. Derivation of Organ/System-specific Reference Doses...................................................2-10

2.1.5. Selection of the Proposed Overall Reference Dose ........................................................2-11

2.1.6. Confidence Statement.....................................................................................................2-12

2.1.7. Previous Reference Dose.................................................................................................2-12

2.2.Inhalation Reference Concentration for Effects Other Than Cancer..........................................2-13

2.2.1. Identification of Studies and Effects for Dose-Response Analysis..................................2-13

2.2.2. Methods of Analysis........................................................................................................2-14

2.2.3. Derivation of Candidate Reference Concentrations .......................................................2-16

2.2.4. Derivation of Organ/System-specific Reference Concentrations....................................2-18
Toxicological Review of Benzo[a]pyrene

This document is a draft for review purposes only and does not constitute Agency policy.

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2.2.5. Selection of the Proposed Overall Reference Concentration .........................................2-19

2.2.6. Confidence Statement.....................................................................................................2-19

2.2.7. Previous Reference Concentration..................................................................................2-19

2.2.8. Uncertainties in the Derivation of the RfD and RfC ........................................................2-20

2.3.Oral Slope Factor for Cancer.......................................................................................................2-21

2.3.1. Analysis of Carcinogenicity Data .....................................................................................2-21

2.3.2. Dose Response Analysis – Adjustments and Extrapolations Methods ...........................2-23

2.3.3. Derivation of the Oral Slope Factor.................................................................................2-24

2.3.4. Uncertainties in the Derivation of the Oral Slope Factor................................................2-26

2.3.5. Previous Oral Slope Factor ..............................................................................................2-28

2.4.Inhalation Unit Risk for Cancer...................................................................................................2-28

2.4.1. Analysis of Carcinogenicity Data .....................................................................................2-28

2.4.2. Dose Response Analysis—Adjustments and Extrapolations Methods............................2-29

2.4.3. Derivation of the Inhalation Unit Risk .............................................................................2-30

2.4.4. Uncertainties in the Derivation of the Inhalation Unit Risk ............................................2-31

2.4.5. Previous Inhalation Unit Risk...........................................................................................2-33

2.5.Dermal Slope Factor for Cancer..................................................................................................2-33

2.5.1. Analysis of Carcinogenicity Data .....................................................................................2-33

2.5.2. Dose Response Analysis – Adjustments and Extrapolations Methods ...........................2-34

2.5.3. Derivation of the Dermal Slope Factor............................................................................2-35

2.5.4. Dermal Slope Factor Cross Species Scaling .....................................................................2-37

2.5.5. Uncertainties in the Derivation of the Dermal Slope Factor...........................................2-38

2.5.6. Previous Dermal Slope Factor .........................................................................................2-40

2.6.Application of Age-dependent Adjustment Factors (ADAFs) .....................................................2-40

REFERENCES...............................................................................................................................................R-1
Toxicological Review of Benzo[a]pyrene

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TABLES 

Table ES- 1. Summary of proposed reference dose (RfD) derivation.......................................................xxviii

Table ES- 2. Summary of proposed reference concentration (RfC) derivation .......................................xxix

Table LS- 1. Details of the literature search strategy employed ..................................................................xxxiv

Table 1-1. Evidence pertaining to the developmental effects of benzo[a]pyrene in humans ............. 1-4

Table 1-2. Evidence pertaining to the developmental effects of benzo[a]pyrene in animals.............. 1-5

Table 1-3. Evidence pertaining to the neurodevelopmental effects of benzo[a]pyrene in humans ....1-

11

Table 1-4. Evidence pertaining to the neurodevelopmental effects of benzo[a]pyrene in animals1-12

Table 1-5. Evidence pertaining to the male reproductive toxicity of benzo[a]pyrene in animals...1-19

Table 1-6. Evidence pertaining to the female reproductive effects of benzo[a]pyrene in humans 1-25

Table 1-7. Evidence pertaining to the female reproductive effects of benzo[a]pyrene in animals.1-26

Table 1-8. Evidence pertaining to the immune effects of benzo[a]pyrene in animals ........................1-33

Table 1-9. Evidence pertaining to the forestomach effects of benzo[a]pyrene in animals.................1-38

Table 1-10. Evidence pertaining to other systemic effects of benzo[a]pyrene in animals .................1-44

Table 1-11. Cancer sites for PAH-related agents reviewed by IARC............................................................1-48

Table 1-12. Summary of epidemiologic studies of benzo[a]pyrene (direct measures) in relation to 

lung cancer risk: Tier 1 studies....................................................................................................................1-48

Table 1-13. Summary of epidemiologic studies of benzo[a]pyrene (direct measures) in relation to 

lung cancer risk: Tier 2 studies....................................................................................................................1-50

Table 1-14. Summary of epidemiologic studies of benzo[a]pyrene (direct measures) in relation to 

bladder cancer risk ...........................................................................................................................................1-52

Table 1-15. Tumors observed in chronic oral animal bioassays. ..................................................................1-54

Table 1-16. Tumors observed in chronic inhalation animal bioassays.......................................................1-56

Table 1-17. Tumor observations in dermal animal bioassays........................................................................1-57

Table 1-18. Experimental support for the postulated key events for the hypothesized mutagenic 

MOA.........................................................................................................................................................................1-65

Table 1-19. Supporting evidence for the carcinogenic to humans cancer descriptor for 

benzo[a]pyrene ..................................................................................................................................................1-74

Table 2-1. Summary of Derivation of Points of Departure................................................................................ 2-7

Table 2-2. Effects and corresponding derivation of candidate RfDs............................................................. 2-8

Table 2-3. Organ/system-specific RfDs and proposed RfD for benzo[a]pyrene.....................................2-11

Table 2-4. Summary of Derivation of Points of Departure..............................................................................2-16

Table 2-5. Effects and corresponding derivation of candidate RfCs ...........................................................2-17

Table 2-6. Organ/system-specific RfCs and proposed RfC for benzo[a]pyrene......................................2-19

Table 2-7. Summary of the Oral Slope Factor Derivations ..............................................................................2-25

Table 2-8. Summary of uncertainties in the benzo[a]pyrene oral cancer slope factor (OSF)..........2-27

Table 2-9. Summary of the Inhalation Unit Risk Derivation...........................................................................2-31

Table 2-10. Summary of uncertainties in the benzo[a]pyrene cancer inhalation unit risk (IUR)...2-32

Table 2-11. Summary of Dermal Slope Factor Derivations.............................................................................2-36

Table 2-12. Summary of uncertainties in the benzo[a]pyrene cancer dermal slope factor (DSF)..2-39

Table 2-13. Application of ADAFs for the estimation of benzo[a]pyrene cancer risk following 

lifetime (70-year) oral exposure .................................................................................................................2-41

Table 2-14. Application of ADAFs for the estimation of benzo[a]pyrene cancer risk following 

lifetime (70-year) inhalation exposure ....................................................................................................2-42

Table 2-15. Application of ADAFs for the estimation of benzo[a]pyrene cancer risk following 

lifetime (70-year) dermal exposure...........................................................................................................2-42
Toxicological Review of Benzo[a]pyrene

This document is a draft for review purposes only and does not constitute Agency policy.

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FIGURES

Figure LS- 1. Study selection strategy......................................................................................................................xxxv

Figure 1-1. Exposure-response array for developmental effects following oral exposure ................. 1-7

Figure 1-2. Exposure-response array for neurodevelopmental effects following oral exposure...1-14

Figure 1-3. Exposure-response array for male reproductive effects following oral exposure ........1-22

Figure 1-4. Exposure-response array for female reproductive effects following oral exposure ....1-27

Figure 1-5. Exposure-response array for immune effects following oral exposure.............................1-35

Figure 1-6. Exposure-response array for forestomach hyperplasia following oral exposure .........1-39

Figure 1-7. Proposed metabolic activation pathways and key events in the carcinogenic mode of 

action for benzo[a]pyrene..............................................................................................................................1-59

Figure 2-1. Candidate RfDs with corresponding POD and composite UF ..................................................2-10

Figure 2-2. Candidate RfCs with corresponding POD and composite UF...................................................2-18
Toxicological Review of Benzo[a]pyrene

This document is a draft for review purposes only and does not constitute Agency policy.

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ABBREVIATIONS

ADAF age-dependent adjustment factor

AhR aryl hydrocarbon receptor

AIC Akaike’s information criterion 

AKR aldo-keto reductase

ALT alanine aminotransferase 

AST aspartate aminotransferase 

ATSDR Agency for Toxic Substances and 

Disease Registry 

B[a]P benzo[a]pyrene

BMC benchmark concentration

BMD benchmark dose 

BMCL benchmark concentration lower 

confidence limit

BMDL benchmark dose lower confidence limit 

BMDS Benchmark Dose Software 

BMR benchmark response 

BPDE benzo[a]pyrene-7,8-diol-9,10-epoxide

BUN blood urea nitrogen 

BW body weight 

CA chromosomal aberration 

CASRN Chemical Abstracts Service Registry 

Number 

CERCLA Comprehensive Environmental 

Response, Compensation, and Liability 

Act

CI confidence interval

CYP450 cytochrome P450 

DAF dosimetric adjustment factor 

DHH dihydrodiol dehydrogenase

DNA deoxyribonucleic acid 

DSF dermal slope factor

EPA Environmental Protection Agency 

ETS environmental tobacco smoke

FSH follicle stimulating hormone

FTOT total fractional deposition

GD gestation day 

GGT γ-glutamyl transferase 

HEC human equivalent concentration 

HED human equivalent dose 

HERO Health and Environmental Research 

Online

HSC hematopoietic stem cells

i.p. intraperitoneal 

IRIS Integrated Risk Information System 

i.t. intratracheal

IUR inhalation unit risk

IVF in vitro fertilization 

LH luteinizing hormone

LOAEL lowest-observed-adverse-effect level 

MMAD mass median aerodynamic diameter

MN micronuclei 

MOA mode of action

MPPP Multi-Path Particle Deposition

NCEA National Center for Environmental 

Assessment 

NMDA N-methyl-D-aspartate

NOAEL no-observed-adverse-effect level 

NPL National Priorities List

NTP National Toxicology Program 

OR odds ratio

ORD Office of Research and Development 

OSF oral slope factor

PAH polycyclic aromatic hydrocarbon

PBPK physiologically based pharmacokinetic 

PCNA proliferating cell nuclear antigen 

PND post-natal day

POD point of departure 

POD[ADJ] duration-adjusted POD 

RBC red blood cells

RDDRER regional deposited dose ratio for 

extrarespiratory effects

RfC inhalation reference concentration 

RfD oral reference dose 

RNA ribonucleic acid 

ROS reactive oxygen species

RR relative risk

SCC squamous cell carcinoma

SCE sister chromatid exchange 

s.c. subcutaneous

SC SA sperm chromatic structure assay

SD standard deviation 

SE standard error 

SIR standardized incidence ratio

SMR standardized mortality ratio

SRBC sheep red blood cells

SSB single strand break

TWA time-weighted average 

UCL upper confidence limit

UF uncertainty factor 

UFA interspecies uncertainty factor 

UFH intraspecies uncertainty factor 

UFL LOAEL-to-NOAEL uncertainty factor

UFS subchronic-to-chronic uncertainty 

factor 

UFD database deficiencies uncertainty factor 

U.S. United States of America

WBC white blood cells

WHO-NCTB World Health Organization 

Neurobehavioral Core Test Battery
Toxicological Review of Benzo[a]pyrene

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AUTHORS | CONTRIBUTORS | REVIEWERS

Assessment Team

Christine Cai, MS

Glinda Cooper, Ph.D.

Louis D’Amico, Ph.D.

Jason Fritz, Ph.D.

Martin Gehlhaus, MHS

Catherine Gibbons, Ph.D.

Karen Hogan, MS

Kathleen Newhouse, MS (Chemical 

Manager)

Margaret Pratt, Ph.D.

Keith Salazar, Ph.D.

John Schaum (retired)

Suryanarayana Vulimiri, DVM

Gene Ching-Hung Hsu, Ph.D. (formerly 

with EPA)

U.S. EPA

Office of Research and Development

National Center for Environmental Assessment

Washington, DC

John Cowden, Ph.D.

John Stanek, Ph.D.

U.S. EPA

Office of Research and Development

National Center for Environmental Assessment

Research Triangle Park, NC

Chris Brinkerhoff, Ph.D. Oak Ridge Institute for Science and Education Fellow

Scott Glaberman, Ph.D. American Association for the Advancement of 

Science Fellow

1

Scientific Support Team

Lynn Flowers, Ph.D., DABT

John Fox

Martin Gehlhaus, MHS

Connie Kang-Sickel, Ph.D.

Jamie B. Strong, Ph.D.

Paul White

U.S. EPA

Office of Research and Development

National Center for Environmental Assessment

Washington, DC

Reeder Sams, Ph.D.

John Cowden, Ph.D.

U.S. EPA

Office of Research and Development

National Center for Environmental Assessment

Research Triangle Park, NC

2

Production Team

Taukecha Cunningham

Maureen Johnson

Terri Konoza

Vicki Soto

U.S. EPA

Office of Research and Development

National Center for Environmental Assessment

Washington, DC

3
Toxicological Review of Benzo[a]pyrene

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Contractor Support

Heather Carlson-Lynch, S.M.

Peter McClure, Ph.D.

Megan Riccardi

Kelly Salinas

Joe Santodonato

Julie Stickney, Ph.D., DABT

SRC, Inc., Chemical, Biological, and 

Environmental Center, Syracuse, NY

George Holdsworth, Ph.D.

Lutz W. Weber, Ph.D., DABT

Oak Ridge Institute for Science and Education, 

Oak Ridge, TN

Janusz Z. Byczkowski, Fairborn, OH, Ph.D., D.Sc., DABT

1

Executive Direction

Rebecca Clark, MPH

Vincent Cogliano, Ph.D.

Lynn Flowers, Ph.D., DABT

Samantha Jones, Ph.D.

Jamie Strong, Ph.D.

U.S. EPA

Office of Research and Development

National Center for Environmental Assessment

Washington, DC

2

Internal Review Team

Stephen Nesnow, Ph.D. U.S. EPA

National Health and Environmental Effects Research 

Laboratory

Research Triangle Park, NC

Rita Schoeny, Ph.D. U.S. EPA

Office of Water

Washington, DC

3
Toxicological Review of Benzo[a]pyrene

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Reviewers

1 This assessment was provided for review to scientists in EPA’s Program and Regional Offices. 

2 Comments were submitted by:

Office of Children’s Health Protection, Washington DC 

Office of Policy, Washington, DC 

Office of Solid Waste and Emergency Response, Washington DC

Office of Water, Washington DC

Region 2, New York, NY

Region 3, Philadelphia, PA

Region 8, Denver, CO

3 This assessment was provided for review to other federal agencies and Executive Offices of the 

4 President. Comments were submitted by:

Agency for Toxic Substances and Disease Registry, Centers for Disease Control, Department of 

Health and Human Services 

Department of Defense

National Aeronautics and Space Administration

National Institute for Occupational Safety and Health, Centers for Disease Control, Department of 

Health and Human Services

Office of Management and Budget, Executive Office of the President

White House Council on Environmental Quality, Executive Office of the President

5
Toxicological Review of Benzo[a]pyrene

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PREFACE

1 This Toxicological Review, prepared under the auspices of EPA’s Integrated Risk 

2 Information System (IRIS) program, critically reviews the publicly available studies on 

3 benzo[a]pyrene in order to identify potential adverse health effects and to characterize exposure4 response relationships. Benzo[a]pyrene is found in the environment and in food, it occurs in 

5 conjunction with other structurally related chemical compounds known as polycyclic aromatic 

hydrocarbons (PAHs). 6 1 Benzo[a]pyrene is universally present in these mixtures and is often used 

7 as an indicator chemical to measure exposure to PAH mixtures. 

8 Benzo[a]pyrene is listed as a hazardous substance under the Comprehensive 

9 Environmental Response, Compensation, and Liability Act of 1980 (CERCLA), is found at 524 

10 hazardous waste sites on the National Priorities List (NPL) and is ranked number 9 out of 275 

11 chemicals on the Priority List of Hazardous Substances for CERCLA (ATSDR, 2011). 

12 Benzo[a]pyrene is also listed as a drinking water contaminant under the Safe Drinking Water Act. 

13 This assessment updates a previous IRIS assessment of benzo[a]pyrene that was developed 

14 in 1987. The previous assessment included a cancer descriptor and oral slope factor. New 

15 information has become available, and this assessment reviews information on all health effects by 

16 all exposure routes. Organ/system-specific reference values are calculated based on 

17 developmental, reproductive and immune system toxicity data. These reference values may be 

18 useful for cumulative risk assessments that consider the combined effect of multiple agents acting 

19 on the same biological system. In addition, in consideration of the Agency’s need to estimate the 

20 potential for skin cancer from dermal exposure (US EPA 2004), especially in children exposed to 

21 contaminated soil, this assessment includes the IRIS Program’s first dermal slope factor. 

22 This assessment was conducted in accordance with EPA guidance, which is cited and 

23 summarized in the Preamble to IRIS Toxicological Reviews. The findings of this assessment and 

24 related documents produced during its development are available on the IRIS website 

25 (http://www.epa.gov/iris). Appendices for chemical and physical properties, toxicokinetic 

26 information, and summaries of toxicity studies and other information are provided as Supplemental 

27 Information to this assessment.

28 For additional information about this assessment or for general questions regarding IRIS, 

29 please contact EPA’s IRIS Hotline at 202-566-1676 (phone), 202-566-1749 (fax), or 

30 hotline.iris@epa.gov.

 

1 PAHs are a large class of chemical compounds formed during the incomplete combustion of 

organic matter. They consist of only carbon and hydrogen arranged in two or more fused rings.
Toxicological Review of Benzo[a]pyrene

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1 Chemical Properties and Uses

2 Benzo[a]pyrene is a five-ring PAH. It is a pale yellow crystalline solid with a faint aromatic 

3 odor. It is relatively insoluble in water and has low volatility. Benzo[a]pyrene is released to the air 

4 from both natural and anthropogenic sources and removed from the atmosphere by photochemical 

5 oxidation; reaction with nitrogen oxides, hydroxy and hydroperoxy radicals, ozone, sulfur oxides, 

6 and peroxyacetyl nitrate; and dry deposition to land or water. In air, benzo[a]pyrene is 

7 predominantly adsorbed to particulates but may also exist as a vapor at high temperatures (NLM, 

8 2010). 

9 There is no known commercial use for benzo[a]pyrene; it is only produced as a research 

10 chemical. Benzo[a]pyrene is ubiquitous in the environment primarily as a result of incomplete 

11 combustion emissions. It is released to the environment via both natural sources (such as forest 

12 fires) and anthropogenic sources including stoves/furnaces burning fossil fuels (especially wood 

13 and coal), motor vehicle exhaust, cigarettes, and various industrial combustion processes (ATSDR, 

14 1995). Benzo[a]pyrene is also found in soot and coal tars. Mahler et al. (2005) has reported that 

15 urban run-off from asphalt-paved car parks treated with coats of coal-tar emulsion seal could 

16 account for the majority of PAHs in many watersheds. Benzo[a]pyrene exposure can also occur to 

17 workers involved in the production of aluminum, coke, graphite, silicon carbide, and in coal tar 

18 distillation. The major sources of non-occupational exposure are cigarettes and food.

19 Assessments by Other National and International Health Agencies

20 Toxicity information on benzo[a]pyrene has been evaluated by California EPA (CalEPA), the 

21 World Health Organization, Health Canada, the International Agency for Research on Cancer, and 

22 the European Union. The results of these assessments are presented in Appendix A. It is important 

23 to recognize that these assessments may have been prepared for different purposes and may utilize 

24 different methods, and that newer studies may be included in the IRIS assessment.
Toxicological Review of Benzo[a]pyrene

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1

2 PREAMBLE TO IRIS TOXICOLOGICAL REVIEWS

3 1. Scope of the IRIS Program

4 Soon after EPA was established in 

5 1970, it was at the forefront of developing 

6 risk assessment as a science and applying it in 

7 decisions to protect human health and the 

8 environment. The Clean Air Act, for example, 

9 mandates that EPA provide “an ample margin 

10 of safety to protect public health”; the Safe 

11 Drinking Water Act, that “no adverse effects 

12 on the health of persons may reasonably be 

13 anticipated to occur, allowing an adequate 

14 margin of safety.” Accordingly, EPA uses 

15 information on the adverse health effects and 

16 on exposure levels below which these effects 

17 are not anticipated to occur.

18 IRIS assessments critically review the 

19 publicly available studies to identify adverse 

20 health effects from long-term exposure to

21 chemicals and to characterize exposure22 response relationships. An assessment may 

23 cover a single chemical, a group of 

24 structurally or toxicologically related 

25 chemicals, or a complex mixture. Exceptions 

26 are chemicals currently used exclusively as 

27 pesticides, ionizing and non-ionizing 

28 radiation, and criteria air pollutants listed 

29 under section 108 of the Clean Air Act 

30 (carbon monoxide, lead, nitrogen oxides, 

31 ozone, particulate matter, and sulfur oxides; 

32 EPA’s Integrated Science Assessments

33 evaluate the effects from these pollutants in 

34 ambient air).

35 Periodically, the IRIS Program asks

36 other EPA programs and regions, other 

37 federal agencies, state government agencies, 

38 and the general public to nominate chemicals 

39 and mixtures for future assessment or 

40 reassessment. These agents may be found in 

41 air, water, soil, or sediment. Selection is based 

42 on program and regional office priorities and 

43 on availability of adequate information to 

44 evaluate the potential for adverse effects. IRIS 

45 may assess other agents as an urgent public 

46 health need arises. IRIS also reassesses agents 

47 as significant new studies are published.

48 2. Process for Developing and Peer

49 Reviewing IRIS Assessments

50 The process for developing IRIS 

51 assessments (revised in May 2009) involves 

52 critical analysis of the pertinent studies, 

53 opportunities for public input, and multiple 

54 levels of scientific review. EPA revises draft 

55 assessments after each review, and external 

56 drafts and comments become part of the 

57 public record (U.S. EPA, 2009).

58 Step 1. Development of a draft 

59 Toxicological Review (usually about 

60 11-1/2 months duration). The draft 

61 assessment considers all pertinent 

62 publicly available studies and applies 

63 consistent criteria to evaluate the studies, 

64 identify health effects, weigh the evidence 

65 of causation for each effect, identify 

66 mechanistic events and pathways, and 

67 derive toxicity values.

68 Step 2. Internal review by scientists in EPA 

69 programs and regions (2 months). The 

70 draft assessment is revised to address 

71 comments from within EPA.

72 Step 3. Interagency science consultation 

73 with other federal agencies and the 

74 Executive Offices of the President

75 (1-1/2 months). The draft assessment is 

76 revised to address the interagency 

77 comments. The science consultation draft, 

78 interagency comments, and EPA’s 

79 response to major comments become part 

80 of the public record.

81 Step 4. External peer review, after public 

82 review and comment (3-1/2 months).

83 EPA releases the draft assessment for 
Toxicological Review of Benzo[a]pyrene

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1 public review and comment, followed by 

2 external peer review. The peer review 

3 meeting is open to the public and includes 

4 time for oral public comments. The peer 

5 reviewers also receive the written public 

6 comments. The peer reviewers assess 

7 whether the evidence has been assembled 

8 and evaluated according to guidelines and 

9 whether the conclusions are justified by 

10 the evidence. The peer review draft, peer 

11 review report, and written public 

12 comments become part of the public 

13 record.

14 Step 5. Revision of draft Toxicological 

15 Review and development of draft IRIS 

16 Summary (2 months). The draft 

17 assessment is revised to reflect the peer 

18 review comments, public comments, and 

19 newly published studies that are critical 

20 to the conclusions of the assessment. The 

21 disposition of peer review comments and 

22 public comments becomes part of the 

23 public record.

24 Step 6. Final EPA review and interagency 

25 science discussion with other federal 

26 agencies and the Executive Offices of 

27 the President (1-1/2 months). The draft 

28 assessment and summary are revised to 

29 address EPA and interagency comments.

30 The science discussion draft, written 

31 interagency comments, and EPA’s 

32 response to major comments become part 

33 of the public record.

34 Step 7. Completion and posting (1 month).

35 The Toxicological Review and IRIS 

36 Summary are posted on the IRIS website 

37 (http://www.epa.gov/iris/).

38 The remainder of this Preamble 

39 addresses step 1, the development of a draft 

40 Toxicological Review. IRIS assessments 

41 follow standard practices of evidence 

42 evaluation and peer review, many of which 

43 are discussed in EPA guidelines (U.S. EPA,

44 1986a, 1986b, 1991, 1996, 1998, 2000a, 

45 2005a, 2005b) and other methods (U.S. EPA,

46 1994, 2000b, 2002, 2006a, 2006b, 2011).

47 Transparent application of scientific 

48 judgment is of paramount importance. To 

49 provide a harmonized approach across IRIS 

50 assessments, this Preamble summarizes 

51 concepts from these guidelines and 

52 emphasizes principles of general 

53 applicability.

54 3. Identifying and Selecting 

55 Pertinent Studies

56 3.1. Identifying studies

57 Before beginning an assessment, EPA 

58 conducts a comprehensive search of the 

59 primary scientific literature. The literature 

60 search follows standard practices and 

61 includes the PubMed and ToxNet databases of 

62 the National Library of Medicine and other 

63 databases listed in EPA’s HERO system 

64 (Health and Environmental Research Online, 

65 http://hero.epa.gov/). Each assessment 

66 specifies the search strategies, keywords, and 

67 cut-off dates of its literature searches. EPA 

68 posts the results of the literature search on 

69 the IRIS website and requests information 

70 from the public on additional studies and 

71 ongoing research.

72 EPA also considers studies received 

73 through the IRIS Submission Desk and studies 

74 (typically unpublished) submitted to EPA 

75 under the Toxic Substances Control Act.

76 Material submitted as Confidential Business 

77 Information is considered only if it includes 

78 health and safety data that can be publicly 

79 released. If a study that may be critical to the 

80 conclusions of the assessment has not been 

81 peer-reviewed, EPA will have it peer82 reviewed.

83 EPA also examines the toxicokinetics 

84 of the agent to identify other chemicals (for 

85 example, major metabolites of the agent) to 

86 include in the assessment if adequate 

87 information is available, in order to more 

88 fully explain the toxicity of the agent and to 

89 suggest dose metrics for subsequent 

90 modeling.

91 In assessments of chemical mixtures, 

92 mixture studies are preferred for their ability 

93 to reflect interactions among components.

94 The literature search seeks, in decreasing 

95 order of preference (U.S. EPA, 1986a, 2000a):
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1 – Studies of the mixture being assessed.

2 – Studies of a sufficiently similar mixture.

3 In evaluating similarity, the assessment 

4 considers the alteration of mixtures in 

5 the environment through partitioning 

6 and transformation.

7 – Studies of individual chemical 

8 components of the mixture, if there are 

9 not adequate studies of sufficiently 

10 similar mixtures.

11 3.2. Selecting pertinent epidemiologic 

12 studies

13 Study design is the key consideration 

14 for selecting pertinent epidemiologic studies 

15 from the results of the literature search.

16 – Cohort studies and case-control studies 

17 provide the strongest epidemiologic 

18 evidence, as they collect information 

19 about individual exposures and effect.

20 – Ecologic studies (geographic 

21 correlation studies) relate exposures 

22 and effects by geographic area. They 

23 can provide strong evidence if there are 

24 large exposure contrasts between 

25 geographic areas, relatively little 

26 exposure variation within study areas, 

27 and population migration is limited.

28 – Case reports of high or accidental 

29 exposure lack definition of the 

30 population at risk and the expected 

31 number of cases. They can provide 

32 information about a rare effect or about 

33 the relevance of analogous results in 

34 animals.

35 The assessment briefly reviews 

36 ecologic studies and case reports but reports

37 details only if they suggest effects not 

38 identified by other epidemiologic studies.

39 3.3. Selecting pertinent experimental 

40 studies

41 Exposure route is a key design 

42 consideration for selecting pertinent 

43 experimental studies from the results of the 

44 literature search.

45 – Studies of oral, inhalation, or dermal 

46 exposure involve passage through an 

47 absorption barrier and are considered 

48 most pertinent to human 

49 environmental exposure.

50 – Injection or implantation studies are 

51 often considered less pertinent but may 

52 provide valuable toxicokinetic or 

53 mechanistic information. They also may 

54 be useful for identifying effects in 

55 animals if deposition or absorption is 

56 problematic (for example, for particles 

57 and fibers).

58 Exposure duration is also a key design 

59 consideration for selecting pertinent 

60 experimental studies.

61 – Studies of effects from chronic exposure 

62 are most pertinent to lifetime human 

63 exposure.

64 – Studies of effects from less-than65 chronic exposure are pertinent but less 

66 preferred than studies of chronic 

67 exposure.

68 Short-duration studies involving 

69 animals or humans may provide toxicokinetic 

70 or mechanistic information. Research 

71 involving human subjects is considered only 

72 if conducted according to ethical principles.

73 For developmental toxicity and 

74 reproductive toxicity, irreversible effects may 

75 result from a brief exposure during a critical 

76 period of development. Accordingly, 

77 specialized study designs are used for these 

78 effects (U.S. EPA, 1991, 1996, 1998, 2006b).

79 4. Evaluating the Quality of 

80 Individual Studies

81 4.1. Evaluating the quality of 

82 epidemiologic studies

83 The assessment evaluates design and 

84 methodologic aspects that can increase or 

85 decrease the weight given to each 

86 epidemiologic study in the overall evaluation 

87 (U.S. EPA, 1991, 1994, 1996, 1998, 2005a):
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1 – Documentation of study design, 

2 methods, population characteristics, 

3 and results.

4 – Definition and selection of the study 

5 group and comparison group.

6 – Ascertainment of exposure to the 

7 chemical or mixture under 

8 consideration.

9 – Ascertainment of disease or health 

10 effect.

11 – Duration of exposure and follow-up and 

12 adequacy for assessing the occurrence 

13 of effects.

14 – Characterization of exposure during the 

15 critical periods.

16 – Sample size and statistical power to 

17 detect anticipated effects.

18 – Participation rates and potential for 

19 selection bias as a result of the achieved 

20 participation rate.

21 – Potential confounding and other 

22 sources of bias addressed in the study 

23 design or in the analysis of results. The 

24 basis for consideration of confounding 

25 is a reasonable expectation that the 

26 confounder is related to both exposure 

27 and outcome.

28 For developmental toxicity, 

29 reproductive toxicity, neurotoxicity, and 

30 cancer there is further guidance on the 

31 nuances of evaluating epidemiologic studies 

32 of these effects (U.S. EPA, 1991, 1996, 1998, 

33 2005a).

34 4.2. Evaluating the quality of 

35 experimental studies

36 The assessment evaluates design and 

37 methodologic aspects that can increase or 

38 decrease the weight given to each 

39 experimental study in the overall evaluation 

40 (U.S. EPA, 1991, 1994, 1996, 1998, 2005a):

41 – Documentation of study design, animals 

42 and methods, basic data, and results.

43 – Relevance to humans of the animal 

44 model and the experimental methods.

45 – Characterization of the nature and 

46 extent of impurities and contaminants 

47 of the administered chemical or 

48 mixture.

49 – Characterization of dose and dosing 

50 regimen (including age at exposure) 

51 and their adequacy to elicit adverse 

52 effects.

53 – Numbers of animals and statistical 

54 power to detect dose-related 

55 differences or trends.

56 – Ascertainment of survival, vital signs, 

57 disease or effects, and cause of death.

58 – Control of other variables that could 

59 influence the occurrence of effects.

60 The assessment uses statistical tests 

61 to evaluate whether the observations may be 

62 due to chance. The standard for determining 

63 statistical significance of a response is a trend 

64 test or comparison of outcomes in the 

65 exposed group with concurrent controls. In 

66 some situations, examination of historical 

67 control data from the same laboratory within 

68 a few years of the study may improve the 

69 analysis. For an uncommon effect that is not 

70 statistically significant compared with 

71 concurrent controls, historical controls may 

72 show that the effect is unlikely to be due to 

73 chance. For a response that appears 

74 significant against a concurrent control 

75 response that is unusual, historical controls 

76 may offer a different interpretation (U.S. EPA,

77 2005a).

78 For developmental toxicity, 

79 reproductive toxicity, neurotoxicity, and 

80 cancer there is further guidance on the 

81 nuances of evaluating experimental studies of 

82 these effects (U.S. EPA, 1991, 1996, 1998, 

83 2005a). In multi-generation studies, agents 

84 that produce developmental effects at doses 

85 that are not toxic to the maternal animal are 

86 of special concern. Effects that occur at doses 

87 associated with mild maternal toxicity are not 

88 assumed to result only from maternal 

89 toxicity. Moreover, maternal effects may be 
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1 reversible, while effects on the offspring may 

2 be permanent (U.S. EPA, 1991, 1998).

3 4.3. Reporting study results

4 The assessment uses evidence tables 

5 to summarize details of the design and key 

6 results of pertinent studies. There may be 

7 separate tables for each site of toxicity or type 

8 of study.

9 If a large number of studies observe 

10 the same effect, the assessment considers the 

11 study characteristics in this section to identify 

12 the strongest studies or types of study. The 

13 tables report details from these studies, and 

14 the assessment explains the reasons for not 

15 reporting details of other studies or groups of 

16 studies that do not add new information.

17 Supplemental material provides references to 

18 all studies considered, including those not 

19 summarized in the tables.

20 The assessment discusses strengths 

21 and limitations that affect the interpretation 

22 of each study. If the interpretation of a study 

23 in the assessment differs from that of the 

24 study authors, the assessment discusses the 

25 basis for the difference.

26 As a check on the selection and 

27 evaluation of pertinent studies, EPA asks peer 

28 reviewers to identify studies that were not 

29 adequately considered.

30 5. Weighing the Overall Evidence of

31 Each Effect

32 5.1. Weighing epidemiologic evidence

33 For each effect, the assessment 

34 evaluates the evidence from the 

35 epidemiologic studies as a whole to 

36 determine the extent to which any observed 

37 associations may be causal. Positive, negative, 

38 and null results are given weight according to 

39 study quality. This evaluation considers 

40 aspects of an association that suggest 

41 causality, discussed by Hill (1965) and 

42 elaborated by Rothman and Greenland 

43 (1998) (U.S. EPA, 1994, 2002, 2005a; DHHS,

44 2004).

45 Strength of association: The finding of a 

46 large relative risk with narrow confidence 

47 intervals strongly suggests that an 

48 association is not due to chance, bias, or 

49 other factors. Modest relative risks, 

50 however, may reflect a small range of 

51 exposures, an agent of low potency, an 

52 increase in an effect that is common, 

53 exposure misclassification, or other 

54 sources of bias.

55 Consistency of association: An inference of 

56 causality is strengthened if elevated risks

57 are observed in independent studies of 

58 different populations and exposure 

59 scenarios. Reproducibility of findings 

60 constitutes one of the strongest 

61 arguments for causality. Discordant 

62 results sometimes reflect differences in 

63 study design, exposure, or confounding 

64 factors.

65 Specificity of association: As originally 

66 intended, this refers to one cause 

67 associated with one effect. Current 

68 understanding that many agents cause 

69 multiple effects and many effects have 

70 multiple causes make this a less 

71 informative aspect of causality, unless the 

72 effect is rare or unlikely to have multiple 

73 causes.

74 Temporal relationship: A causal 

75 interpretation requires that exposure 

76 precede development of the effect.

77 Biologic gradient (exposure-response 

78 relationship): Exposure-response 

79 relationships strongly suggest causality. A 

80 monotonic increase is not the only 

81 pattern consistent with causality. The 

82 presence of an exposure-response 

83 gradient also weighs against bias and 

84 confounding as the source of an 

85 association.

86 Biologic plausibility: An inference of 

87 causality is strengthened by data 

88 demonstrating plausible biologic 

89 mechanisms, if available.

90 Coherence: An inference of causality is 

91 strengthened by supportive results from 
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1 animal experiments, toxicokinetic studies, 

2 and short-term tests. Coherence may also 

3 be found in other lines of evidence, such 

4 as changing disease patterns in the 

5 population.

6 “Natural experiments”: A change in 

7 exposure that brings about a change in 

8 disease frequency provides strong 

9 evidence of causality, for example, an 

10 intervention to reduce exposure in the 

11 workplace or environment that is 

12 followed by a reduction of an adverse 

13 effect.

14 Analogy: Information on structural 

15 analogues or on chemicals that induce 

16 similar mechanistic events can provide 

17 insight into causality.

18 These considerations are consistent 

19 with guidelines for systemic reviews that 

20 evaluate the quality and strength of evidence.

21 Confidence is increased if the magnitude of 

22 effect is large, if there is evidence of an 

23 exposure-response relationship, or if an 

24 association was observed and the plausible 

25 biases would tend to decrease the reported 

26 effect. Confidence is decreased for study 

27 limitations, inconsistency of results, 

28 indirectness of evidence, imprecision, or 

29 reporting bias (Guyatt et al., 2008a,b).

30 To make clear how much the 

31 epidemiologic evidence contributes to the 

32 overall weight of the evidence, the 

33 assessment may choose a descriptor such as 

34 sufficient evidence, suggestive evidence, 

35 inadequate evidence, or evidence suggestive 

36 of no causal relationship to characterize the 

37 epidemiologic evidence of each effect (DHHS,

38 2004).

39 5.2. Weighing experimental animal 

40 evidence

41 For each effect, the assessment 

42 evaluates the evidence from the animal 

43 experiments as a whole to determine the 

44 extent to which they indicate a potential for 

45 effects in humans. Consistent results across 

46 various species and strains increase 

47 confidence that similar results would occur in 

48 humans. Several concepts discussed by Hill 

49 (1965) are pertinent to the weight of 

50 experimental results: consistency of 

51 response, dose-response relationships, 

52 strength of response, biologic plausibility, and 

53 coherence (U.S. EPA, 1994, 2002, 2005a).

54 In weighing evidence from multiple 

55 experiments, U.S. EPA (2005a) distinguishes 

56 Conflicting evidence (that is, mixed positive 

57 and negative results in the same sex and 

58 strain using a similar study protocol) 

59 from 

60 Differing results (that is, positive results and 

61 negative results are in different sexes or 

62 strains or use different study protocols).

63 Negative or inconclusive results do 

64 not invalidate positive results in a different 

65 experimental system. EPA regards both as 

66 valid observations and looks to 

67 methodological differences or, if available, 

68 mechanistic information to reconcile differing 

69 results. 

70 It is well established that there are

71 critical periods for some developmental and 

72 reproductive effects. Accordingly, the 

73 assessment determines whether critical 

74 periods have been adequately investigated 

75 (U.S. EPA, 1991, 1996, 1998, 2005a, 2005b, 

76 2006b). Similarly, the assessment determines 

77 whether the database is adequate to evaluate 

78 other critical sites and effects.

79 In evaluating evidence of genetic 

80 toxicity:

81 – Demonstration of gene mutations, 

82 chromosome aberrations, or aneuploidy 

83 in humans or experimental mammals 

84 (in vivo) provides the strongest 

85 evidence.

86 – This is followed by positive results in 

87 lower organisms or in cultured cells 

88 (in vitro) or for other genetic events.

89 – Negative results carry less weight, 

90 partly because they cannot exclude the 

91 possibility of effects in other tissues 

92 (IARC 2006).
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1 For germ-cell mutagenicity, EPA has 

2 defined categories of evidence, ranging from 

3 positive results of human germ-cell 

4 mutagenicity to negative results for all effects 

5 of concern (EPA 1986b).

6 5.3. Characterizing modes of action

7 For each effect, the assessment

8 discusses the available information on its 

9 modes of action and associated key events 

10 (key events being empirically observable, 

11 necessary precursor steps or biologic 

12 markers of such steps; mode of action being a 

13 series of key events involving interaction with 

14 cells, operational and anatomic changes, and 

15 resulting in disease). Pertinent information 

16 may also come from studies of metabolites or 

17 of compounds that are structurally similar or 

18 that act through similar mechanisms. The 

19 assessment addresses several questions 

20 about each hypothesized mode of action (U.S. 

21 EPA, 2005a). Information on mode of action 

22 is not required for a conclusion that an effect 

23 is causally related to an agent (EPA 2005a).

24 1) Is the hypothesized mode of action 

25 sufficiently supported in test animals?

26 Strong support for a key event being 

27 necessary to a mode of action can come 

28 from experimental challenge to the 

29 hypothesized mode of action, in which 

30 studies that suppress a key event observe

31 suppression of the effect. Support for a 

32 mode of action is meaningfully 

33 strengthened by consistent results in 

34 different experimental models, much 

35 more so than by replicate experiments in 

36 the same model. The assessment may 

37 consider various aspects of causality in 

38 addressing this question.

39 2) Is the hypothesized mode of action 

40 relevant to humans? The assessment 

41 reviews the key events to identify critical 

42 similarities and differences between the 

43 test animals and humans. Site 

44 concordance is not assumed between 

45 animals and humans, though it may hold 

46 for certain effects or modes of action.

47 Information suggesting quantitative 

48 differences in doses where effects would 

49 occur in animals or humans is considered 

50 in the dose-response analysis but is not 

51 used to determine relevance. Similarly, 

52 anticipated levels of human exposure are 

53 not used to determine relevance.

54 3) Which populations or life-stages can 

55 be particularly susceptible to the 

56 hypothesized mode of action? The 

57 assessment reviews the key events to 

58 identify populations and life-stages that 

59 might be susceptible to their occurrence.

60 Quantitative differences may result in 

61 separate risk estimates for susceptible 

62 populations or life-stages.

63 The assessment discusses the 

64 likelihood that an agent operates through 

65 multiple modes of action. An uneven level of 

66 support for different modes of action can 

67 reflect disproportionate resources spent 

68 investigating them (U.S. EPA, 2005a). It 

69 should be noted that in clinical reviews, the 

70 credibility of a series of studies is reduced if 

71 evidence is limited to studies funded by one 

72 interested sector (Guyatt et al., 2008b).

73 For cancer, the assessment evaluates 

74 evidence of a mutagenic mode of action to 

75 guide extrapolation to lower doses and 

76 consideration of susceptible lifestages. Key 

77 data include the ability of the agent or a 

78 metabolite to react with or bind to DNA, 

79 positive results in multiple test systems, or 

80 similar properties and structure-activity 

81 relationships to mutagenic carcinogens (EPA 

82 2005a).

83 5.4. Characterizing the overall weight 

84 of the evidence

85 After weighing the epidemiologic and 

86 experimental studies pertinent to each effect, 

87 the assessment may select a standard 

88 descriptor to characterize the overall weight 

89 of the evidence. For example, the following 

90 standard descriptors combine epidemiologic, 

91 experimental, and mechanistic evidence of 

92 carcinogenicity (U.S. EPA, 2005a).

93 Carcinogenic to humans: There is 

94 convincing epidemiologic evidence of a 
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1 causal association (that is, there is 

2 reasonable confidence that the 

3 association cannot be fully explained by 

4 chance, bias, or confounding), or there is 

5 strong human evidence of cancer or its 

6 precursors, extensive animal evidence, 

7 identification of key precursor events in 

8 animals, and strong evidence that they 

9 are anticipated to occur in humans.

10 Likely to be carcinogenic to humans: The 

11 evidence demonstrates a potential hazard 

12 to humans but does not meet the criteria 

13 for carcinogenic. There may be a plausible 

14 association in humans, multiple positive 

15 results in animals, or a combination of 

16 human, animal, or other experimental 

17 evidence.

18 Suggestive evidence of carcinogenic 

19 potential: The evidence raises concern 

20 for effects in humans but is not sufficient 

21 for a stronger conclusion. This descriptor 

22 covers a range of evidence, from a 

23 positive result in the only available study 

24 to a single positive result in an extensive 

25 database that includes negative results in 

26 other species.

27 Inadequate information to assess 

28 carcinogenic potential: No other 

29 descriptors apply. Conflicting evidence can 

30 be classified as inadequate information if 

31 all positive results are opposed by 

32 negative studies of equal quality in the 

33 same sex and strain. Differing results, 

34 however, can be classified as suggestive 

35 evidence or as likely to be carcinogenic.

36 Not likely to be carcinogenic to humans:

37 There is robust evidence for concluding 

38 that there is no basis for concern. There 

39 may be no effects in both sexes of at least 

40 two appropriate animal species; positive 

41 animal results and strong, consistent 

42 evidence that each mode of action in 

43 animals does not operate in humans; or

44 convincing evidence that effects are not 

45 likely by a particular exposure route or 

46 below a defined dose.

47 Multiple descriptors may be used if 

48 there is evidence that carcinogenic effects 

49 differ by dose range or exposure route (EPA 

50 2005a).

51 EPA is investigating and may on a trial 

52 basis propose standard descriptors to 

53 characterize the overall weight of the 

54 evidence for effects other than cancer.

55 6. Selecting Studies for Derivation of 

56 Toxicity Values

57 For each effect where there is credible 

58 evidence of an association with the agent, the 

59 assessment derives toxicity values if there are 

60 suitable epidemiologic or experimental data. 

61 The decision to derive toxicity values may be 

62 linked to the weigh-of-evidence descriptor. 

63 For example, EPA typically derives toxicity 

64 values for agents classified as carcinogenic to 

65 humans or as likely to be carcinogenic (U.S. 

66 EPA, 2005a).

67 Dose-response analysis requires 

68 different information than is needed to 

69 identify the occurrence of effects. Although 

70 dose-response relationships may contribute 

71 to a qualitative assessment quantitative 

72 measures of dose and response are essential 

73 to dose-response analysis. Then, other factors 

74 being equal (U.S. EPA, 1994, 2005a):

75 – Epidemiologic studies are preferred 

76 over animal studies, if quantitative 

77 measures of exposure are available and 

78 estimated effects can be attributed to 

79 the agent.

80 – Among experimental animal models, 

81 those that respond most like humans 

82 are preferred, if the comparability of 

83 response can be determined.

84 – Studies by a route of human 

85 environmental exposure are preferred, 

86 although a validated toxicokinetic 

87 model can also be used to extrapolate 

88 across exposure routes.

89 – Studies of longer exposure duration and 

90 follow-up are preferred, to minimize 

91 uncertainty about whether effects 

92 change with time.
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1 – Studies with multiple exposure levels 

2 are preferred for their ability to provide 

3 information about the shape of the 

4 exposure-response curve.

5 – Studies with adequate power to detect 

6 effects at lower exposure levels are 

7 preferred, to minimize the extent of 

8 extrapolation to levels found in the 

9 environment.

10 Studies with non-monotonic 

11 exposure-response relationships are not 

12 necessarily excluded from the analysis. A 

13 diminished effect at higher exposure levels 

14 may be satisfactorily explained by factors 

15 such as competing toxicity, saturation of 

16 absorption or metabolism, exposure 

17 misclassification, or selection bias.

18 If a large number of studies are 

19 suitable for dose-response analysis, the 

20 assessment considers the study 

21 characteristics in this section to focus on the 

22 most informative data. The assessment 

23 explains the reasons for not analyzing other 

24 groups of studies. As a check on the selection 

25 of studies for dose-response analysis, EPA 

26 asks peer reviewers to identify studies that 

27 were not adequately considered.

28 7. Deriving Toxicity Values

29 7.1. General framework for dose30 response analysis

31 EPA uses a two-step approach that 

32 distinguishes analysis of the observed dose33 response data from inferences about lower 

34 doses (U.S. EPA, 2005a).

35 Within the observed range, the 

36 preferred approach is to use modeling to 

37 incorporate a wide range of data into the 

38 analysis. The modeling yields a point of 

39 departure (an exposure level near the lower 

40 end of the observed range, without significant 

41 extrapolation to lower doses) (Sections 7.2

42 and 7.3).

43 Extrapolation to lower doses 

44 considers what is known about the modes of 

45 action for each effect (Sections 7.4 and 7.5).

46 When response estimates at lower doses are 

47 not required, an alternative is to derive

48 reference values, which are calculated by 

49 applying factors that account for sources of 

50 uncertainty and variability to the point of 

51 departure (Section 7.6).

52 For a group of agents that induce an 

53 effect through a common mode of action, the 

54 dose-response analysis may derive a relative 

55 potency factor for each agent. A full dose56 response analysis is conducted for one well57 studied index chemical in the group, and then 

58 the potencies of other members are 

59 expressed in relative terms based on relative 

60 toxic effects, relative absorption or metabolic 

61 rates, quantitative structure-activity 

62 relationships, or receptor binding 

63 characteristics (EPA 2000a, 2005a).

64 Increasingly, EPA is basing toxicity 

65 values on combined analyses of multiple data 

66 sets or multiple responses. EPA also 

67 considers multiple dose-response approaches 

68 when they can be supported by robust data.

69 7.2. Modeling dose

70 The preferred approach for analysis 

71 of dose is toxicokinetic modeling because of 

72 its ability to incorporate a wide range of data.

73 The preferred dose metric would refer to the 

74 active agent at the site of its biologic effect or 

75 to a close, reliable surrogate measure. The 

76 active agent may be the administered 

77 chemical or a metabolite. Confidence in the 

78 use of a toxicokinetic model depends on the 

79 robustness of its validation process and on 

80 the results of sensitivity analyses (U.S. EPA,

81 1994, 2005a, 2006a).

82 Because toxicokinetic modeling can 

83 require many parameters and more data than 

84 are typically available, EPA has developed 

85 standard approaches that can be applied to 

86 typical data sets. These standard approaches 

87 also facilitate comparison across exposure 

88 patterns and species.

89 – Intermittent study exposures are 

90 standardized to a daily average over the 

91 duration of exposure. For chronic 

92 effects, daily exposures are averaged 

93 over the lifespan. Exposures during a 

94 critical period, however, are not 
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1 averaged over a longer duration (U.S. 

2 EPA, 1991, 1996, 1998, 2005a).

3 – Doses are standardized to equivalent 

4 human terms to facilitate comparison of 

5 results from different species:

6 Oral doses can be scaled 

allometrically using mg/kg3/4 7 -d as the 

8 equivalent dose metric across species.

9 Allometric scaling pertains to 

10 equivalence across species, not across 

11 lifestages, and is not used to scale 

12 doses from adult humans or mature 

13 animals to infants or children (U.S. 

14 EPA, 2005a, 2011).

15 Inhalation exposures are scaled using 

16 dosimetry models that apply species17 specific physiologic and anatomic 

18 factors and consider whether the 

19 effect occurs at the site of first contact 

20 or after systemic circulation (U.S. EPA,

21 1994).

22 It can be informative to convert doses 

23 to target sites across different exposure 

24 routes. If this approach is followed, the 

25 assessment describes the underlying data, 

26 algorithms, and assumptions (U.S. EPA,

27 2005a).

28 In the absence of study-specific data 

29 on, for example, intake rates or body weight, 

30 EPA has developed recommended values for 

31 use in dose-response analysis (EPA 1988).

32 7.3. Modeling response in the range of 

33 observation

34 Toxicodynamic (“biologically based”) 

35 modeling can incorporate data on biologic 

36 processes leading to an effect. Such models 

37 require sufficient data to ascertain a mode of 

38 action and to quantitatively support model 

39 parameters associated with its key events.

40 Because different models may provide 

41 equivalent fits to the observed data but 

42 diverge substantially at lower doses, critical 

43 biologic parameters should be measured 

44 from laboratory studies, not by model fitting.

45 Confidence in the use of a toxicodynamic 

46 model depends on the robustness of its 

47 validation process and on the results of 

48 sensitivity analyses. Peer review of the 

49 scientific basis and performance of a model is 

50 essential (U.S. EPA, 2005a).

51 Because toxicodynamic modeling can 

52 require many parameters and more 

53 knowledge and data than are typically 

54 available, EPA has developed a standard set 

55 of empirical (“curve-fitting”) models (http://

56 www.epa.gov/ncea/bmds/) that can be 

57 applied to typical data sets, including those 

58 that are nonlinear. EPA has also developed 

59 guidance on modeling dose-response data, 

60 assessing model fit, selecting suitable models, 

61 and reporting modeling results (U.S. EPA,

62 2000b). Additional judgment or alternative 

63 analyses are used when the procedure fails to 

64 yield reliable results, for example, if the fit is 

65 poor, modeling may be restricted to the lower 

66 doses, especially if there is competing toxicity 

67 at higher doses (U.S. EPA, 2005a).

68 Modeling is used to derive a point of 

69 departure (U.S. EPA, 2000b, 2005a). (See 

70 Section 7.6 for alternatives if a point of 

71 departure cannot be derived by modeling.)

72 – For dichotomous responses, the point 

73 of departure is often the 95% lower 

74 bound on the dose associated with a 

75 10% response, but a lower response 

76 that falls within the observed range 

77 may be used instead. For example, 

78 reproductive or developmental studies 

79 often have power to detect a 5% 

80 response; epidemiologic studies, 1% or 

81 lower.

82 – For continuous responses, the point of 

83 departure is ideally the dose where the 

84 effect becomes biologically significant.

85 In the absence of such definition, both 

86 statistical and biologic factors are 

87 considered.

88 7.4. Extrapolating to lower doses

89 The purpose of extrapolating to lower 

90 doses is to estimate risks from exposures 

91 below the observed data. Low-dose 

92 extrapolation is typically used for known and 

93 likely carcinogens. Low-dose extrapolation 
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1 considers what is known about modes of 

2 action (U.S. EPA, 2005a).

3 1) If a biologically based model has been 

4 developed and validated for the agent, 

5 extrapolation may use the fitted model 

6 below the observed range if significant 

7 model uncertainty can be ruled out with 

8 reasonable confidence.

9 2) Linear extrapolation is used if the dose10 response curve is expected to have a 

11 linear component below the point of 

12 departure. This includes:

13 – Agents or their metabolites that are 

14 DNA-reactive and have direct 

15 mutagenic activity.

16 – Agents or their metabolites for which 

17 human exposures or body burdens 

18 are near doses associated with key 

19 events leading to an effect.

20 Linear extrapolation is also used if the 

21 evidence is insufficient to establish a 

22 mode of action.

23 The result of linear extrapolation is 

24 described by an oral slope factor or an 

25 inhalation unit risk, which is the slope of 

26 the dose-response curve at lower doses

27 or concentrations, respectively.

28 3) Nonlinear extrapolation is used if there 

29 are sufficient data to ascertain the mode 

30 of action and to conclude that it is not 

31 linear at lower doses, and the agent does 

32 not demonstrate mutagenic or other 

33 activity consistent with linearity at lower 

34 doses. If nonlinear extrapolation is 

35 appropriate but no model is developed, 

36 an alternative is to calculate reference 

37 values.

38 If linear extrapolation is used, the 

39 assessment develops a candidate slope factor 

40 or unit risk for each suitable data set. These 

41 results are arrayed, using common dose 

42 metrics, to show the distribution of relative 

43 potency across various effects and 

44 experimental systems. The assessment then 

45 calculates an overall slope factor and an 

46 overall unit risk for the agent, considering the 

47 various dose-response analyses, the study 

48 preferences discussed in Section 6, and the 

49 possibility of basing a more robust result on 

50 multiple data sets.

51 7.5. Considering susceptible 

52 populations and life-stages

53 The assessment analyzes the available 

54 information on populations and life-stages 

55 that may be particularly susceptible to each 

56 effect. If adequate data are available, the 

57 assessment derives separate risk estimates 

58 for susceptible populations or life-stages. A 

59 tiered approach is used (U.S. EPA, 2005a).

60 1) If an epidemiologic or experimental study 

61 reports quantitative results for a 

62 susceptible population or life-stage, these 

63 data are analyzed to derive separate risk

64 estimates for susceptible individuals.

65 2) If data on risk-related parameters allow 

66 comparison of the general population and 

67 susceptible individuals, these data are 

68 used to adjust the general-population risk 

69 estimate for application to susceptible 

70 individuals.

71 3) In the absence of chemical-specific data, 

72 EPA has developed age-dependent 

73 adjustment factors early-life exposure to 

74 suspected carcinogens that have a 

75 mutagenic mode of action. There is 

76 evidence of early-life susceptibility to 

77 various carcinogenic agents, but most 

78 epidemiologic studies and cancer 

79 bioassays do not include early-life 

80 exposure. To address the potential for 

81 early-life susceptibility, EPA recommends

82 (U.S. EPA, 2005b):

83 – 10-fold adjustment for exposures 

84 before age 2 years.

85 – 3-fold adjustment for exposures 

86 between ages 2 and 16 years.

87 7.6. Reference values and uncertainty 

88 factors

89 An oral reference dose or an inhalation 

90 reference concentration is an estimate of an 

91 exposure (including in susceptible 
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xxiv DRAFT—DO NOT CITE OR QUOTE

1 subgroups) that is likely to be without an 

2 appreciable risk of adverse health effects over 

3 a lifetime (U.S. EPA, 2002). Reference values 

4 are typically calculated for effects other than 

5 cancer and for suspected carcinogens if a well 

6 characterized mode of action indicates that a 

7 necessary key event does not occur below a 

8 specific dose. Reference values provide no 

9 information about risks at higher exposure 

10 levels.

11 The assessment characterizes effects 

12 that form the basis for reference values as 

13 adverse, considered to be adverse, or a 

14 precursor to an adverse effect. For 

15 developmental toxicity, reproductive toxicity, 

16 and neurotoxicity there is guidance on 

17 adverse effects and their biologic markers 

18 (U.S. EPA, 1991, 1996, 1998).

19 To account for uncertainty and 

20 variability in the derivation of a lifetime 

21 human exposure where effects are not 

22 anticipated to occur, reference doses and 

23 reference concentrations are calculated by 

24 applying a series of uncertainty factors to the 

25 point of departure. If a modeled point of 

26 departure is not available, a no-observed27 adverse-effect level or a lowest-observed28 adverse-effect level is used instead. The 

29 assessment discusses scientific 

30 considerations involving several areas of 

31 variability or uncertainty.

32 Human variation. A factor of 10 is applied to 

33 account for variation in susceptibility 

34 across the human population and the 

35 possibility that the available data may not 

36 be representative of individuals who are 

37 most susceptible to the effect. This factor 

38 is reduced only if the point of departure is 

39 derived specifically for susceptible 

40 individuals (not for a general population 

41 that includes both susceptible and non42 susceptible individuals) (U.S. EPA, 1991, 

43 1994, 1996, 1998, 2002).

44 Animal-to-human extrapolation. A factor of 

45 10 is applied if animal results are used to 

46 make inferences about humans. This 

47 factor is often regarded as comprising 

48 toxicokinetics and toxicodynamics in 

49 equal parts. Accordingly, if the point of 

50 departure is based on toxicokinetic 

51 modeling, dosimetry modeling, or 

52 allometric scaling across species, a factor 

of 101/2 53 (rounded to 3) is applied to 

54 account for the remaining uncertainty 

55 involving toxicodynamic differences. An 

56 animal-to-human factor is not applied if a 

57 biologically based model adjusts fully for 

58 toxicokinetic and toxicodynamic 

59 differences across species (U.S. EPA,

60 1991, 1994, 1996, 1998, 2002, 2011).

61 Adverse-effect level to no-observed62 adverse-effect level. If a point of 

63 departure is based on a lowest-observed64 adverse-effect level, the assessment must 

65 infer a dose where such effects are not 

66 expected. This can be a matter of great 

67 uncertainty, especially if there is no 

68 evidence available at lower doses. A 

69 factor of 10 is applied to account for the 

70 uncertainty in making this inference. A 

71 factor other than 10 may be used, 

72 depending on the magnitude and nature 

73 of the response and the shape of the dose74 response curve (U.S. EPA, 1991, 1994, 

75 1996, 1998, 2002).

76 Subchronic-to-chronic exposure. If a point 

77 of departure is based on subchronic 

78 studies, the assessment considers 

79 whether lifetime exposure could have 

80 effects at lower levels of exposure. A 

81 factor of 10 is applied to account for the 

82 uncertainty in using subchronic studies to 

83 make inferences about lifetime exposure.

84 This factor may also be applied for 

85 developmental or reproductive effects if 

86 exposure covered less than the full critical 

87 period. A factor other than 10 may be 

88 used, depending on the duration of the 

89 studies and the nature of the response 

90 (U.S. EPA, 1994, 1998, 2002).

91 Incomplete database. If an incomplete 

92 database raises concern that further 

93 studies might identify a more sensitive 

94 effect, organ system, or life-stage, the 

95 assessment may apply a database 

96 uncertainty factor (U.S. EPA, 1991, 1994, 

97 1996, 1998, 2002). The size of the factor 
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xxv DRAFT—DO NOT CITE OR QUOTE

1 depends on the nature of the database 

2 deficiency. For example, EPA typically 

3 follows the suggestion that a factor of 10 

4 be applied if both a prenatal toxicity study 

5 and a two-generation reproduction study 

are missing, and a factor of 101/2 6 if either 

7 is missing (U.S. EPA, 2002).

8 In this way, the assessment derives 

9 candidate reference values for each suitable 

10 data set and effect that is credibly associated 

11 with the agent. These results are arrayed, 

12 using common dose metrics, to show where 

13 effects occur across a range of exposures (U.S. 

14 EPA, 1994). The assessment then selects an 

15 overall reference dose and an overall 

16 reference concentration for the agent to 

17 represent lifetime human exposure levels 

18 where effects are not anticipated to occur.

19 The assessment may also report 

20 reference values for each effect. This would 

21 facilitate subsequent cumulative risk 

22 assessments that consider the combined 

23 effect of multiple agents acting at a common 

24 site or through common mechanisms (U.S. 

25 EPA, 2002).

26 7.7. Confidence and uncertainty in the 

27 reference values

28 The assessment selects a standard 

29 descriptor to characterize the level of 

30 confidence in each reference value, based on 

31 the likelihood that the value would change 

32 with further testing. Confidence in reference 

33 values is based on quality of the studies used 

34 and completeness of the database, with more 

35 weight given to the latter. The level of 

36 confidence is increased for reference values 

37 based on human data supported by animal 

38 data (U.S. EPA, 1994).

39 High confidence: The reference value is not 

40 likely to change with further testing, 

41 except for mechanistic studies that might 

42 affect the interpretation of prior test 

43 results.

44 Medium confidence: This is a matter of 

45 judgment, between high and low 

46 confidence.

47 Low confidence: The reference value is 

48 especially vulnerable to change with 

49 further testing.

50 These criteria are consistent with

51 guidelines for systematic reviews that 

52 evaluate the quality of evidence. These also 

53 focus on whether further research would be 

54 likely to change confidence in the estimate of 

55 effect (Guyatt et al., 2008a).

56 All assessments discuss the significant 

57 uncertainties encountered in the analysis.

58 EPA provides guidance on characterization of 

59 uncertainty (U.S. EPA, 2005a). For example, 

60 the discussion distinguishes model 

61 uncertainty (lack of knowledge about the 

62 most appropriate experimental or analytic 

63 model) and parameter uncertainty (lack of 

64 knowledge about the parameters of a model).

65 Assessments also discuss human variation 

66 (interpersonal differences in biologic 

67 susceptibility or in exposures that modify the 

68 effects of the agent).

69

70 DHHS (2004) The health consequences of 

71 smoking: a report of the Surgeon 

72 General. Dept of Health and Human 

73 Services, http://www.cdc.gov/

74 tobacco/data_statistics/sgr/2004/

75 index.htm.

76 Guyatt GH, Oxman AD, Vist GE, Kunz R, Falck77 Ytter Y, Alonso-Coello P, 

78 Schünemann HJ (2008a) GRADE: an 

79 emerging consensus on rating quality 

80 of evidence and strength of 

81 recommendations. British Medical 

82 Journal 336: 924-926, http://

83 www.bmj.com/content/336/7650/

84 924.full.

85 Guyatt GH, Oxman AD, Kunz R, Vist GE, Falck86 Ytter Y, Schünemann HJ (2008b) 

87 GRADE: what is “quality of evidence” 

88 and why is it important to clinicians? 

89 British Medical Journal 336: 995-998, 

90 http:// www.bmj.com/content/336/

91 7651/995.full.

92 Hill AB (1965) The environment and disease: 

93 association or causation? Proceedings 
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xxvi DRAFT—DO NOT CITE OR QUOTE

1 of the Royal Society of Medicine 58: 

2 295-300.

3 IARC (2006) Preamble to the IARC 

4 Monographs. 

5 http://monographs.iarc.fr/.

6 Rothman KJ, Greenland S (1998) Modern 

7 Epidemiology. Philadelphia: 

8 Lippincott Williams and Wilkins.

9 U.S. EPA (1986a) Guidelines for the Health 

10 Risk Assessment of Chemical 

11 Mixtures. EPA/630/R-98/002.

12 U.S. EPA (1986b) Guidelines for Mutagenicity 

13 Risk Assessment. EPA/630/R14 98/003.

15 U.S. EPA (1988) Recommendations for and 

16 Documentation of Biological Values 

17 for Use in Risk Assessment. 

18 EPA/600/6-87/008.

19 U.S. EPA (1991) Guidelines for Developmental 

20 Toxicity Risk Assessment. 

21 EPA/600/FR-91/001.

22 U.S. EPA (1994) Methods for Derivation of 

23 Inhalation Reference Concentrations 

24 and Application of Inhalation 

25 Dosimetry. EPA/600/8-90/066F.

26 U.S. EPA (1996) Guidelines for Reproductive 

27 Toxicity Risk Assessment. 

28 EPA/630/R-96/009.

29 U.S. EPA (1998) Guidelines for Neurotoxicity 

30 Risk Assessment. EPA/630/R31 95/001F.

32 U.S. EPA (2000a) Supplementary Guidance 

33 for Conducting Health Risk 

34 Assessment of Chemical Mixtures. 

35 EPA/630/R-00/002.

36 U.S. EPA (2000b) Benchmark Dose Technical 

37 Guidance Document. EPA/630/R38 00/001 (external review draft).

39 U.S. EPA (2002) A Review of the Reference 

40 Dose and Reference Concentration 

41 Processes. EPA/630/P-02/002F.

42 U.S. EPA (2005a) Guidelines for Carcinogen 

43 Risk Assessment. EPA/630/P44 03/001F.

45 U.S. EPA (2005b) Supplemental Guidance for 

46 Assessing Susceptibility from Early47 Life Exposure to Carcinogens. 

48 EPA/630/R-03/003F.

49 U.S. EPA (2006a) Approaches for the 

50 Application of Physiologically Based 

51 Pharmacokinetic (PBPK) Models and 

52 Supporting Data in Risk Assessment. 

53 EPA/600/R-05/043F.

54 U.S. EPA (2006b) A Framework for Assessing 

55 Health Risks of Environmental 

56 Exposures to Children. EPA/600/R57 05/093F.

58 U.S. EPA (2009) IRIS Process. 

59 http://www.epa.gov/iris/process.ht

60 m.

61 U.S. EPA (2011) Recommended Use of Body 

Weight 62 3/4 as the Default Method in 

63 Derivation of the Oral Reference Dose. 

64 EPA/100/R11/0001.

65

66 May 2012 version

67
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EXECUTIVE SUMMARY

1 Occurrence and Health Effects

2

3 Benzo[a]pyrene is a five ring polycyclic aromatic hydrocarbon (PAH). 

4 Benzo[a]pyrene (along with other PAHs) is released into the atmosphere as a 

5 component of smoke from forest fires, industrial processes, vehicle exhaust, 

6 cigarettes, and through the burning of fuel (such as wood, coal, and petroleum 

7 products). Oral exposure to benzo[a]pyrene can occur by eating foods grown in 

8 areas contaminated with benzo[a]pyrene (from the air and soil) or by eating certain 

9 food products, such as charred meats, where benzo[a]pyrene is formed during the 

10 cooking process. Dermal exposure may occur from contact with soils or materials 

11 that contain soot, tar, or crude petroleum products or by using certain 

12 pharmaceutical products containing coal tars, such as those used to treat the skin 

13 conditions eczema and psoriasis. The magnitude of human exposure to 

14 benzo[a]pyrene and other PAHs depends on factors such as lifestyle (e.g., diet, 

15 tobacco smoking), occupation, and living conditions (e.g., urban versus rural setting, 

16 domestic heating, and cooking methods). 

17 Animal studies demonstrate that exposure to benzo[a]pyrene may be 

18 associated with developmental, reproductive, and immunological effects. In 

19 addition, epidemiology studies involving exposure to PAH mixtures have reported

20 associations between adverse birth outcomes (including reduced birth weight, 

21 postnatal body weight, and head circumference) and decreased fertility with 

22 internal biomarkers of exposure to benzo[a]pyrene (benzo[a]pyrene diol epoxide23 DNA adducts).

24 Studies in multiple animal species demonstrate that benzo[a]pyrene is 

25 carcinogenic at multiple tumor sites (alimentary tract, liver, kidney, respiratory 

26 tract, pharynx, and skin) by all routes of exposure. In addition, there is strong 

27 evidence of carcinogenicity in occupations involving exposure to PAH mixtures 

28 containing benzo[a]pyrene, such as aluminum production, chimney sweeping, coal 

29 gasification, coal-tar distillation, coke production, iron and steel founding, and 

30 paving and roofing with coal tar pitch. An increasing number of occupational 

31 studies demonstrate a positive exposure-response relationship with cumulative 

32 benzo[a]pyrene exposure and lung cancer. 

33

34 Effects Other Than Cancer Following Oral Exposure

35 In animals, oral exposure to benzo[a]pyrene has been shown to result in developmental 

36 toxicity, reproductive toxicity, and immunotoxicity. Developmental effects in rats and mice include

37 altered spatial learning and memory and cardiovascular effects following gestational exposures. 

38 Reproductive and immune effects include decreased sperm counts, ovarian weight and follicle 

39 numbers, and decreased immunoglobulin and B-cell numbers and thymus weight following oral 

40 exposures in adult animals. In humans, benzo[a]pyrene exposure occurs in conjunction with other
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1 PAHs and, as such, attributing the observed effects to benzo[a]pyrene is complicated. However, 

2 human studies report associations between particular health endpoints and internal measures of 

3 exposure, such as benzo[a]pyrene-DNA adducts, or external measures of benzo[a]pyrene exposure. 

4 Overall, the human studies report developmental and reproductive effects that are generally 

5 analogous to those observed in animals, and provide qualitative, supportive evidence for the

6 hazards associated with benzo[a]pyrene exposure. 

7 Developmental toxicity, represented by neurodevelopmental impairments, was chosen as 

8 the basis for the porposed RfD as the available data indicate that neurodevelopmental effects 

9 represent a sensitive hazard of benzo[a]pyrene exposure. The neurodevelopmental study by Chen 

10 et al. (2012) and the observed impaired spatial learning were used to derive the RfD. The endpoint 

11 of impaired spatial learning, as measured by the increase in latency time to find a hidden platform 

12 in the Morris water maze, was selected as the critical effect due to the sensitivity of this endpoint 

13 and the observed dose-response relationship of effects across dose groups. Benchmark dose 

14 (BMD) modeling was utilized to derive the BMDL1SD of 0.06 mg/kg-day that was used as the point of 

15 departure (POD) for RfD derivation.

16 The RfD was calculated by dividing the POD by a composite UF of 300 to account for the 

17 extrapolation from animals to humans (10), for interindividual differences in human susceptibility 

18 (10), and for deficiencies in the toxicity database (3) as shown in Table ES-1.

19 Table ES- 1. Summary of proposed reference dose (RfD) derivation 

Critical effect Point of departure UF Chronic RfD

Neurodevelopmental impairment

PND 5-11 rat study, gavage

(Chen et al., 2012)

BMDL1SD: 0.06 mg/kg-day 300 2 x 10-4 mg/kg-day

20 Confidence in the Chronic Oral RfD

21 The overall confidence in the RfD is medium. Confidence in the principal study (Chen et al., 

22 2012) is medium-to- high. The design, conduct, and reporting of this neurodevelopmental study 

23 was excellent and a wide variety of neurotoxicity endpoints were measured. Several subchronic 

24 and developmental studies covering a wide variety of endpoints are available; however, the lack of 

25 a multi-generation toxicity study with exposure throughout development is not available. 

26 Therefore, confidence in the database is medium.

27 Effects Other Than Cancer Following Inhalation Exposure

28 In animals, inhalation exposure to benzo[a]pyrene has been shown to result in 

29 developmental and reproductive toxicity. Studies in rats following inhalation exposure show 

30 decreased fetal survival and brain effects in offspring, and decreased testes weight and sperm 

31 counts in adult animals. Overall, the available human PAH mixtures studies report developmental 
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1 and reproductive effects that are generally analogous to those observed in animals, and provide 

2 qualitative, supportive evidence for the hazards associated with benzo[a]pyrene exposure.

3 Developmental toxicity, represented by decreased fetal survival, was chosen as the basis for 

4 the proposed RfC as the available data indicate that developmental effects represent a sensitive 

5 hazard of benzo[a]pyrene exposure. The developmental inhalation study in rats by Archibong et al. 

6 (2002) and the observed decreased fetal survival following exposure to benzo[a]pyrene on GD 11-

7 20 were used to derive the overall RfC. The LOAEL of 25 μg/m3 based on decreased fetal survival 

8 was selected as the POD. The LOAEL was adjusted to account for the discontinuous daily exposure 

9 to derive the PODADJ and the human equivalent concentration (HEC) was calculated from the PODADJ

10 by multiplying by the regional deposited dose ratio (RDDRER) for extrarespiratory (i.e., systemic) 

11 effects, as described in Methods for Derivation of Inhalation Reference Concentrations and 

12 Application of Inhalation Dosimetry (U.S. EPA, 1994b). These adjustments resulted in a PODHEC of 

13 4.6 μg/m3 which was used as the POD for RfC derivation.

14 The RfC was calculated by dividing the POD by a composite UF of 3000 to account for

15 toxicodynamic differences between animals and humans (3), interindividual differences in human 

16 susceptibility (10), LOAEL-to-NOAEL extrapolation (10), and deficiencies in the toxicity database 

17 (10) as shown in Table ES-2.

18 Table ES- 2. Summary of proposed reference concentration (RfC) derivation

Critical effect Point of departure* UF Chronic RfC

Decreased fetal survival

GD 11-20 rat study

(Archibong et al., 2002)

LOAELHEC: 4.6 μg/m3(4.6 x 10-3 mg/m3) 3000 2 x 10-6 mg/m3

19 * The POD was adjusted for continuous daily exposure: PODADJ= POD × hours exposed per day/24 hours, and 

20 was further adjusted to a human equivalent concentration by adjusting by the regional deposited dose ratio 

21 calculated using MPPD software (see Section 2.2.2. and Appendices)

22 Confidence in the Chronic Oral RfC

23 The overall confidence in the RfC is low-to-medium. Confidence in the principal study 

24 (Archibong et al., 2002) is medium. The conduct and reporting of this developmental dietary study 

25 were adequate; however, a NOAEL was not identified. Confidence in the database is low due to the 

26 lack of a multigeneration toxicity study and the lack of information on diverse toxicity endpoints 

27 following subchronic and chronic inhalation exposure. However, confidence in the RfC is bolstered 

28 by consistent systemic effects observed by the oral route (including reproductive and 

29 developmental effects) and similar effects observed in human populations exposed to PAH 

30 mixtures.
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1 Evidence for Human Carcinogenicity

2 Under EPA’s Guidelines for Carcinogen Risk Assessment (U.S. EPA, 2005a), benzo[a]pyrene is 

3 “carcinogenic to humans” based on strong and consistent evidence in animals and humans. The 

4 evidence includes an extensive number of studies demonstrating carcinogenicity in multiple animal 

5 species exposed via all routes of administration and increased cancer risks, particularly in the lung 

6 and skin, in humans exposed to different PAH mixtures containing benzo[a]pyrene. Mechanistic 

7 studies provide strong supporting evidence that links the metabolism of benzo[a]pyrene to DNA8 reactive agents with key mutational events in genes that can lead to tumor development. These 

9 events include formation of specific DNA adducts and specific mutations in oncogenes and tumor 

10 suppressor genes that have been observed in humans exposed to PAH mixtures. This combination 

11 of human, animal, and mechanistic evidence provides the basis for characterizing benzo[a]pyrene 

12 as “carcinogenic to humans.”

13 Quantitative Estimate of Carcinogenic Risk From Oral Exposure

14 Lifetime oral exposure to benzo[a]pyrene has been associated with forestomach, liver, oral 

15 cavity, jejunum or duodenum, and auditory canal tumors in male and female Wistar rats, 

16 forestomach tumors in male and female Sprague-Dawley rats, and forestomach, esophagus, tongue, 

17 and larynx tumors in female B6C3F1 mice (male mice were not tested). Less-than-lifetime oral 

18 exposure to benzo[a]pyrene has also been associated with forestomach tumors in more than 

19 10 additional bioassays with several strains of mice. The Kroese et al. (2001) and Beland and Culp 

20 (1998) studies were selected as the best available studies for dose-response analysis and 

21 extrapolation to lifetime cancer risk following oral exposure to benzo[a]pyrene. These studies 

22 included histological examinations for tumors in many different tissues, contained three exposure 

23 levels and controls, contained adequate numbers of animals per dose group(~50/sex/group), 

24 treated animals for up to 2 years, and included detailed reporting methods and results (including 

25 individual animal data). 

26 EPA used the multistage-Weibull model for the derivation of the oral slope factor because it 

27 incorporates the time at which death-with-tumor occurred and can account for differences in 

28 mortality observed between the exposure groups. Using linear extrapolation from the BMDL10, 

29 human equivalent oral slope factors were derived for each gender/tumor site combination (slope 

30 factor = 0.1/BMDL10) reported by Kroese et al. (2001) and Beland and Culp (1998). The oral slope 

31 factor of 1 per mg/kg-day based on the tumor response in the alimentary tract (forestomach, 

32 esophagus, tongue, and larynx) of female B6C3F1 mice (Beland and Culp, 1998) was selected as the 

33 factor with the highest value (most sensitive) among a range of slope factors derived.

34 Quantitative Estimate of Carcinogenic Risk From Inhalation Exposure

35 Inhalation exposure to benzo[a]pyrene has been associated with squamous cell neoplasia in 

36 the larynx, pharynx, trachea, esophagus, and forestomach, of male Syrian golden hamsters exposed 

37 to benzo[a]pyrene condensed onto NaCl particles (Thyssen et al., 1981). Supportive evidence for 
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1 the carcinogenicity of inhaled benzo[a]pyrene comes from additional studies with hamsters 

2 exposed to benzo[a]pyrene via intratracheal instillation. The Thyssen et al. (1981) bioassay 

3 represents the only available data that exhibits a dose-response relationship for cancer from 

4 inhaled benzo[a]pyrene.

5 A time-to-tumor dose-response model was fit to the time-weighted average exposure 

6 concentrations and the individual animal occurrence data for tumors in the larynx, pharynx, 

7 trachea, esophagus, and forestomach. The inhalation unit risk of 5 × 10-4 per μg/m3 was calculated 

by linear extrapolation (slope factor = 0.1/BMCL10) from a BMCL10 of 0.20 mg/m3 8 for the 

9 occurrence of upper respiratory and upper digestive tract tumors in male hamsters chronically 

10 exposed by inhalation to benzo[a]pyrene (Thyssen et al., 1981).

11 Quantitative Estimate of Carcinogenic Risk From Dermal Exposure

12 Skin cancer in humans has been documented to result from occupational exposure to 

13 complex mixtures of PAHs including benzo[a]pyrene, such as coal tar, coal tar pitches, unrefined 

14 mineral oils, shale oils, and soot. In animal models, numerous dermal bioassays have demonstrated 

15 an increased incidence of skin tumors with increasing dermal exposure of benzo[a]pyrene in all 

16 species tested (mice, rabbits, rats, and guinea pigs), although most benzo[a]pyrene bioassays have 

17 been conducted in mice. The analysis in this assessment focuses on chronic carcinogenicity 

18 bioassays in several strains of mice demonstrating increasing incidence of benign and malignant 

19 skin tumors following repeated dermal exposure to benzo[a]pyrene for the animals’ lifetime. 

20 The Poel (1959) and Sivak et al. (1997) studies were selected as the best available studies 

21 for dose-response analysis and extrapolation to lifetime cancer risk following dermal exposure to 

22 benzo[a]pyrene. Both studies included at least three exposure levels (including several low doses), 

23 group sizes of 30−50 mice, and reporting of intercurrent mortality. Following the modeling, the 

24 BMDL10 was adjusted for interspecies differences by allometric scaling. The dermal slope factor of 

25 0.005 per μg/day was calculated by linear extrapolation (slope factor = 0.1/BMDL10-HED) from the 

26 human equivalent POD for the occurrence of skin tumors in male mice chronically exposed 

27 dermally to benzo[a]pyrene. As this slope factor has been developed for a local effect, it is not 

28 intended to estimate systemic risk of cancer following dermal absorption of benzo[a]pyrene into 

29 the systemic circulation.

30 Susceptible Populations and Lifestages

31 Benzo[a]pyrene has been determined to be carcinogenic by a mutagenic mode of action in 

32 this assessment. According to the Supplemental Guidance for Assessing Susceptibility from Early Life 

33 Exposure to Carcinogens (U.S. EPA, 2005b), individuals exposed during early life to carcinogens with 

34 a mutagenic mode of action are assumed to have an increased risk for cancer. The oral slope factor 

35 of 1 per mg/kg-day, inhalation unit risk of 0.005 per μg/day, and dermal slope factor of 0.004 per 

36 μg/day for benzo[a]pyrene, calculated from data applicable to adult exposures, do not reflect 

37 presumed early life susceptibility to this chemical. Although some chemical-specific data exist for 
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1 benzo[a]pyrene that demonstrate increased early life susceptibility to cancer, these data were not 

2 considered sufficient to develop separate risk estimates for childhood exposure. In the absence of 

3 adequate chemical-specific data to evaluate differences in age-specific susceptibility, the 

4 Supplemental Guidance (U.S. EPA, 2005b) recommends that ADAFs be applied in estimating cancer 

5 risk. The ADAFs are 10- and 3-fold adjustments that are combined with age specific exposure 

6 estimates when estimating cancer risks from early life (<16 years of age) exposures to 

7 benzo[a]pyrene. 

8 Regarding effects other than cancer, there are epidemiological studies that report

9 associations between developmental effects (decreased postnatal growth, decreased head 

10 circumference and neurodevelopmental delays) and internal biomarkers of exposure to 

11 benzo[a]pyrene. 

12 Studies in animals also indicate alterations in neurological development and heightened 

13 susceptibility to reproductive effects following gestational or early postnatal exposure to 

14 benzo[a]pyrene.

15 Key Issues Addressed in Assessment

16 The dermal slope factor was developed based on data in animals, and because there is no 

17 established methodology for extrapolating dermal toxicity from animals to humans. As such,

18 several alternative approaches were evaluated (See Appendix C in Supplemental Information).

19 Allometric scaling using body weight to the 3⁄4 power was selected based on known species 

20 differences in dermal metabolism and penetration of benzo[a]pyrene. 
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LITERATURE SEARCH STRATEGY|STUDY SELECTION

1 The literature search strategy used to identify primary, peer-reviewed literature pertaining 

2 to benzo[a]pyrene was conducted using the databases and keywords listed in Table LS-1. 

3 References that were evaluated in other Agency and international health assessments were also 

4 examined. A comprehensive literature search was last conducted in February 2012. 

5 Figure LS-1 depicts the literature search, study selection strategy, and the number of 

6 references obtained at each stage of literature screening. Approximately 20,700 references were 

7 identified with the initial keyword search. Based on a secondary keyword search followed by a 

8 preliminary manual screen of titles or abstracts by a toxicologist, approximately 1,190 references 

9 were identified that provided information potentially relevant to characterizing the health effects 

10 or physical and chemical properties of benzo[a]pyrene. A more detailed manual review of titles, 

11 abstracts, and/or papers was then conducted. Notable exclusions from the Toxicological Review 

12 are large numbers of animal in vivo or in vitro studies designed to identify potential therapeutic 

13 agents that would prevent the carcinogenicity or genotoxicity of benzo[a]pyrene and toxicity 

14 studies of benzo[a]pyrene in nonmammalian species (e.g., aquatic species, plants).
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1 Table LS- 1. Details of the literature search strategy employed

Database Keywordsa

Pubmed Chemical name (CASRN): benzo[a]pyrene (50-32-8)a

Synonyms: benzo[d,e,f]chrysene, benzo[def]chrysene, 3,4-benzopyrene, 1,2-benzpyrene, 3,4-bp, 

benz(a)pyrene, 3,4-benzpyren, 3,4-benzpyrene, 4,5-benzpyrene, 6,7-benzopyrene, benzopirene, 

benzo(alpha)pyrene

Initial keyword search

Standard toxicology search

Toxicity (including duration, effects to children and occupational exposure); development; 

reproduction; teratogenicity; exposure routes; pharmacokinetics; toxicokinetics; metabolism; body 

fluids; endocrinology; carcinogenicity; genotoxicity; antagonists; inhibitors

Secondary keyword searchb

Chemical-specific keywords

Cancer; genotoxicity, neurotoxicity, immunotoxicity, reproductive toxicity, developmental toxicity

Additional keywords

lung OR skin combined AND tumor, neoplasm, papilloma, OR carcinoma; leukemia; forestomach; 

tongue; auditory canal; esophagus; larynx; pharynx; fertility; sperm; epididymis; seminiferous; 

testosterone; cervical hyperplasia; corpus luteum; estrous; testicular, ovarian OR thymus with 

atrophy; weight AND testis, ovary, thymus, spleen, pup; spleen AND cells, lymphocytes; 

immunoglobulin; immunosuppression; motor; functional observational battery; neurobehavioral; 

rotarod; nerve conduction; locomotor; neuromuscular; neurodevelopment; cognitive; learning; 

memory; righting.

Toxcenter

Toxline

TSCATS Searched by chemical names (including synonyms) and CASRNs

ChemID

Chemfinder

CCRIS

HSDB

GENETOX

RTECS

a Keywords and synonyms were applied to the Pubmed, Toxcenter, and Toxline databases. 

b Secondary keywords were selected from an understanding of the targets of benzo[a]pyrene toxicity gained 

from review of papers identified in literature searches conducted at the start of document development and 

relevant review documents.

2
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1

2 Figure LS- 1. Study selection strategy.
Toxicological Review of Benzo[a]pyrene

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1 Selection of studies for inclusion in the Toxicological Review was based on consideration of 

2 the extent to which the study was informative and relevant to the assessment and general study 

3 quality considerations. In general, the relevance of health effect studies was evaluated as outlined 

4 in the Preamble and EPA guidance (A Review of the Reference Dose and Reference Concentration 

5 Processes (U.S. EPA, 2002) and Methods for Derivation of Inhalation Reference Concentrations and 

6 Application of Inhaled Dosimetry (U.S. EPA, 1994)). The reasons for excluding epidemiological and 

7 animal studies from the approximately 1,190 references identified by the keyword search are 

8 provided in Figure LS-1.

9 The available studies examining the health effects of benzo[a]pyrene exposure in humans 

10 are discussed and evaluated in the hazard identification sections of the assessment (Section 1), with 

11 specific limitations of individual studies and of the collection of studies noted. The common major 

12 limitation of the human epidemiological studies (with respect to identifying potential adverse 

13 health outcomes specifically from benzo[a]pyrene) is that they all involve exposures to complex 

14 mixtures containing other PAHs and other compounds. The evaluation of the epidemiological 

15 literature focuses on studies in which possible associations between external measures of exposure 

16 to benzo[a]pyrene or biomarkers of exposure to benzo[a]pyrene (e.g., benzo[a]pyene-DNA adducts 

17 or urinary biomarkers) and potential adverse health outcomes were evaluated. Pertinent 

18 mechanistic studies in humans (e.g., identification of benzo[a]pyene-DNA adducts and 

19 characteristics of mutations in human tumors) were also considered in assessing the weight of 

20 evidence for the carcinogenicity of benzo[a]pyrene. 

21 The health effects literature for benzo[a]pyrene is extensive. All animal studies of 

22 benzo[a]pyrene involving repeated oral, inhalation, or dermal exposure that were considered to be 

23 of acceptable quality, whether yielding positive, negative, or null results, were considered in 

24 assessing the evidence for health effects associated with chronic exposure to benzo[a]pyrene. In 

25 addition, animal toxicity studies involving short-term duration and other routes of exposure were 

26 evaluated to inform conclusions about health hazards. 

27 The references considered and cited in this document, including bibliographic information 

28 and abstracts, can be found on the Health and Environmental Research Online (HERO) website2

29 (Insert link when available).

 

2HERO is a database of scientific studies and other references used to develop EPA’s risk assessments aimed

at understanding the health and environmental effects of pollutants and chemicals. It is developed and 

managed in EPA’s Office of Research and Development (ORD) by the National Center for Environmental 

Assessment (NCEA). The database includes more than 300,000 scientific articles from the peer-reviewed 

literature. New studies are added continuously to HERO.
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1. HAZARD IDENTIFICATION

1 1.1. Synthesis of Evidence

2 NOTE: In the environment, benzo[a]pyrene occurs in conjunction with other 

3 structurally related chemical compounds known as polycyclic aromatic 

hydrocarbons (PAHs). 4 3 Accordingly, there are no epidemiologic studies designed to 

5 solely investigate the effects of benzo[a]pyrene. There are, however, many 

6 epidemiologic studies that have investigated the effects of exposure to PAH 

7 mixtures. Benzo[a]pyrene is universally present in these mixtures and is often used 

8 as an indicator chemical to measure exposure to PAH mixtures.

9 1.1.1. Developmental Toxicity

10 Human and animal studies provide evidence for PAH and benzo[a]pyrene-induced 

11 developmental effects. Effects on fetal survival, postnatal growth, and development have been 

12 demonstrated in human populations exposed to PAH mixtures during gestation. Animal studies 

13 demonstrate various effects including changes in pup weight, blood pressure, fertility, reproductive 

14 organ weight and histology, and neurological function in gestationally or early postnatally treated 

15 animals.

16 Altered Birth Outcomes

17 Human and animal studies provide evidence that benzo[a]pyrene exposure may lead to 

18 altered outcomes reflecting growth and development in utero or early childhood. Two cohort 

19 studies in pregnant women in China and the United States examined cord blood levels of 

20 benzo[a]pyrene-DNA adducts in relation to measures of child growth following exposure to PAH 

21 mixtures (Tang et al., 2006; Perera et al., 2005a; 2004) (Table 1-1). In the Chinese cohort, high 

22 benzo[a]pyrene-adduct levels were associated with reduced weight at 18, 24, and 30 months of age, 

23 but not at birth (Tang et al., 2006). In the U.S. cohort, an independent effect on birth weight was not 

24 observed with either benzo[a]pyene-adducts or environmental tobacco smoke (ETS); however, a 

25 doubling of cord blood adducts in combination with ETS exposure in utero was seen, corresponding

26 to an 8% reduction in birth weight (Perera et al., 2005a; 2004). Environmental tobacco smoke, also 

27 called secondhand smoke, is the smoke given off by a burning tobacco product and the smoke 

28 exhaled by a smoker that contains over 7,000 chemicals including benzo[a]pyrene. No associations 

29 were seen with birth length (or height at later ages) in either of these cohort studies.

 

3 PAHs are a large class of chemical compounds formed during the incomplete combustion of organic matter.
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1 In animals (Table 1-2), reduced bodyweight in offspring has been noted in some 

2 developmental studies. Decreases in body weight (up to 13%) were observed in mice following 

3 prenatal gavage exposure (GD 7-16), and as time from exposure increased (PND 20 to 42) the dose 

4 at which effects were observed decreased (40 to 10 mg/kg-day, respectively) (MacKenzie and 

5 Angevine, 1981). In addition, decreases in body weight (approximately 10 to 15%) were observed 

6 in rats on PNDs 36 and 71 following gavage exposure at only 2 mg/kg-day on PNDs 5-11 (Chen et 

7 al., 2012). At doses up to 1.2 mg/kg-day and follow-up to PND 30, two developmental studies in 

8 rats did not observe decrements in pup body weight following treatment from GD14-17 (Jules et al., 

9 2012; McCallister et al., 2008). Maternal toxicity was not observed in mouse or rat dams exposed to 

10 up to 160 mg/kg-day benzo[a]pyrene (Jules et al., 2012, McCallister et al., 2008; Brown et al., 2007; 

11 Kristensen et al., 1995; MacKenzie and Angevine, 1981).

12 Decreased fetal survival has also been noted in gestationally treated animals at relatively 

13 high doses by the oral and inhalation route. An approximate 40% decrease in fetal survival was 

14 noted in mouse dams treated by gavage on GD7-16 at doses of 160 mg/kg-day, but no decreases

15 were observed at 10 or 40 mg/kg-day (MacKenzie and Angevine, 1981). Several lower dose studies 

16 of rats treated on GD14-17 with doses of up to 1.2 mg/kg-day benzo[a]pyrene did not observe any 

17 difference in fetal survival (Jules et al., 2012, McCallister et al., 2008; Brown et al., 2007). By the 

18 inhalation route, fetal survival was decreased by 19% following exposure to 25 μg/m3 

19 benzo[a]pyrene on GD 11-20 (Archibong et al., 2002) and 65% following exposure to 100 μg/m3 

20 benzo[a]pyrene on GD11-21 in (Wormley et al., 2004) F344 rats. Wu et al. (2003) also evaluated 

21 fetal survival as part of a study analyzing metabolites of benzo[a]pyrene and activation of the aryl 

22 hydrocarbon receptor (AhR) and cytochrome P450 (CYP) 1A1. The study authors reported 

decreased fetal survival at 75 and 100 μg/m3, but not at 25 μg/m3 23 following exposure to 

24 benzo[a]pyrene on GD 11-20. This study did not report number of dams or litters and no numerical 

25 data were reported.

26 Fertility in Offspring

27 Several studies suggest that gestational exposure to maternal tobacco smoke decreases the 

28 future fertility of female offspring (Ye et al., 2010; Jensen et al., 1998; Weinberg et al., 1989) (Table 

29 1-1). In animal models, marked effects on the development of male and female reproductive organs 

30 and the fertility of animals exposed gestationally has been demonstrated (Kristensen et al., 1995; 

31 MacKenzie and Angevine 1981) (Table 1-2). In two studies examining reproductive effects in mice, 

32 decreased fertility and fecundity in F1 animals was observed following exposure to doses ≥ 10 

33 mg/kg-day during gestation (Kristensen et al., 1995; MacKenzie and Angevine, 1981). When F1 

34 females were mated with untreated males, a dose-related decrease in fertility of > 30% was 

35 observed, in addition to a 20% decrease in litter size starting at the lowest dose tested (10 mg/kg36 day). A dose-related decrease in fertility was also observed in male mice treated gestationally with 

37 benzo(a)pyrene. At the lowest dose tested (10 mg/kg-day), a 35% decrease in fertility was 
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1 observed when gestationally exposed animals were mated with untreated females (MacKenzie and 

2 Angevine, 1981). Similar effects on fertility were observed in another developmental study in mice 

3 (Kristensen et al., 1995). F1 females (bred continuously for 6 months) in this study had 63% fewer 

4 litters, and litters were 30% smaller as compared to control animals. The fertility of male offspring 

5 was not assessed in this study.

6 Reproductive Effects in Offspring

7 The above mentioned studies also demonstrated dose-related effects on male and female 

8 reproductive organs in animals exposed gestationally to benzo[a]pyrene (Table 1-2). Testicular 

9 weight was decreased and atrophic seminiferous tubules and vacuolization was increased at ≥ 10 

10 mg/kg-day in male mice exposed to benzo[a]pyrene gestationally from GD 7-16; severe atrophic 

11 seminiferous tubules were observed at 40 mg/kg-day (MacKenzie and Angevine, 1981). 

12 In female mice treated with doses ≥ 10 mg/kg-day during gestation, ovarian effects were 

13 observed including decreases in ovary weight, numbers of follicles, and corpora lutea (Kristensen et 

14 al., 1995; MacKenzie and Angevine, 1981). Specifically, ovary weight in F1 offspring was reduced 

15 30% following exposure to 10 mg/kg-day benzo[a]pyrene (Kristensen et al., 1995) while in another 

16 gestational study at the same dose level, ovaries were so drastically reduced in size (or absent) that 

17 they were not weighed (MacKenzenie and Angevine 1981). Hypoplastic ovaries with few or no 

18 follicles and corpora lutea (numerical data not reported), and ovaries with few or no small, 

19 medium, or large follicles and corpora lutea (numerical data not reported) have also been observed 

20 in mouse offspring exposed gestationally to benzo[a]pyrene (MacKenzenie and Angevine 1981; 

21 Kristensen et al., 1995).

22 Cardiovascular Effects in Offspring

23 Increased systolic and diastolic blood pressure was observed in adult animals following 

24 gestational treatment with benzo[a]pyrene (Jules et al., 2012) (Table 1-2). Approximate elevations 

25 in systolic and diastolic blood pressure of 20- 30% and 50-80% were noted in the 0.6 mg/kg-day 

26 and 1.2 mg/kg-day dose groups, respectively. Heart rate was decreased at 0.6 mg/kg-day, but was 

27 increased at 1.2 mg/kg-day.
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1 Table 1-1. Evidence pertaining to the developmental effects of benzo[a]pyrene

2 in humans

Reference and Study Design:

Study Type/Period/Study 

Size/Location/Exposure Estimate Results

Tang et al., 2006

Pregnancy cohort

150 non-smoking women that delivered 

babies between March 2002 – June 2002

Tongliang, China

Exposure: mean hours per day exposed to

environmental tobacco smoke 0.42 (SD 

1.19); lived within 2.5 km of power plant 

that operated from December 2001 – May 

2002; B[a]P-DNA adducts from maternal and 

cord blood samples; cord blood mean 0.33 

(SD 0.14) (median 0.36) adducts/10-8

nucleotides; maternal blood mean 0.29 (SD 

0.13) adducts/10-8nucleotides

Relation between cord blood B[a]P-DNA adducts and logtransformed weight and height

Weight

Beta (pvalue)

Length (Height)

Beta (p-value)

Birth –0.007 

(0.73)

–0.001 (0.89)

18 months –0.048 

(0.03)

–0.005 (0.48)

24 months –0.041 

(0.027)

–0.007 (0.28)

30 months –0.040 

(0.049)

–0.006 (0.44)

Adjusted for environmental tobacco smoke, sex of child, maternal 

height, maternal weight, cesarean section delivery, maternal head 

circumference, and gestational age (for measures at birth)

Perera et al. (2005a; 2004)

Pregnancy cohort

214 African-American and Dominican nonsmoking women that delivered babies 

between April 1998 – October 2002; 

approximately 40% with smoker in the 

home

New York, United States

Exposure: B[a]P-DNA adducts in cord blood 

samples; mean 0.22 (SD 0.14) adducts/10-8

nucleotides; median of detectable values 

0.36 adducts/10-8nucleotides)

Relation between cord blood B[a]P-DNA adducts and logtransformed birth weight and length: 

Weight

Beta (pvalue)

Length

Beta (p-value)

Interaction 

term 

–0.088

(0.05)

–0.014 (0.39)

B[a]P-DNA 

adducts 

–0.012 

(0.49)

–0.048 (0.64)

ETS in home –0.003

(0.90)

–0.007 (0.32)

Adjusted for ethnicity, sex of newborns, maternal body 

mass index, dietary PAHs, gestational age. ETS = 

environmental tobacco smoke
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1 Table 1-2. Evidence pertaining to the developmental effects of benzo[a]pyrene

2 in animals

Study Design and Reference Results

Birth outcomes

MacKenzie and Angevine, 1981

CD-1 mice, 30 or 60 F0 females/ 

dose

0, 10, 40, or 160 mg/kg-d by 

gavage 

GD 7–16

↓ Number of F0 females with viable litters: 46/60, 21/30, 44/60, 13/30*

↓ F1 body weight at PND 20

Response relative to control: 0, 4, -5*, -13*

↓ F1 body weight at PND 42 

Response relative to control: 0, -6*, -6*, -10*

Kristensen et al., 1995

NMRI mice, 9 F0 females/dose

0 or 10 mg/kg-d by gavage 

GD 7–16 

Exposed F0 females showed no gross signs of toxicity and no effects on 

fertility (data not reported)

Jules et al., 2012

Long-Evans rats, 6-17 F0

females/dose

0, 0.15, 0.3, 0.6, or 1.2 mg/kg-day 

by gavage 

GD 14-17

No overt signs of toxicity in dams or offspring, differences in pup body 

weight, or number of pups per litter

McCallister et al. (2008)

Long Evans Hooded rats, 5-

6/group

0 or 0.3 mg/kg-day by gavage

GD14- 17

No difference in number of pups/litter

No overt maternal or pup toxicity

No difference in liver:body weight

Increased brain:body weight ratio at PND15 and 30

Brown et al. (2007) 

Long Evans Hooded rats, 6/group

0, 0.025 or 0.15 mg/kg-day by 

gavage

GD 14- 17

No difference in number of pups/litter or overt maternal or pup toxicity

Chen et al. (2012)

Sprague-Dawley rats, 20 pups (10 

male and 10 female)/group

0, 0.02, 0.2, or 2 mg/kg-day by 

gavage

PND5-PND11

Statistically significant decrease in pup bodyweight (approximate 10-15% 

decrease) at 2 mg/kg-day measured on PND36 and 71 

No differences among treatment groups in developmental milestones: 

incisor eruption, eye opening, development of fur, testis decent or vaginal 

opening

Fertility in offspring

MacKenzie and Angevine, 1981

CD-1 mice, 30 or 60 F0 females/ 

dose

0, 10, 40, or 160 mg/kg-d by 

gavage 

GD 7-16

↓ Number of F1 females with viable litters: 35/35, 23/35*, 0/55*, 0/20*

↓F1 female fertility index (females pregnant/females exposed to males x 

100): 100 , 66*, 0*, 0*

↓ F1 male fertility index (females pregnant/females exposed to males x 

100): 80, 52*, 5*, 0*
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↓F1 litter size (20%) at 10 mg/kg-day (no litters were produced at high 

doses)

↓ size or absence of F1 ovaries (weights not collected)

hypoplastic ovaries with few or no follicles and corpora lutea (numerical 

data not reported)

↓ testicular weight 

% change from control: 0, -42, -82, ND (statistical significance not reported) 

↑ atrophic seminiferous tubules and vacuolization at ≥ 10 mg/kg-day; 

severe atrophic seminiferous tubules at 40 mg/kg-day (numerical data not 

reported)

Kristensen et al., 1995

NMRI mice, 9 F0 females/dose

0 or 10 mg/kg-d by gavage 

GD 7–16 

↓ Number of F1 litters (63%) 

↓F1 litter size (30%)

↓ ovary weight (31%) in F1 females

Few or no small, medium, or large follicles and corpora lutea (numerical 

data not reported)

Archibong et al., 2002 

F344 rats, 10 females/group

0, 25, 75, or 100 μg/m3 nose-only 

inhalation for 4 hrs/day

GD 11–20

↓ Fetal survival ([pups/litter]/[implantation sites/litter] x 100)

% fetal survival 97, 78*, 38*, and 34*% 

↓ pup weight 0, 0, 14*, and 16*%

Wormley et al., 2004

F344 rats, 10 females/group

0 or 100 μg/m3 nose-only

inhalation for 4 hrs/day

GD 11–21

↓ 65% decrease in pups/litter

Reproductive effects in offspring

MacKenzie and Angevine, 1981

CD-1 mice, 30 or 60 F0 females/ 

dose

0, 10, 40, or 160 mg/kg-d by 

gavage 

GD 7-16

↓ size or absence of F1 ovaries (weights not collected)

hypoplastic ovaries with few or no follicles and corpora lutea (numerical 

data not reported) 

↓ testicular weight 

% change from control: 0, -42, -82, ND (statistical significance not reported) 

↑ atrophic seminiferous tubules and vacuolization at ≥ 10 mg/kg-day; 

severe atrophic seminiferous tubules at 40 mg/kg-day (numerical data not 

reported)

Kristensen et al., 1995

NMRI mice, 9 F0 females/dose

0 or 10 mg/kg-d by gavage 

GD 7–16 

↓ ovary weight (31%) in F1 females

Few or no small, medium, or large follicles and corpora lutea (numerical 

data not reported)

Cardiac Effects in Offspring

Jules et al., 2012 ↑ systolic blood pressure (measured at PND53)
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Long-Evans rats, 6-17 F0

females/dose

0, 0.15, 0.3, 0.6, or 1.2 mg/kg-day 

by gavage 

GD 14-17

~20%* increase at 0.6 mg/kg-day

~50% *increase at 1.2 mg/kg-day

(other dose groups not reported)

↑ diastolic blood pressure (measured at PND53)

33%* increase at 0.6 mg/kg-day

83% *increase at 1.2 mg/kg-day

(other dose groups not reported)

Statistically significant increase in heart rate at 0.6 mg/kg-day and 

statistically significant decrease at 1.2 mg/kg-day

1 Figure 1-1. Exposure-response array for developmental effects following oral 

2 exposure

3

4 Neurodevelopment Effects

5 There is evidence in humans and animals that benzo[a]pyrene induces developmental 

6 neurotoxicity. In addition to the persistent reductions in cognitive ability observed in epidemiology 
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1 studies of prenatal PAH exposure, the two epidemiology studies that examined benzo[a]pyrene2 specific measures observed effects on neurodevelopment and behavior in young children. Altered 

3 learning and memory, motor activity and anxiety, and electrophysiological changes have also been 

4 observed in animals following oral and inhalation exposure to benzo[a]pyrene.

5 The mammalian brain undergoes a period of rapid brain growth during the last 3 months of 

6 pregnancy through the first 2 years of life in humans (Dobbing and Sands, 1973, 1979) and the first 

7 1–2 weeks of life in the rat and mouse neonate (Chen et al., 2011a). This period is characterized by 

8 the maturation of axonal and dendritic outgrowth and the establishment of neuronal connections. 

9 Also during this critical period, animals acquire many new motor and sensory abilities (Kolb and 

10 Whishaw, 1989). There is a growing literature of animal studies that shows subtle changes in motor 

11 and cognitive function following acute or repeated perinatal or lactation exposure to 

12 benzo[a]pyrene (Bouayed et al., 2009; McCallister et al., 2008; Wormley et al., 2004).

13 Cognitive function

14 Head circumference at birth is associated with measures of intelligence in children, even 

15 among term infants (Broekman et al., 2009; Gale et al., 2006). The two pregnancy cohort studies 

16 that examined maternal or cord blood levels of benzo[a]pyrene-DNA adducts in relation to head 

17 circumference provide some evidence of an association, most strongly within the context of an 

18 interaction with environmental tobacco smoke (Tang et al., 2006; Perera et al., 2005a; 2004) (Table 

19 1-3). One of these studies also examined aspects of neurodevelopment in two-year old children in 

20 China (Tang et al., 2008). In this follow-up of 2-year old children in an area surrounding a coal 

21 power plant, average scores on the Gesell Development Scale decreased with increased exposure 

22 (measured by benzo[a]pyrene-DNA adducts in cord blood) with similar effects seen in the motor, 

23 adaptive, and language domains (approximately 15 point decrement per unit increase in exposure). 

24 Animal studies have also provided evidence of altered learning and memory behaviors 

25 following lactational or postnatal exposure to benzo[a]pyrene (Bouayed et al., 2009b, Chen et al., 

26 2012) (Table 1-4). In mice, working memory was measured using the Y-maze spontaneous 

27 alternation test (Bouayed et al, 2009b). This test records alternations between arm entries in a Y28 shaped maze as a measure of memory, as rodents typically prefer to investigate a new arm of the 

29 maze. This test may also be used to measure exploratory behavior in rodents and thus may be 

30 reflective of changes in anxiety-like behavior. A decrease in working memory was evident in mice, 

31 as exhibited by significant increases in spontaneous alternations in the Y-maze test in mice on PND 

32 40 following lactational exposure to 2 mg/kg-day benzo[a]pyrene (but not 20 mg/kg-day) from 

33 PND 0 to PND 14 (Bouayed et al., 2009b). The total number of arm entries in the Y-maze was 

34 unaffected by lactational exposure. In rats, spatial learning and memory was measured using the 

35 Morris water maze, which measures the ability of a rat to navigate to a target platform using 

36 external spatial cues (Chen et al., 2012). Increased escape latency, decreased time in the target 

37 quadrant, and decreased number of platform crossings were observed in PND 39 – PND 40 rats 

38 following postnatal exposure to 2 mg/kg-day benzo[a]pyrene (Chen et al., 2012). These effects 
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1 were more pronounced in animals tested at PND 74 –PND 75. No difference in swim speed was 

2 observed between treatment groups, suggesting the difference observed is not attributable to 

3 general motor impairment.

4 Negative geotaxis and surface righting are discrete endpoints routinely used as part of a 

5 neurobehavioral test battery to assess acquisition of developmental milestones. In typical 

protocols, animals are tested on successive days (usually PND 3-7+ and PND 6-9+ 6 , respectively) and 

successful acquisition of these phenotypes is indicated when righting occurs within 7 ~2 seconds or 

orienting 180° occurs within 8 ~60 seconds (although latency for these events is often reported), 

9 respectively; in rats, both phenotypes are nearly always established (aka fully matured) before PND 

10 10. Chen et al. (2012) performed these tests in a slightly atypical manner as quantitative measures 

11 of sensorimotor function at PND 12 and beyond, with control animals already to right within 0.8-

12 1.8 seconds and able to orient 180° within 5-9 seconds. Although informative in terms of possible 

13 developmental delays, the sensitivity of these measures at these later timepoints has not been well 

14 established in the literature. Specifically, statistically significant differences observed by Chen et al. 

(2012) in the surface righting test were on the order of 15 ~0.2-0.3 seconds and in the negative 

geotaxis test, 16 ~3-4 seconds, with no automated recording of latency (such as use of video 

17 recordings). Additionally, male and female rats (which often show differences in the maturation of 

18 these developmental landmarks) were pooled for these measures. Due to these uncertainties, EPA 

19 considered the elevated plus maze and Morris water maze tests to be the most informative and 

20 appropriate measures of neurobehavioral function performed by Chen et al. (2012). The open field 

21 tests were considered less informative as male and female data were pooled and the test paradigm 

22 could not separate effects on motor activity from anxiety responses.

23 Neuromuscular function, coordination, and motor activity

24 Motor behavior, assessed by locomotion, reaching, balance, comprehension, drawing and 

25 hand control was one of the specific domains assessed in the Chinese pregnancy cohort evaluated 

26 by Tang et al. (2008). In children aged 2 years, decreased scores were seen in relation to increasing 

27 benzo[a]pyrene-DNA adducts measured in cord blood, with a Beta per unit increase in adducts of –

28 16 (p = 0.004), and an approximate two-fold increased risk of development delay per unit increase 

29 in adducts (Table 1-3). 

30 In laboratory animals (Table 1-4), impaired neuromuscular function and coordination have 

31 been consistently observed in mice lactationally exposed to ≥2mg/kg-day benzo[a]pyrene from 

32 PND 0 to PND 14 (Bouayed et al., 2009b) and in rat pups postnatally exposed to ≥0.02 mg/kg-day 

33 benzo[a]pyrene from PND5 to PND11 (Chen et al., 2012). In the righting reflex test, significant 

34 increases in righting time were observed in PND 3 – PND 5 mice and in PND 12 – PND 16 rats. In 

35 rats, increased righting time did not show a monotonic dose response on PND 12 or PND 14. In 

36 another test of neuromuscular function and coordination, dose-dependent increases in latency in 

37 the negative geotaxis test were observed in PND 5 – PND 9 mice and in PND 12 - PND 14 rats. The 

38 forelimb grip strength test was also evaluated in both mice and rats, but the results differed 
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1 between the species. In mice, a dose-dependent increase in duration of forelimb grip was observed 

2 on PND 9 and PND 11 during lactational exposure to benzo[a]pyrene, but not on PND 12. The 

3 Water Escape Pole Climbing test was also used to evaluate neuromuscular function and 

4 coordination in mice (Bouayed et al., 2009). No effect on climbing time was observed, whereas 

5 increased latency in pole grasping and pole escape in PND 20 male pups was observed. Increased 

6 locomotor activity on PND 69, measured using the open field test, has been reported in rats 

7 postnatally exposed to ≥0.2 mg/kg-day benzo[a]pyrene on PNDs 5-11 (Chen et al., 2012). This 

8 increase in locomotor activity was observed at doses that did not cause maternal toxicity and at 

9 several weeks post-exposure, suggesting that the developmental neurotoxic effects of 

10 benzo[a]pyrene persist through neurodevelopment and may become exacerbated over time. One 

11 issue in interpreting these results, however, is that it is difficult to separate an anxiety response 

12 from effects solely on motor function with this test.

13 Anxiety

14 Anxiety, attention, and hyperactivity in children ages 6 – 7 years were examined in relation 

15 to benzo[a]pyrene-DNA adducts measured at birth in a follow-up of a pregnancy cohort study 

16 conducted in New York City (Perera et al., 2012). The associations were stronger using the 

17 measures in cord blood compared with maternal samples, with indications of a 4-fold increased 

18 risk (p=0.051) of attention problems (Table 1-3). Exposure was treated as a dichotomy (i.e., 

19 detectable compared with non-detectable levels) in these analyses. 

20 Decreased anxiety was reported in both rat and mouse pups following postnatal oral 

21 exposure to benzo[a]pyrene (Bouayed et al., 2009b, Chen et al., 2012) (Table 1-4). Anxiety-related 

22 behaviors were measured in both species using an elevated plus maze, where an increase in the 

23 time spent in the closed arms of the maze is considered evidence of anxious behavior. In mice, 

24 significant increases in the entries and time spent in open arms of the maze, as well as significantly 

25 decreased entries into closed arms of the maze, were observed on PND 32 following lactational 

26 exposure to ≥2 mg/kg-day benzo[a]pyrene (Bouayed et al., 2009b). The mice exhibited decreased 

27 latency in the elevated plus maze following lactational exposure to 20 mg/kg-day benzo[a]pyrene, 

28 while there was no exposure-related effect on the total number of times the mice entered arms of 

29 the maze. Decreased anxiety-like behaviors were also reported in rats following oral 

30 benzo[a]pyrene exposure from PND 5 to PND 11, although sex-specific differences were observed 

31 (Chen et al., 2012). In females, postnatal exposure to ≥0.2 mg/kg-day benzo[a]pyrene was 

32 associated with a significant increase in the number of open arm entries and significant decreases 

33 in the number of closed arm entries on PND 70. Significantly increased time in open arms of the 

34 maze was reported in PND 70 female rats following postnatal exposure to ≥0.02 mg/kg-day. Male 

35 rats also showed decreased anxiety-like behavior on PND 70, although the doses of benzo[a]pyrene 

36 were higher than females. A significant increase in the number of open arm entries and significant 

37 decreases in the number of closed arm entries were observed in male rat pups exposed to 2 mg/kg38 day, while a significant increase in time in the open arms of the maze was reported for males 
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1 exposed to ≥0.2 mg/kg-day. A significant decrease in latency to enter an open arm of the maze was 

2 observed in both male and female rat pups exposed to 2 mg/kg-day benzo[a]pyrene. These data 

3 indicate that oral postnatal exposure to benzo[a]pyrene resulted in decreased anxiety-like behavior 

4 in both mice and rats, with rats more sensitive than mice and female rats more sensitive than male 

5 rats.

6 Electrophysiological changes

7 Electrophysiological effects of gestational exposure to benzo[a]pyrene have been examined 

8 in animal studies through implanted electrodes in the rat cortex (Table 1-4). Inhalation exposure to 

9 0.1 mg/m3 benzo[a]pyrene resulted in reduced long-term potentiation in the dentate gyrus of male 

10 rats between PND 60 to PND 70 (Wormley et al., 2004). Oral exposure to 0.3 mg/kg-day 

11 benzo[a]pyrene resulted in decreased evoked neuronal activity in male rats following mechanical 

12 whisker stimulation between PND 90 to PND 120 (McCallister et al., 2008). The authors also noted

13 reduced spike numbers in both the short and long latency responses to whisker stimulation, 

14 although no quantitative data were presented by the study authors. These effects were observed 

15 several months post-exposure, suggesting that gestational benzo[a]pyrene exposure has long16 lasting functional effects on neuronal activity.

17 Table 1-3. Evidence pertaining to the neurodevelopmental effects of 

18 benzo[a]pyrene in humans

Reference, study design Results

Tang et al. (2008, 2006)

Tongliang, China

Pregnancy cohort

150 non-smoking women (110 for 

Developmental Quotient analysis), delivered 

March 2002 – June 2002; lived within 2.5 km 

of power plant that operated from December 

2001 – May 2002

Outcomes: head circumference at birth; 

Gesell Developmental Schedule, administered 

by physicians at 2 years of age (4 domainsa: 

motor, adaptive, language, and social); 

standardized mean score = 100 ± SD 15 (score 

< 85 = developmental delay)

Exposure: B[a]P-DNA adducts from maternal 

and cord blood samples; cord blood mean 

0.32 (SD 0.14), range 0.125 – 0.812 

adducts/10-8nucleotides

Relation between cord blood B[a]P-DNA adducts and log-transformed 

head circumference 

Beta (p-value)

Birth –0.011 (0.057)

18 months –0.012 (0.085)

24 months –0.006 (0.19)

30 months –0.005 (0.31)

High versus low, dichotomized at median, adjusted for environmental 

tobacco smoke, sex of child, maternal height, maternal weight, cesarean 

section delivery, maternal head circumference, and gestational age (for 

measures at birth)

Association between B[a]P adducts and development 

Beta (95% CI)aOR (95% CI)b

Motor –16.0 (–31.3, –0.72)* 1.91 (1.22, 2.97)* 

Adaptive –15.5 (–35.6, 4.61) 1.16 (0.76, 1.76) 

Language –16.6 (–33.7, 0.46) 1.31 (0.84, 2.05) 

Social –9.29 (–25.3, 6.70) 1.52 (0.93, 2.50) 

Average –14.6 (–28.8, –0.37)* 1.67 0.93, 3.00) 

a Linear regression of change in Developmental Quotient per unit increase 

in B[a]P adducts

b

Logistic regression of risk of developmental delay (defined as normalized 

score < 85) per 1 unit (0.1 adducts/10-8nucleotides) increase in adducts

Both analyses adjusted for sex, gestational age, maternal education, 

environmental tobacco smoke, and cord lead levels.
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Reference, study design Results

Perera et al. (2012; 2005a; 2004)

United States, New York

Pregnancy cohort

214, non-smoking women (253 for behavior 

analysis), African-American and Dominican, 

delivered April 1998 – October 2002 

Outcomes: head circumference at birth; Child

Behavior Checklist (118 items), completed by 

mothers for children ages 6 – 7 years. Two 

domains: anxious/depression, attention 

problems (normalized T-score ≤ 65 = 

borderline or clinical syndrome); also used for 

scales of anxiety problems and attention 

deficit hyperactivity problems based on DSM 

classification

Exposure: B[a]P-DNA adducts from maternal 

and cord blood samples; mean 0.22 (SD 0.14) 

adducts/10-8nucleotides; median of 

detectable values 0.36 adducts/10-8

nucleotides. 

Relation between cord blood B[a]P-DNA adducts and log-transformed 

head circumference 

Beta (p-value)

Interaction term –0.088 (0.05)

B[a]P-DNA adducts –0.012 (0.49)

ETS in home –0.003 (0.90)

High versus low, dichotomized at than 0.36 adducts/10-8nucleotides, 

adjusted for ethnicity, sex of newborns, maternal body mass index, 

dietary PAHs, gestational age. ETS = environmental tobacco smoke

Logistic regression of risk of borderline or clinical status in relation to 

detectable levels of B[a]P adducts

Maternal Cord Blood

Prevalence OR (95% CI) OR (95% CI)

Anxious/depressed 6.32 % 1.4 (0.38, 5.4) 2.6 (0.69, 9.4) 

Attention problems 6.72% 2.2 (0.74, 6.8) 4.1 (0.99, 16.6) 

Anxiety (DSM) 9.48% 2.2 (0.79, 6.1) 2.5 (0.84, 7.7) 

Attention deficit –

hyperactivity (DSM) 7.91% 1.8 (0.66, 5.1) 2.6 (0.68, 10.3) 

Exposure dichotomized as detectable (n=87 maternal, 56 cord blood 

samples) versus non-detectable; adjusted for sex, gestational age, maternal 

education, maternal IQ, prenatal environmental tobacco smoke, ethnicity, 

age, heating season, prenatal demoralization, and HOME inventory

1 Table 1-4. Evidence pertaining to the neurodevelopmental effects of 

2 benzo[a]pyrene in animals

Study Design and Reference Results

Cognitive function

Chen et al. (2012)

Sprague-Dawley rats, 20 pups (10 

male and 10 female)/group

0, 0.02, 0.2, or 2 mg/kg-day by 

gavage

PND5-PND11

Escape latency in the Morris water maze test:

PND39: significant increase in escape latency at 2 mg/kg-day only 

PND74: significant increase in escape latency at ≥0.2 mg/kg-day

Similar differences were seen at PND 36-38 or PND 71-73 time points; 

not statistically significantly different from control

Time spent in the target quadrant in the Morris water maze test:

PND40: significant decrease at 2 mg/kg-day only 

PND75: significant decrease at ≥ 0.2 mg/kg-day

Number of platform crossings in the Morris water maze test:

PND40: significant decrease at 2 mg/kg-day only 

PND75: significant decrease at ≥ 0.2 mg/kg-day (in females) and 2 

mg/kg-day (in males)

Bouayed et al. (2009b)

Female Swiss albino mice, 5/group

0, 2, or 20 mg/kg-day maternal 

gavage

PND 0 – 14 (lactational exposure)

Significant increase in the percent of spontaneous alternations in the Ymaze alternation test at 2 mg/kg-day but not at 20 mg/kg-day

No effect on the total number of arm entries in the Y-maze alternation test
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Neuromuscular function and coordination

Chen et al. (2012)

Sprague-Dawley rats, 20 pups (10 

male and 10 female)/group

0, 0.02, 0.2, or 2 mg/kg-day by 

gavage

PND5-PND11

Latency in the surface righting reflex test 

PND12: significant increase at 0.2 mg/kg-day only

PND14: Significant increase at 0.02 and 2 mg/kg-day only 

PND16 Significant difference at 2 mg/kg-day only 

PND 18 No significant difference

Latency in the negative geotaxis test

PND12: significant increase at all doses

PND14: Significant increase at 2 mg/kg-day only

PND16 and 18: No significant difference

No effect on duration of forelimb grip in forelimb grip strength test

Bouayed et al. (2009b)

Female Swiss albino mice, 5/group

0, 2, or 20 mg/kg-day maternal 

gavage

PND 0 – 14 (lactational exposure)

Significant increase in righting time in the surface righting reflex test at 

both doses on PNDs 3 and 5 (but not PNDs 7 and 9)

Significant increase in latency in the negative geotaxis time for 20 mg/kgday dose group at PNDs 5, 7, and 9 (no significant difference at PND 11)

Significant increase in duration of forelimb grip in forelimb grip strength 

test at both dose groups on PND 9 (statistically significant at PND 11 only 

at high dose)

Significant increase in pole grasping latency in male pups in the water 

escape pole climbing test at 20 mg/kg-day

No effect on climbing time in the water escape pole climbing test

Significant increase in pole escape latency in the water escape pole 

climbing test in male rats at 20 mg/kg-day

Anxiety and/or motor activity

Chen et al. (2012)

Sprague-Dawley rats, 20 pups (10 

male and 10 female)/group

0, 0.02, 0.2, or 2 mg/kg-day by 

gavage

PND5-PND11

Significant increase in the number of entries into open arms in the 

elevated plus maze at PND 70 at ≥ 0.2 mg/kg-day (in females) and ≥ 2 

mg/kg-day (in males) (no difference at PND 35)

Significant decrease in the number of entries into closed arms in the 

elevated plus maze at PND 70 at ≥ 0.2 mg/kg-day (in females) and ≥ 2 

mg/kg-day (in males) (no difference at PND 35)

Significant increase in the time spent in open arms in the elevated plus 

maze at PND35 at ≥ 2 mg/kg-day in females and at PND70 at doses ≥ 0.02 

mg/kg-day in females and ≥ 0.2 mg/kg-day in males

Significant decrease in latency time to first entry an open arm in the 

elevated plus maze on PND 70 at ≥ 0.2 mg/kg-day (no difference at PND 

35)

Significant increase in the number of squares crossed in the open-field 

activity test: 

PND34- significant increase at 2 mg/kg-day; PND69- significant increase 

at ≥ 0.02 mg/kg-day (no difference at PND18 and 20) 

Significant increase in rearing activity in open-field activity at 0.2 mg/kgday on PND 69 (no difference at PNDs 18, 20, and 34)

Bouayed et al. (2009b)

Female Swiss albino mice, 5/group

0, 2 or 20 mg/kg-day maternal 

gavage

PND 0 – 14 (lactational exposure)

Significantly increased time in open arms in the elevated plus maze at ≥ 2 

mg/kg-day

Significantly increased percentage of entries into open arms in the 

elevated plus maze at ≥ 2 mg/kg-day

Significantly decreased entries into closed arms in the elevated plus maze 

at 2 mg/kg-day, but not at 20 mg/kg-day

Significantly decreased latency time in the elevated plus maze at 20 
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mg/kg-day

No effect on the total number of arm entries in the elevated plus maze

No effect on the latency to retract from the edge in the cliff aversion test

Electrophysiological changes

McCallister et al. (2008)

Long Evans Hooded rats, 5-

6/group

0 or 0.3 mg/kg-day by gavage

GD14- 17

Statistically significant decreases in stimulus-evoked cortical neuronal 

activity on PND90 – PND120

Reduction in the number of spikes in both the short and long latency 

periods on PND90 – PND120 (no quantitative data presented by authors)

Wormley et al., 2004

F344 rats, 10 females/group

0 or 100 μg/m3nose-only 

inhalation for 4 hrs/day 

GD 11–21

Electrophysiological changes in the hippocampus: 

Consistently lower LTP following gestational exposure (statistical analysis 

not reported) 

Response relative to control: -26%

1 Figure 1-2. Exposure-response array for neurodevelopmental effects 

2 following oral exposure

3
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1 Summary of Developmental Effects

2 Developmental effects following in utero exposure to benzo[a]pyrene have been reported in 

3 human and in animal models. In human populations, decreased head circumference, decreased 

4 birth weight, and decreased postnatal weight have been reported (Tang et al., 2006; Perera et al., 

5 2005a, b). Analogous effects in laboratory animals, including decreased pup weight and decreased 

6 fetal survival, have been noted following gestational or early postnatal exposure to benzo[a]pyrene 

7 by the oral or inhalation route (Chen et al., 2012; Archibong et al., 2002; MacKenzie and Angevine, 

8 1981). Reproductive function is also altered in mice treated gestationally with benzo[a]pyrene 

9 (Kristensen et al., 1995; MacKenzie and Angevine, 1981). These effects include impaired 

10 reproductive performance in F1 offspring (male and female) and alterations of the weight and 

11 histology of reproductive organs (ovaries and testes). 

12 The available human and animal data also support the conclusion that benzo[a]pyrene is 

13 developmental neurotoxicant. Human studies of environmental PAH exposure in two cohorts have 

14 observed neurotoxic effects, including suggestions of reduced head circumference (Tang et al., 

15 2006; Perera et al., 2005a; 2004), impaired cognitive ability (Tang et al., 2008, Perera et al., 2009), 

16 impaired neuromuscular function (Tang et al., 2008), and increased attention problems following 

17 prenatal exposure (Perera et al., 2012). These effects were seen in pregnancy cohort studies in 

18 different populations (New York City and China), in studies using specific benzo[a]pyrene measures 

19 (i.e., adduct levels measured in cord blood samples) (Tang et al., 2008; 2006; Perera et al., 2012; 

20 2005a; 2004). This type of measure covers a relevant time window of exposure with respect to 

21 gestational development. The analytical method was similar in the two studies (with a common set 

22 of investigators). The coefficient of variation of the exposure measures was relatively small (12%), 

23 but a high proportion of samples were below the detection limit and thus these studies were 

24 limited in terms of ability to examine a broad range of exposure. The available evidence from mice 

25 and rats also demonstrates significant and persistent developmental impairments following 

26 exposure to benzo[a]pyrene. Impaired learning and memory behaviors and impaired 

27 neuromuscular function were consistently observed in multiple neurobehavioral tests in two 

28 separate species at comparable doses in the absence of maternal or fetal toxicity (Bouayed et al., 

29 2009b; Chen et al., 2012). 

30 In conclusion, the available human and animal data suggest that developmental toxicity and 

31 developmental neurotoxicity is a hazard of benzo[a]pyrene exposure.

32 Susceptible Populations and Lifestages

33 Childhood susceptibility to benzo[a]pyrene toxicity is indicated by epidemiological studies 

34 reporting associations between adverse birth outcomes and developmental effects and internal 

35 biomarkers of exposure to benzo[a]pyrene, presumably via exposure to complex PAH mixtures 

36 (Tang et al., 2008, 2006; Perera et al., 2005a, b). The occurrence of BPDE-DNA adducts in maternal 

37 and umbilical cord blood in conjunction with exposure to ETS was associated with reduced birth 
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1 weight and head circumference in pregnant women living in the vicinity of fires from the 

2 09/11/2001 disaster site in New York City (Perera et al., 2005a). In other studies, elevated levels of 

3 BPDE-DNA adducts in umbilical cord blood were associated with: (1) reduced birth weights or 

4 reduced head circumference in the offspring of 529 Dominican or African-American nonsmoking 

5 women (Perera et al., 2005b); and (2) decreased body weight at 18, 24, and 30 months and deficits 

6 in several areas of development as assessed by the Gesell Developmental Schedules at 24 months in 

7 the offspring of nonsmoking Chinese women living in the vicinity of a coal-fired power plant (Tang 

8 et al., 2008, 2006).

9 Studies in humans and experimental animals indicate that exposure to PAHs in general, and 

10 benzo[a]pyrene in particular, may impact neurological development. Observational studies in 

11 humans have suggested associations between gestational exposure to PAHs and later measures of 

12 neurodevelopment (Perera et al., 2009; Tang et al., 2008). An observational study of a Chinese 

13 population living in close proximity to a coal fired power plant found increased levels of 

14 benzo[a]pyrene-DNA adducts in cord blood were associated with decreased developmental 

15 quotients in offspring (Tang et al., 2008). 

16 Evidence in animals of the effects of benzo[a]pyrene on neurological development includes: 

17 1) decreased electrophysiological response to electrical stimulation of the dentate gyrus of the 

18 hippocampus and increased brain concentrations of benzo[a]pyrene metabolites in offspring of 

19 F344 rats exposed by inhalation to benzo[a]pyrene:carbon black aerosols on GD 11-21 (Wormley et 

20 al., 2004; Wu et al., 2003); 2) decreased evoked response in the field cortex and decreased 

21 cerebrocortical levels of mRNA for the NMDA receptor subunit in offspring of Long Evans rats 

22 exposed to 300 μg/kg on GD 14-17 (McCallister et al., 2008); and 3) decreased righting reflex and 

23 altered disinhibition behavior in offspring of lactating rats exposed to oral doses of 2 or 20 mg/kg24 day on PND 1-14 (Bouayed et al., 2009b).

25 1.1.2. Reproductive Toxicity

26 Human and animal studies provide evidence for benzo[a]pyrene-induced male and female 

27 reproductive toxicity. Effects on sperm quality and male fertility have been demonstrated in human 

28 populations highly exposed to PAH mixtures (Soares and Melo, 2008; Hsu et al., 2006). The use of 

29 internal biomarkers of exposure in humans (e.g., BPDE-DNA adducts) support associations between 

30 benzo[a]pyrene exposure and these effects. In females, numerous epidemiological studies indicate 

31 cigarette smoking reduces fertility; however, few studies have specifically examined levels of 

32 benzo[a]pyrene exposure and female reproductive outcomes (see Table 1-6). Animal studies 

33 demonstrate decrements in sperm quality, changes in testicular histology, and hormone alterations 

34 following benzo[a]pyrene exposure in adult male animals, and decreased fertility and ovotoxic 

35 effects in adult females following exposure to benzo[a]pyrene.
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1 Male Reproductive Effects

2 Fertility

3 Effects on male fertility have been demonstrated in populations exposed to mixtures of 

4 PAHs. Spermatozoa from smokers have reduced fertilizing capacity, and embryos display lower 

5 implantation rates (Soares and Melo, 2008). Occupational PAH exposure has been associated with 

6 higher levels of PAH-DNA adducts in sperm and male infertility (Gaspari et al., 2003). In addition, 

7 men with higher urinary levels of PAH metabolites have been shown to be more likely to be infertile 

8 (Xia et al., 2009). Studies were not identified which directly examined the reproductive capacity of 

9 adult animals following benzo[a]pyrene exposure. However, a dose-related decrease in fertility 

10 was observed in male mice treated in utero with benzo[a]pyrene, as discussed in Section 1.1.1.

11 Sperm parameters

12 Effects on semen quality have been demonstrated in populations exposed to mixtures of 

13 PAHs including coke oven workers and smokers (Soares and Melo, 2008; Hsu et al., 2006). Coke 

14 oven workers had higher frequency of oligospermia (19 vs. 0% in controls) and twice the number 

15 of morphologically abnormal sperm (Hsu et al., 2006). Elevated levels of BPDE-DNA adducts have 

16 been measured in the sperm of populations exposed to PAHs occupationally (Gaspari et al., 2003)

17 and through cigarette smoke (Phillips, 2002; Zenzes et al., 1999). A higher concentration of BPDE18 DNA adducts was observed in sperm not selected for intrauterine insemination or IVF based on 

19 motility and morphology in patients of fertility clinics (Perrin et al., 2011a, b). An association 

20 between benzo[a]pyrene exposure levels and increased sperm DNA fragmentation using the sperm 

21 chromatin structure assay (SCSA) was observed by Rubes et al. (2012). However, it is currently 

22 unclear whether the SCSA, which measures sperm fragmentation following denaturation, is 

23 predictive of fertility (Sakkas et al., 2009; American Society of Reproductive Medicine 2008). 

24 In several studies in rats and mice, decreased sperm count, motility, and production, and an 

25 increase in morphologically abnormal sperm have been observed (Table 1-5). Alterations in these 

26 sperm parameters have been observed in different strains of rats and mice and across different 

27 study designs and routes of exposure. 

28 Decreases in epididymal sperm (25 to 50% compared to controls) counts have been 

29 observed in Sprague-Dawley rats and C57BL6 mice treated with 1- 5 mg/kg-day benzo[a]pyrene 

30 following oral exposure 42 or 90 days (Chen et al., 2011a; Mohamed et al., 2010). Additionally, a 

31 15% decrease in epididymal sperm count was observed at a dose 100-fold lower in Sprague32 Dawley rats exposed to benzo[a]pyrene for 90 days (Chung et al., 2011). However, confidence in 

33 this study is limited as the authors dosed animals with 0.001, 0.01, and 0.1 mg/kg-day 

34 benzo[a]pyrene, but only reported on sperm parameters at the mid-dose and no other available 

35 studies demonstrated findings in the range of the mid- and high-dose. A short term study in mice 

36 and a subchronic inhalation study in rats lend support for the endpoint of decreased sperm count 

37 (Arafa et al., 2009; Archibong et al., 2008; Ramesh et al., 2008). Significantly decreased sperm 
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1 count and daily sperm production (~40% decrease from control in each parameter) were observed 

2 following 10 days of gavage exposure to 50 mg/kg-day benzo[a]pyrene in mice (Arafa et al., 2009). 

3 In addition, decrements in sperm count were observed following inhalation exposure to 

benzo[a]pyrene in rats for 60 days to 75 μg/m3 4 (Archibong et al., 2008; Ramesh et al., 2008). 

5 Both oral and inhalation exposure to benzo[a]pyrene have been shown to lead to decreased 

6 epididymal sperm motility and altered morphology in rodents. Decreased motility of 20 to 30% 

7 compared to controls was observed in C57BL6 mice (≥ 1mg/kg-day) and Sprague-Dawley rats (at 

8 0.01 mg/kg-day) (Chung et al. 2011; Mohamed et al., 2010). However, the effective doses spanned 

9 two degrees of magnitude and as noted above, confidence in the study observing effects at 0.01 

10 mg/kg-day benzo[a]pyrene (Chung et al., 2011) is limited by poor reporting. A short term oral 

11 study in mice also reported a significantly decreased number of motile sperm (~40% decrease) 

12 following 10 days of gavage exposure to 50 mg/kg-day benzo[a]pyrene in rats (Arafa et al., 2009). 

13 In addition, decreased sperm motility was observed following inhalation exposure to 

benzo[a]pyrene in rats for 60 days to 75 μg/m3 14 (Archibong et al., 2008; Ramesh et al., 2008) and for

10 days at ≥ 75 μg/m3 15 (Inyang et al., 2003). Abnormal sperm morphology was observed in 

16 Sprague-Dawley rats treated with 5 mg/kg-day benzo[a]pyrene by gavage for 84 days (Chen et al., 

17 2011a) and in rats exposed to 75 μg/m3 benzo[a]pyrene by inhalation for 60 days (Archibong et al., 

18 2008; Ramesh et al., 2008).

19 Testicular changes

20 Several studies have demonstrated dose-related effects on male reproductive organs in 

21 adult animals exposed subchronically to benzo[a]pyrene (Table 1-5). Decreases in testicular 

22 weight of approximately 35% have been observed in a 10 day gavage study in adult rats exposed to 

23 50 mg/kg-day benzo(a)pyrene (Arafa et al., 2009) and following subchronic inhalational exposure 

of adult F344 rats to 75 μg/m3 24 (Archibong et al., 2008; Ramesh et al., 2008). No effects on testes 

25 weight were observed in Wistar rats exposed for 35 days to gavage doses up to 36 mg/kg-day 

26 (Kroese et al., 2001); F344 rats exposed for 90 days to dietary doses up to 100 mg/kg-day 

27 (Knuckles et al., 2001); or Sprague-Dawley rats exposed for 90 days to gavage doses up to 0.1 

28 mg/kg-day (Chung et al., 2011). Strain differences may have contributed to the lack of response, 

29 however, F344 rats exposed to benzo[a]pyrene via inhalation showed effects on testicular weight 

30 (Archibong et al., 2008; Ramesh et al., 2008). In addition, decreased testicular weight has also been 

31 observed in offspring following in utero exposure to benzo[a]pyrene as discussed in Section 1.1.1. 

32 Histological changes in the testis have often been reported to accompany decreases in 

33 testicular weight. Apoptosis, as evident by increases in Terminal deoxynucleotidyl transferase 

34 dUTP nick end labeling (TUNEL) positive germ cells and increases in caspase-3 staining, was 

35 evident in seminiferous tubules of Sprague-Dawley rats following 90 days of exposure to ≥0.001

36 and 0.01 mg/kg-day, respectively, benzo[a]pyrene by gavage (Chung et al., 2011). However, the 

37 study authors did not observe testicular atrophy or azospermia in any dose group. Seminiferous 

38 tubules were reported to look qualitatively similar between controls and animals exposed to 
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benzo[a]pyrene by inhalation doses of 75 μg/m3 1 for 60d (Archibong et al., 2008; Ramesh et al., 

2 2008). However, when histologically examined, statistically significantly reduced tubular lumen 

3 size and length were observed in treated animals. Seminiferous tubule diameters also appeared to 

4 be reduced in exposed animals, although this difference did not reach statistical significance 

5 (Archibong et al., 2008; Ramesh et al., 2008). In addition, histological changes in the seminiferous 

6 tubules have also been observed in offspring following in utero exposure to benzo[a]pyrene as 

7 discussed in Section 1.1.1. 

8 Epididymal changes

9 In addition to testicular effects, histological effects in the epididymis have been observed 

10 following 90-day gavage exposure to benzo[a]pyrene (Chung et al., 2011) (Table 1-5). Specifically, 

11 statistically significant decreased epididymal tubule diameter (for caput and cauda) was observed 

12 at doses ≥ 0.001 mg/kg-day. At the highest dose tested (0.1 mg/kg-day) diameters were reduced 

13 approximately 25%. No changes in epididymis weights were observed following an 84 day 

14 treatment in Sprague-Dawley rats of 5 mg/kg-day benzo[a]pyrene (Chen et al., 2011a).

15 Hormone changes

16 Several animal models have reported decreases in testosterone following both oral and 

17 inhalation exposure to benzo[a]pyrene (Table 1-5). In male Sprague-Dawley rats, decreases in 

18 testosterone have been observed following 90-day oral exposures (Chung et al., 2011; Zheng et al., 

19 2010). Statistically significant decreases of 15% in intratesticular testosterone were observed at 5 

20 mg/kg-day in one study (Zheng et al., 2010), while a second study in the same strain of rats 

21 reported statistically significant decreases of 40% in intratesticular testosterone and 70% in serum 

22 testosterone (70%) at 0.1 mg/kg-day (Chung et al., 2011). Statistically significant decreases in 

23 intratesticular testosterone (80%) and serum testosterone (60%) were also observed following

24 inhalation exposure to 75 μg/m3 benzo[a]pyrene in F344 rats for 60 days (Archibong et al., 2008; 

25 Ramesh et al., 2008). Statistically significant increases in serum luteinizing hormone (LH) have 

26 also been observed in Sprague-Dawley rats following gavage exposure to benzo[a]pyrene at doses

27 of ≥ 0.01 mg/kg-day and in F344 rats following inhalation exposure to 75 μg/m3 benzo[a]pyrene 

28 for 60 days (Archibong et al., 2008; Ramesh et al., 2008).

29 Table 1-5. Evidence pertaining to the male reproductive toxicity of 

30 benzo[a]pyrene in animals

Study Design and Reference Results

Sperm quality

Mohamed et al., 2010

C57BL/6 mice, 10 males/dose

(treated before mating with unexposed 

females)

0, 1, or 10 mg/kg-day, daily gavage 

↓ epididymal sperm count in F0 mice

Approximate % change from control: 0, -50*, -70*

(numerical data not reported)

↓ epididymal sperm motility in F0 mice
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Study Design and Reference Results

(F0 males only) 

42 days 

Approximate % change from control: 0, -20*, -50*

(numerical data not reported)

↓ epididymal sperm count in untreated F1 and F2 generations

(numerical data not reported)

No effects were observed in the F3 generation

Chen et al., 2011a

Sprague-Dawley rats , 10 males/dose

0 or 5 mg/kg-day by gavage

84d 

↓ epididymal sperm count(% change from control)

0, - 25%*

↑ % abnormal epididymal sperm

5, 8*

Chung et al., 2011

Sprague-Dawley rats, 20-25 males/dose

0, 0.001, 0.01, 0.1 mg/kg-day by gavage 

90 days

↓ epididymal sperm motility (% response relative to control; 

reported only for 0.01 mg/kg-day)

0, -30* 

No statistically significant decrease in epididymal sperm count

(reported only for 0.01 mg/kg-day)

Archibong et al., 2008; Ramesh et al., 

2008

F344 rats, 10 males/group

0 or 75 μg/m3, 4 hrs/day by inhalation 

60 days

↓ epididymal sperm motility (% change from control)

0, -73*

↓ epididymal sperm count (% change from control)

0, -69*

↑ % abnormal epididymal sperm

33, 87*

↓ spermatids/g testis (approximate % change from control; 

numerical data not reported)

0, -45*

Testicular changes (weight, histology)

Mohamed et al., 2010

C57BL/6 mice, 10 males/dose

(treated before mating with unexposed 

females)

0, 1, or 10 mg/kg-day, daily gavage 

(F0 males only) 

42 days

↓ seminiferous tubules with elongated spermatids (approximate % 

change from control; numerical data not reported))

0, -20*, -35*

No statistically significant change in area of seminiferous epithelium 

of testis (approximate % change from control; numerical data not 

reported)

0, 5, 20

Chung et al., 2011

Sprague-Dawley rats, 20-25 males/dose

0, 0.001, 0.01, 0.1 mg/kg-day by gavage 

90 days

↑ number of apoptotic germ cells per tubule (TUNEL or caspase 3 

positive)

No change in testis weight or histology

Chen et al., 2011a

Sprague-Dawley rats, 10/dose

0 or 5 mg/kg-day by gavage 

84 days

↑ testicular lesions characterized as irregular arrangement of germ 

cells and absence of spermatocytes (numerical data not reported)

No change in testis weight
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Study Design and Reference Results

Archibong et al., 2008; Ramesh et al., 

2008

F344 rats, 10 adult males/group

0 or 75 μg/m3, 4 hrs/day by inhalation

60 days

↓ decreased testis weight (% change from control)

0, 34*

↓ size of seminiferous tubule lumens and reduced tubular length 

No change in % of tubules with elongated spermatids

Kroese et al., 2001

Wistar rats, 10/sex/dose

0, 1.5, 5, 15, or 50 mg/kg-day by gavage 

35 days

No change in testis weight 

Knuckles et al., 2001

F344 rats, 20/sex/dose

0, 5, 50, or 100 mg/kg-day in diet 

90 days

No change in testis weight

Epididymal changes (weight, histology)

Chung et al., 2011

Sprague-Dawley rats, 20-25 males/dose

0, 0.001, 0.01, 0.1 mg/kg-day by gavage 

90 days

↓ diameter of caput epididymal tubule (n=5; numerical data not 

reported)

↓ diameter of cauda epididymal tubule (n=5; numerical data not 

reported)

Chen et al., 2011a

Sprague-Dawley rats, 10/dose

0 or 5 mg/kg-day by gavage 

84 days

No change in epididymis weight

Hormone changes

Chung et al., 2011

Sprague-Dawley rats, 20-25 males/dose

0, 0.001, 0.01, 0.1 mg/kg-day by gavage 

90 days

↓ Intratesticular testosterone (approximate % change from control; 

numerical data not reported)

0, -12, -25, -40*

↓ Serum testosterone (approximate % change from control ; 

numerical data not reported)

0, 0, -35, -70*

↑ serum LH (approximate % change from control; numerical data 

not reported)

0, 33, 67*, 87*

↓ hCG or dbcAMP-stimulated testosterone production in Leydig 

cells 

Zheng et al., 2010

Sprague-Dawley rats, 8 males/dose

0, 1 or 5 mg/kg-day by gavage 

90 days

↓ Intratesticular testosterone (pproximate % change from control ; 

numerical data not reported)

0, -15, -15*

Archibong et al., 2008; Ramesh et al., 

2008

F344 rats, 10 adult males/group

0 or 75 μg/m3, 4 hrs/day by inhalation

60 days

↓ intratesticular testosterone (approximate % change from control;

numerical data not reported)

0, -80*

↓ serum testosterone (approximate % change from control)

0, -60*
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Study Design and Reference Results

↑ serum LH (approximate % change from control)

0, 50*

1 Figure 1-3. Exposure-response array for male reproductive effects following 

2 oral exposure

3

4 Mode of Action Analysis—Male reproductive effects

5 Data regarding the potential mode of action for male reproductive effects associated with 

6 benzo[a]pyrene exposure are limited. Hypothesized modes of action include benzo[a]pyrene 

7 mediated DNA damage to male germ cells leading to genotoxicity, cytotoxicity, and apoptosis 
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1 (Chung et al., 2011; Olsen et al., 2010; Perrin et al., 2011 a,b; Revel et al., 2001), compromised 

2 function of Leydig and Sertoli cells (Chung et al., 2011; Raychoudhury and Kubinski, 2003), altered 

3 androgen hormone regulation (Inyang et al., 2003; Vinggaard et al., 2000), and decreased embryo 

4 viability post fertilization associated with sperm DNA damage (Borini et al., 2006; Seli et al., 2004).

5 Female Reproductive Effects

6 Fertility

7 In women, exposure to cigarette smoke has been shown to affect fertility, including effects 

8 related to pregnancy, ovulatory disorders, and spontaneous abortion (as reviewed in Waylen et al., 

9 2009; Cooper and Moley, 2008; Soares and Melo, 2008). In addition, several studies suggest that in 

10 utero exposure to maternal tobacco smoke also decreases the future fertility of female offspring (Ye 

11 et al., 2010; Jensen et al., 1998; Weinberg et al., 1989). Benzo[a]pyrene levels in follicular fluid and 

12 benzo[a]pyrene-DNA adducts in granulosa-lutein cells and oocytes and in human cervical cells have 

13 been associated with smoking status and with amount smoked (Neal et al., 2008; Zenzes et al., 

14 1998; Mancini et al., 1999; Melikian et al., 1999; Shamsuddin and Gan, 1988). 

15 Few epidemiological studies have examined the specific influence of components of PAH 

16 mixtures on fertility or other reproductive outcomes; EPA identified only two studies with specific 

17 data on benzo[a]pyrene (Table 1-6). One of these studies addressed the probability of conception 

18 among women undergoing in vitro fertilization (Neal et al., 2008). Although no association was 

19 seen with serum levels of benzo[a]pyrene, follicular fluid benzo[a]pyrene levels were significantly 

20 higher among the women who did not conceive with women who did get pregnant. The other study 

21 examined risk of what was termed a “missed abortion” (i.e., fetal death before 14 weeks gestation), 

22 using a case-control design with controls selected from women undergoing elective abortion (Wu et 

23 al., 2010). Benzo[a]pyrene-DNA adduct levels were similar in the aborted tissue of cases compared 

24 with controls, but a strong association was seen between maternal blood benzo[a]pyrene-DNA 

25 adduct levels and risk of missed abortion, with a 4-fold increased risk for levels above compared 

26 with below the median. 

27 Experimental studies in animals also provide evidence that benzo[a]pyrene exposure 

28 affects fertility (Table 1-7). Decreased fertility and fecundity (decreased number of F0 females 

29 producing viable litters at parturition) was statistically significantly reduced by about 35% in adult 

30 females exposed to 160 mg/kg-day of benzo[a]pyrene (MacKenzie and Angevine, 1981). In another 

31 study, F0 females showed no signs of general toxicity or effects on fertility following gavage 

32 exposure to 10 mg/kg-day on GD 7-16 (Kristensen et al., 1995). Decrements in fertility were more 

33 striking in the offspring from these studies, as described in Section 1.1.1.

34 Ovarian effects

35 Human epidemiological studies which directly relate ovotoxicity and benzo[a]pyrene 

36 exposure are not available; however, smoking, especially during the time of the peri-menopausal 
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1 transition, has been shown to accelerate ovarian senescence (Midgette and Baron, 1990). 

2 Benzo[a]pyrene-induced ovarian toxicity has been demonstrated in animal studies. In adult female 

3 rats treated by gavage, statistically significant, dose-related decreases in ovary weight has been 

4 observed in female rats treated for 60 days at doses ≥ 5 mg/kg (2.5 mg/kg-day adjusted) (Xu et al., 

5 2010). At 10 mg/kg in adult rats (5 mg/kg-day adjusted) ovary weight was decreased 15% (Xu et 

6 al., 2010). Changes in ovarian weight were not observed in two subchronic studies in rats. 

7 Specifically, no changes in ovary weight were seen in Wistar rats exposed for 35 days to gavage 

8 doses up to 36 mg/kg-day (Kroese et al., 2001) or in F344 rats exposed for 90 days to dietary doses 

9 up to 100 mg/kg-day (Knuckles et al., 2001). 

10 In adult female rats treated by gavage dose-related decreases in the number of primordial 

11 follicles have been observed in female rats treated for 60 days at doses ≥ 2.5 mg/kg-day, with a 

12 statistically significant decrease of approximately 20% at the high dose (Xu et al., 2010) (Table 1-7). 

13 No notable differences in other follicle populations and corpora lutea were observed. However, in 

14 utero studies exposing dams to the same doses produced offspring with few or no follicles or 

15 corpora lutea (Kristensen et al., 1995; MacKenzenie and Angevine 1981). Additional support for 

16 the alteration of female reproductive endpoints comes from IP experiments in animals and in vitro 

17 experiments. Several studies have observed ovarian effects (decreased numbers of ovarian follicles 

18 and corpora lutea, absence of folliculogenesis, oocyte degeneration, and decreased fertility) in rats 

19 and mice exposed via intraperitoneal (i.p.) injection (Mattison et al., 1980; Swartz and Mattison, 

20 1985; Miller et al., 1992; Borman et al., 2000). Further evidence is available from in vitro studies

21 showing inhibition of antral follicle development and survival, as well as decreased production of 

22 estradiol, was demonstrated in mouse ovarian follicles cultured with benzo[a]pyrene for 13 days 

23 (Sadeu and Foster, 2011). Likewise, FSH-stimulated growth of cultured rat ovarian follicles was 

24 inhibited by exposure to benzo[a]pyrene (Neal et al., 2007).

25 Hormone levels

26 Alteration of hormone levels has been observed in female rats following oral or inhalation 

27 exposure to benzo[a]pyrene (Table 1-7). Inhalation exposure to benzo(a)pyrene:carbon black 

28 particles during gestation resulted in decreases in plasma progesterone, estradiol, and prolactin in 

29 pregnant rats (Archibong et al., 2002). In addition, statistically significant, dose-related decreases 

30 in estradiol along with altered estrus cyclicity was observed in female rats treated for 60 days at 

31 doses ≥ 2.5 mg/kg-day by gavage (Xu et al., 2010). Mechanistic experiments have also noted 

32 decreased estradiol output in murine ovarian follicles cultured with benzo[a]pyrene in vitro for 13 

33 days, but did not find any decrease in progesterone (Sadeu and Foster 2011).

34 Cervical effects

35 One subchronic animal study is available which investigated effects in the cervix following 

36 oral exposure to benzo[a]pyrene (Table 1-7). Statistically-significant dose-related increases in the 

37 incidence of cervical inflammatory cells were observed in mice exposed twice a week for 98 days to 
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1 benzo[a]pyrene via gavage at doses ≥ 2.5 mg/kg (Gao et al., 2011a, 2010). Cervical effects of 

2 increasing severity, including epithelial hyperplasia, atypical hyperplasia, apoptosis, and necrosis, 

3 were observed at higher doses. There are no data on cervical effects in other species or in other 

4 mouse strains. Gao et al. (2011) considered the hyperplasia responses to be preneoplastic lesions. 

5 Cervical neoplasia was not reported in the available chronic bioassays, but this tissue was not 

6 subjected to histopathology examination in either bioassay (Beland and Culp, 1998; Kroese et al., 

7 2001). Thus, the relationship of the cervical lesions to potential development of neoplasia is 

8 uncertain.

9 Table 1-6. Evidence pertaining to the female reproductive effects of 

10 benzo[a]pyrene in humans 

Reference and Study Design:

Study Type/Period/Study Size/Location/Exposure 

Estimate Results

Probability of Conception

Neal et al., 2008

36 women undergoing in vitro fertilization (19 

smokers, 7 passive smokers, and 10 non smokers)

Exposure: B[a]P in serum and follicular fluid

B[a]P levels (ng/ml)a

Conceived

Did not 

Conceive

(p-value)

Follicular fluid 0.1 1.7 (< 0.001)

Serum 0.01 0.05 (not reported)

a

estimated from Figure 3 of Neal et al., 2008

Fetal Death

Wu et al., 2010

Case control study: 81 cases (96% participation 

rate) - fetal death confirmed by ultrasound before 

14 weeks gestation; 81 controls (91% participation 

rate) - elective abortions); matched by age, 

gestational age, and gravidity; excluded smokers 

and occupational PAH exposure

Tianjin, China

Exposure: B[a]P in aborted tissue and maternal 

blood samples (51 cases and controls, 2 of 4 

hospitals)

B[a]P adduct levels (/108nucleotides), mean (±SD)

Cases Controls (p-value)

Maternal blood 6.0 (± 4.7) 2.7 (± 2.2) (< 0.001)

Aborted tissue 4.8 (± 6.0) 6.0 (± 7.4) (0.29)

Low correlation between blood and tissue levels (r = –0.02 in 

cases, r = –0.21 in controls)

Association between B[a]P adducts and missed abortiona

OR (95% CI)

Per unit increase in adducts 1.37 (1.12, 1.67)

Dichotomized at median 4.56 (1.46, 14.3)

a

Conditional logistic regression, adjusted for maternal 

education, household income, and gestational age; age 

also considered as potential confounder
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1 Table 1-7. Evidence pertaining to the female reproductive effects of 

2 benzo[a]pyrene in animals 

Study Design and Reference Results

Fertility

MacKenzie and Angevine, 1981

CD-1 mice, 30 or 60 F0 females/ dose

0, 10, 40, or 160 mg/kg-d by gavage 

GD 7–16

↓ Number of F0 females with viable litters 

46/60, 21/30, 44/60, 13/30*

Kristensen et al., 1995

NMRI mice, 9 females/dose

0 or 10 mg/kg-d by gavage 

GD 7–16 

No changes in fertility of F0 females

Ovarian effects (weight, histology, follicle numbers)

Xu et al., 2010 

Sprague-Dawley rats, 6 females/

dose 

0, 5 or 10 mg/kg by gavage every 

other day (2.5 and 5 mg/kg-day, 

adjusted)

60 days 

↓ Ovary weight (% change from control)

0, 11*, 15*

↓ number of primordial follicles (20%* decrease at high dose)

↑ increased apoptosis of ovarian granulosa cells (approximate % 

apoptosis)

2, 24*, 14*

Knuckles et al., 2001

F344 rats, 20/sex/dose

0, 5, 50, or 100 mg/kg-d in diet 

90 days

No changes in ovary weight

Kroese et al., 2001 

Wistar rats, 10/sex/dose

0, 1.5, 5, 15, or 50 mg/kg-day by 

gavage 5x/week

35 days

No changes in ovary weight 

Hormone levels

Xu et al., 2010 

Sprague-Dawley rats, 6 females/

dose 

0, 5 or 10 mg/kg by gavage every 

other day (2.5 and 5 mg/kg-day, 

adjusted)

60 days 

↓ Serum estradiol (approximate change from control)

0, 16%, 25%

Altered estrous cyclicity

Archibong et al., 2002

F344 rats, 10 females/group

0, 25, 75, or 100 μg/m3 by inhalation

4 hrs/day

GD 11–20 (serum hormones tested 

at GD 15 and 17 in 0, 25, and 75 

μg/m3 dose groups)

↓ F0 estradiol, approximately 50% decrease at 75 μg/m3 at GD 17

↓ F0 prolactin, approximately 70% decrease at 75 μg/m3 at GD 17

↑ F0 plasma progesterone approximately 17% decrease at 75 μg/m3 at 

GD 17

Cervical effects
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Study Design and Reference Results

Gao et al, 2011

ICR mice, 26 females/dose

0, 2.5, 5, or 10 mg/kg by gavage 2 

days/week

98 days 

↑ cervical epithelial hyperplasia: 0/26, 4/26, 6/25*, 7/24*

↑ cervical atypical hyperplasia: 0/26, 0/26, 2/25, 4/24*

↑ inflammatory cells in cervical epithelium: 3/26, 10/26, 12/26*, 18/24* 

↑ mortality: 0/26, 0/26, 1/26, 2/26

1 Figure 1-4. Exposure-response array for female reproductive effects following 

2 oral exposure

3

4 Mode of Action Analysis—Female reproductive effects

5 Although the mechanisms underlying female reproductive effects following benzo[a]pyrene 

6 exposure are not fully established, associations with stimulation of apoptosis, impairment of 
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1 steroidogenesis, and cytotoxicity have been made. Ovarian lesions in benzo[a]pyrene-exposed rats 

2 have been associated with increased apoptosis in ovarian granulosa cells and alteration in 

3 hormone-mediated regulation of folliculogenesis (Xu et al., 2010), and results from in vitro 

4 experiments provide support for an association between benzo[a]pyrene exposure and impaired 

5 folliculogenesis, steroidogenesis, and oocyte maturation (Sadeu and Foster, 2011; Neal et al., 2007). 

6 A growing body of research suggests that benzo(a)pyrene triggers the induction of apoptosis in 

7 oocytes through AhR driven expression of pro-apoptotic genes including Bax (Kee et al., 2010; Neal 

8 et al., 2010; Pru et al., 2009; Matkainen et al., 2002, 2001; Robles et al., 2000). Other proposed 

9 mechanisms include the impairment of folliculogenesis from reactive metabolites (Takizawa et al., 

10 1984; Mattison and Thorgeirsson, 1979; 1977) or by a decreased sensitivity to FSH-stimulated 

11 follicle growth (Neal et al., 2007). Based on findings that an ERα antagonist counteracted effects of 

12 subcutaneously administered benzo[a]pyrene on uterine weight (decreased in neonatal rats and 

13 increased in immature rats), interactions with ERα have been proposed, possibly via occupation of 

14 ERα binding sites or via AhR-ER-crosstalk (Kummer et al., 2008; 2007). However, several in vitro 

15 studies have demonstrated low affinity binding of benzo[a]pyrene to the estrogen receptor and 

16 alteration of estrogen-dependent gene expression (Liu et al., 2006; Van Lipzig et al., 2005; 

17 Vondracek et al., 2002; Fertuck et al., 2001; Charles et al., 2000); so the role of the ER in 

18 benzo[a]pyrene-induced reproductive toxicity is unclear.

19 Summary of Reproductive Effects

20 Male Reproductive Effects

21 Exposure to benzo[a]pyrene in laboratory animals induces male reproductive effects 

22 including decreased levels of testosterone and increased levels of LH, decreased sperm count and 

23 motility, histological changes in the testis, and decreased reproductive success. These findings in 

24 animals are supported by decrements in sperm quality and decreased fertility in human 

25 populations exposed to PAH mixtures (Soares and Melo, 2008; Hsu et al., 2006). In laboratory 

26 animals, male reproductive toxicity has been observed after oral and inhalation exposure to rats or 

27 mice. Effects seen after oral exposures include impaired fertility, effects on sperm parameters, 

28 decreased reproductive organ weight, testicular lesions, and hormone alterations (Chen et al., 

29 2011a; Chung et al., 2011; Mohamed et al., 2010; Zheng et al., 2010; MacKenzie and Angevine, 

30 1981). In addition to oral exposure, male reproductive effects of benzo[a]pyrene have also been 

31 observed following inhalation exposure in rats (Archibong et al., 2008; Ramesh et al., 2008; Inyang 

32 et al., 2003). The male reproductive effects associated with benzo[a]pyrene exposure are 

33 considered to be biologically plausible and adverse. The evidence for male reproductive toxicity

34 seen across multiple human and animal studies identifies the male reproductive system effects as a 

35 potential hazard associated with exposure to benzo[a]pyrene.
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1 Female Reproductive Effects

2 A large body of mechanistic data, both in vivo and in vitro, suggests that benzo[a]pyene 

3 impacts fertility through the disruption of folliculogenensis. This finding is supported, albeit 

4 indirectly, by observations of premature ovarian senescence in women exposed to cigarette smoke 

5 (Midgette and Baron, 1990). Evidence for female reproductive toxicity of benzo(a)pyrene comes 

6 from studies of human populations exposed to PAH mixtures as well as laboratory animal and in 

7 vitro studies. In addition, two human studies observed associations specifically between 

8 benzo(a)pyrene measures and two fertility-related endpoints: decreased ability to conceive (Neal 

9 et al., 2008, 2007) and increased risk of early fetal death (i.e., before 14 weeks gestation) (Wu et al., 

10 2010). Studies in multiple strains of rats and mice indicate fertility related effects including 

11 decreases in ovarian follicle populations and decreased fecundity. Decreased serum estradiol has 

12 also been noted in two different strains of rats exposed by oral or inhalation exposure. The 

13 reproductive effects associated with benzo(a)pyrene exposure are biologically supported and 

14 relevant to humans. In consideration of the evidence from human, animal, and mechanistic studies, 

15 female reproductive effects are identified as a potential hazard associated with exposure to

16 benzo(a)pyrene.

17 Susceptible Populations and Lifestages 

18 Epidemiological studies indicate that exposure to complex mixtures of PAHs, such as 

19 through cigarette smoke, is associated with measures of decreased fertility in humans (Neal et al., 

20 2005; El Nemr et al., 1998) and that prenatal exposure to cigarette smoking is associated with 

21 reduced fertility of women later in life (Weinberg et al., 1989). A case-control study in a Chinese 

22 population has also indicated that women with elevated levels of benzo[a]pyrene-DNA adducts in 

23 maternal blood were 4 times more likely to have experienced a missed abortion (Wu et al., 2010).

24 Inhalation exposure of pregnant female rats to benzo[a]pyrene:carbon black aerosols on GD

25 11- 20 caused decreased fetal survival and number of pups per litter associated with decreased 

26 levels of plasma progesterone, estradiol, and prolactin (Archibong et al., 2002). Decreased numbers 

27 of live pups were also seen in pregnant mice following i.p. exposure to benzo[a]pyrene (Mattison et 

28 al., 1980). These results indicate that benzo[a]pyrene exposure can decrease the ability of females 

29 to maintain pregnancy.

30 Oral multigenerational studies of benzo[a]pyrene exposure in mice demonstrated effects on 

31 fertility and the development of reproductive organs (decreased ovary and testes weight) in both 

32 male and female offspring of pregnant mice exposed to 10 to 160 mg/kg-day on GD 7-16 

33 (Kristensen et al., 1995; Mackenzie and Angevine, 1981).

34 Reductions in female fertility associated with decreased ovary weight following gestational 

35 exposure are supported by observations of: 1) destruction of primordial follicles (Borman et al., 

36 2000; Mattison et al., 1980) and decreased corpora lutea (Miller et al., 1992; Swartz and Mattison, 

37 1985) in adult female mice following i.p. exposure; 2) decreased ovary weight in adult female rats 
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1 following oral exposure (Xu et al. 2010); 3) stimulation of oocyte apoptosis (Matkainen et al., 2002, 

2 2001) or by a decreased sensitivity to FSH-stimulated follicle growth (Neal et al., 2007).

3 Reductions in male fertility associated with decreased testes weight following gestational 

4 exposure are supported by observations of: 1) decreased sperm count, altered serum testosterone 

5 levels, testicular lesions, and/or increased numbers of apoptotic germ cells in adult rats following 

6 repeated oral exposure to benzo[a]pyrene (Chung et al., 2011; Chen et al., 2010; Zheng et al., 2010; 

7 Arafa et al., 2009); 2) decreased epididymal sperm counts in adult F0 and F1 generations of male 

8 mice following 6 weeks oral exposure of the F0 to benzo[a]pyrene (Mohammed et al., 2010); and 3) 

9 decreased testis weight, decreased testicular or plasma testosterone levels, and/or decreased 

10 sperm production, motility and density in adult male rats following repeated inhalation exposure to 

11 aerosols of benzo[a]pyrene:carbon black (Archibong et al., 2008; Ramesh et al., 2008; Inyang et al., 

12 2003).

13 1.1.3. Immunotoxicity

14 Human studies evaluating immune effects following exposure to benzo[a]pyrene alone are 

15 not available for any route of exposure. However, a limited number of occupational human studies, 

16 particularly in coke oven workers, show effects on immune parameters associated with exposure to 

17 PAH mixtures. These studies are of limited utility because effects associated specifically with 

18 benzo[a]pyrene cannot be distinguished from other constituents of the PAH mixture. Subchronic 

19 and short-term animal studies have reported immunotoxic effects of benzo[a]pyrene by multiple 

20 routes of exposure (Table 1-8). Effects include changes in thymus weight and histology, decreased 

21 B cell percentages and other alterations in the spleen, and immune suppression. Data obtained 

22 from subchronic oral gavage studies are supported by short term, intraperitoneal (i.p.), 

23 intratracheal (i.t.), and subcutaneous (s.c.) studies. Additionally, there is evidence in animals for 

24 effects of benzo[a]pyrene on the developing immune system. No studies were located that 

25 examined immune system endpoints following inhalation exposure of animals to benzo[a]pyrene.

26 Thymus Effects

27 Decreased thymus weights (up to 62% compared to controls) were observed in male and 

28 female Wistar rats exposed by gavage to 3-90 mg/kg-day benzo[a]pyrene for 35 or 90 days (Kroese 

29 et al., 2001; De Jong et al., 1999). This effect may be due to thymic atrophy. The incidence of slight 

30 thymic atrophy was increased in males (6/10) and females (3/10) at a dose of 30 mg/kg-day in a 

31 90-day study, although there was no evidence of atrophy at any lower dose (Kroese et al., 2001). 

32 Additionally, at the highest dose tested (90 mg/kg-day) in one of the 35-day studies, the relative 

33 cortex surface area of the thymus and thymic medullar weight were significantly reduced (De Jong 

34 et al., 1999). Other histopathological changes in the thymus (increased incidence of brown 

35 pigmentation of red pulp; hemosiderin) were observed in Wistar rats of both sexes at 50 mg/kg36 day in a 35-day study; however, this tissue was not examined in intermediate-dose groups (Kroese 
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1 et al., 2001). Consistent with the effects observed in these studies, decreased thymus weights and 

2 reduced thymic cellularity were observed in i.p. injection studies that exposed mice to doses 

3 ranging from 50 to 150 mg/kg in utero (Holladay and Smith 1995, 1994; Urso and Johnson, 1988).

4 Spleen Effects

5 Reduced splenic cellularity indicated by decreased relative and absolute number of B cells 

6 in the spleen (decreased 13-41% and 61% compared to controls, respectively) and decreased 

7 absolute number of splenic cells (31% decrease at the highest dose) was observed in a subchronic 

8 study in male Wistar rats administered 3-90 mg/kg-day benzo[a]pyrene by gavage for 35 days (De 

9 Jong et al., 1999). While the effect on the relative number of B cells was dose-related, the lower 

10 doses did not affect the number of B cells or the absolute splenic cell number. The reduced splenic 

11 cell counts were attributed by the study authors to the decreased B cells, and suggest a possible 

12 selective toxicity of benzo[a]pyrene to B cell precursors in the bone marrow. The spleen effects 

13 observed in the single subchronic study are supported by observations of reduced spleen cellularity 

14 and decreased spleen weights following i.p. injection or in utero benzo[a]pyrene exposure to doses 

15 ranging from 50 to 150 mg/kg (Holladay and Smith, 1995; Urso and Johnson, 1988). 

16 In addition to physical effects on the spleen, several studies have demonstrated functional 

17 suppression of the spleen following benzo[a]pyrene exposure. Dose-related decreases in sheep red 

18 blood cell (SRBC) specific serum IgM levels after SRBC challenge were reported in rats (10 or 40 

19 mg/kg-day) and mice (5,20,or 40 mg/kg-day) following s.c. injection of benzo[a]pyrene for 14 days 

20 (Temple et al., 1993). Similarly, reduced spleen cell responses, including decreased numbers of 

21 plaque forming cells and reduced splenic phagocytosis to SRBC and lipopolysaccharide challenge, 

22 were observed in B6C3F1 mice exposed to doses ≥40 mg/kg-day benzo[a]pyrene by i.p. or s.c. 

23 injection for 4-14 days (Lyte and Bick, 1985; Dean et al., 1983; Munson and White, 1983) or by 

24 intratracheal instillation for 7 days (Schnizlein et al., 1987). 

25 Immunoglobulin Alterations

26 Alterations in immunoglobulin levels have been associated with exposure to PAH mixtures 

27 in a limited number of human studies. Some occupational studies have reported evidence of 

28 immunosuppression following PAH exposure. For example, reductions in serum IgM and/or IgA 

29 titers were reported in coke oven workers (Szczeklik et al., 1994; Wu et al., 2003). Conversely, 

30 immunostimulation of immunoglobulin levels has also been observed in humans, specifically 

31 elevated IgG (Karakaya et al., 1999) and elevated IgE (Wu et al., 2003) following occupational PAH 

32 exposure. 

33 Decreases in serum IgM (13 to 33% compared to controls) and IgA levels (22-61% 

34 compared to controls) were observed in male Wistar rats exposed to 3-90 mg/kg-day 

35 benzo[a]pyrene by gavage for 35 days (De Jong et al., 1999); however, these reductions were not 

36 dose-dependent. Similarly, reductions in IgA (9-38% compared to controls) were also observed in 
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1 male and female B6C3F1 mice exposed to doses of 5-40 mg/kg benzo[a]pyrene by s.c. injection for 

2 14 days (Munson and White, 1983). Reductions in serum IgG levels of 18-24%, although not 

3 statistically significant, were observed in female B6C3F1 mice exposed to doses ≥50 mg/kg 

4 benzo[a]pyrene by i.p. injection for 14 days (Dean et al., 1983).

5 Immune Suppression and Sensitization

6 Some occupational studies of coke oven emissions have reported evidence of 

7 immunosuppression following PAH exposure. Reduced mitogenic responses in T cells (Winkler et 

8 al., 1996) and reduced T-lymphocyte proliferative responses (Karakaya et al., 2004) have been 

9 observed following occupational exposure to PAH. Increased levels of apoptosis were observed in 

10 the peripheral blood mononuclear cells (a population of lymphocytes and monocytes) of 

11 occupationally exposed coke oven workers; a response that may contribute to immunodeficiency in 

12 this population (Zhang et al. 2012). However, a limitation of this study is that it does not attribute 

13 the proportion of apoptotic activity to a specific class of cells and does not include assessment of 

14 other potential markers of immunotoxicity in peripheral blood. 

15 Results of functional immune assays in laboratory animals following short term i.p. and s.c. 

16 exposures add to the evidence for benzo[a]pyrene immunotoxicity. Resistance to Streptococcus 

17 pneumonia or Herpes simplex type 2 was dose dependently reduced in B6C3F1 mice following s.c. 

18 injection of ≥5mg/kg-day benzo[a]pyrene for 14 days (Munson et al., 1985). Reduced cell 

19 proliferation, IFN-γ release, and IL-4 release were observed in male and female C56BL/6 mice 

20 following short-term exposure to a gavage dose of 13 mg/kg benzo[a]pyrene as measured in a 

21 modified local lymph node assay (van den Berg et al., 2005). A statistically significant decrease in 

22 NK cell activity was observed in male Wistar rats (Effector:Target cell ratio was 40.9 ± 28.4% that 

23 of controls) exposed to 90 mg/kg-day by gavage for 35 days (De Jong et al., 1999); however, splenic 

24 NK cell activity was not affected in B6C3F1 mice after s.c. injection of 40mg/kg-day benzo[a]pyrene 

25 for 14 days (Munson et al., 1985). The magnitude of the dose and duration of the exposure may 

26 account for the discrepancy between these two studies. Single i.p. injections of 50 mg/kg 

27 benzo[a]pyrene decreased pro and/or pre B-lymphocytes and neutrophils in the bone marrow of 

28 C57BL/6J mice without affecting the numbers of immature and mature B-lymphocytes or GR-1+ 

29 myeloid cells (Galvan et al., 2006).

30 In contrast to studies that have shown immunosuppression, benzo[a]pyrene may also 

31 induce sensitization responses. Epicutaneous application of benzo[a]pyrene (100 μg 

32 benzo[a]pyrene to C3H/HeN mice followed by ear challenge with 20 μg benzo[a]pyrene 5 days 

33 later) produced a contact hypersensitivity (a significant ear swelling) response (Klemme et al., 

34 1987).
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1 Developmental Immunotoxicity

2 As noted above, several i.p. injection studies suggest that cell-mediated and humoral 

3 immunity may be altered by exposure to high doses of benzo[a]pyrene during gestation. 

4 Suppression of the mixed lymphocyte response, the graft-versus-host response, and suppression of 

5 the plaque-forming cell response to SRBCs was observed in mice exposed in utero to 150 mg/kg 

6 during mid (GD 11-13), late (GD 16-18), or both (GD 11-17) stages of gestation; these effects 

7 persisted until 18 months of age (Urso and Gengozian, 1984, 1982, 1980). Fetal thymic atrophy, as 

8 assessed by reductions in cellularity (74-95%, compared to controls), was observed in mice 

9 exposed to 50-150 mg/kg benzo[a]pyrene from GD 13-17, when examined on GD18 (Holladay and 

10 Smith 1994). Analysis of cell surface markers (e.g., CD4, CD8) from the same study indicate that 

11 benzo[a]pyrene may inhibit and/or delay thymocyte maturation, possibly contributing to the 

12 observed thymic atrophy (Holladay and Smith 1994). Consistent with these findings, several other 

13 studies have noted decreased thymocyte numbers and disrupted T cell maturation after in utero 

14 exposure to benzo[a]pyrene (Rodriguez et al., 1999; Holladay and Smith, 1995; Lummus and 

15 Henningsen 1995, Urso et al., 1992; Urso and Johnson, 1987). 

16 The fetal liver is the primary hematopoietic organ during gestation and a major source of 

17 thymocyte precursors beginning around GD 10 or 11 in mice (Landreth and Dodson, 2005; Penit 

18 and Vasseur 1989). Statistically significant reductions in total cellularity in the fetal liver of 54% 

19 and 67% were reported in offspring after gavage exposures of 50 or 100 mg/kg benzo[a]pyrene to 

20 the dams on GD 13-17, respectively (Holladay and Smith, 1994). The decreased fetal liver 

21 cellularity was accompanied by decreased expression of terminal deoxynucleotidyl transferase and 

22 CD45R cellular markers, which are known to be present in cortical thymocyte progenitors in the 

23 fetal liver (Holladay and Smith, 1994; Fine et al., 1990; Silverstone et al., 1976). These data also 

24 suggest that benzo[a]pyrene disrupts liver hematopoiesis during gestation and may interfere with 

25 prolymphoid seeding of the thymus, possibly contributing to thymic atrophy and cell-mediated 

26 immunosuppression. Decreased numbers of CD4+ T-cells have been reported in the spleen of 1-

27 week-old mice following in utero benzo[a]pyrene exposure by i.p. injection to the dams, 

28 demonstrating the potential for downstream effects on T cell development (Rodriguez et al., 1999). 

29 The decreased numbers of CD4+ T-cells correspond with observations of decreased proliferation in 

30 the presence of ConA and a weak response compared to controls in an allogeneic mixed lymphocyte 

31 reaction assay (Urso et al., 2008).

32 Table 1-8. Evidence pertaining to the immune effects of benzo[a]pyrene in 

33 animals

Study Design and Reference Results

Thymus Effects

Kroese et al., 2001

Wistar rats, 10/sex/dose

0, 3, 10, 30 mg/kg-d by gavage 5 days/week

↓ thymus weight 

Females (% change from controls): 0, -3, -6, -28*

Males (% change from controls): 0, 0, -13, -29*
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90 days

↑ slight thymic atrophy

Females (incidence): 0/10, 0/10, 0/10, 3/10

Males (incidence): 0/10, 2/10, 1/10, 6/10*

De Jong et al., 1999

Wistar rats, 8 males/dose

0, 3, 10, 30, 90 mg/kg-d by gavage 5 days/week

35 days

↓ thymus weight

% change from control: 0, -9, -15*, -25*, -62*

Kroese et al., 2001

Wistar rats, 10/sex/dose 

0, 1.5, 5, 15, 50 mg/kg-d by gavage 5 days/week

35 days

↓ thymus weight

Females (% change from controls): 0, +13, +8, -3, -17*

Males (% change from controls): 0, -8, -11, -27*, -33*

Spleen Effects

De Jong et al., 1999

Wistar rats, 8 males/dose 

0, 3, 10, 30, 90 mg/kg-d by gavage 5 days/week 

35 days

↓ in relative number (%) of B cells in spleen

% change from control: 0, -8, -13*, -18*, -41*

↓ in total number of B cells in spleen

% change from control: 0, +13, -13, -13, -61*

Change in total cell number in the spleen

% change from control: 0, +20, 0, +7, -31*

Immunoglobulin Alterations

De Jong et al., 1999

Wistar rats, 8 males/dose

0, 3, 10, 30, 90 mg/kg-d by gavage 5 days/week

35 days

↓ Serum IgM

% change from control: 0, -13, -14, -33*, -19

↓ Serum IgA

% change from control: 0, -27, -22, -28, -61*

*

1 Statistically significant (p<0.05) as reported by study authors.
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1 Figure 1-5. Exposure-response array for immune effects following oral 

2 exposure

3

4 Mode of Action Analysis—Immune Effects

5 Exposure to benzo[a]pyrene induces immunosuppressive effects such as decreased 

6 numbers of B cells in the spleen and decreased thymus weight and cellularity following oral, i.p., 

7 s.c., or i.t. exposure in experimental animals. However, the key events underlying benzo[a]pyrene 

8 immunotoxicity have not been identified. 

9 Benzo[a]pyrene is a well known ligand for the aryl hydrocarbon receptor (AhR) (Okey et al., 

10 1994; Nebert et al., 1993; Postlind et al., 1993). Ligands of the AhR have been shown to have a role 

11 in regulating hematopoietic stem cells (HSC) in the bone marrow, a major site of B cell proliferation 

12 and antibody production (Esser et al 2009). Benzo[a]pyrene induces greater reductions in bone 

13 marrow cultures from Ah-responsive (C57BL/6) mice compared to Ah-nonresponsive (DBA/2) 

14 mice (Hardin et al. 1992). Furthermore, B cell lymphopoiesis in Ah-responsive mice was dose 
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1 dependently suppressed but not in Ah-nonresponsive mice (Hardin et al., 1992). Addition of the 

2 AhR antagonist and cytochrome P450 (CYP450) inhibitor, α-NF, prevented the benzo[a]pyrene3 induced inhibition of B cell lymphopoiesis in a concentration-dependent fashion. Similarly, the

4 CYP1A1 inhibitor, 1-(1-propynyl) pyrene, blocked benzo[a]pyrene-induced B-cell growth 

5 inhibition, but not through the metabolite, BPDE (Allen et al., 2006). Altogether, these data suggest 

6 that benzo[a]pyrene may regulate B cell proliferation and antibody production in the bone marrow 

7 via the AhR.

8 Summary of Immune Effects

9 Immunotoxic effects of benzo[a]pyrene exposure are based on data from animal studies 

10 that vary in route and duration of exposure. There are no human epidemiological studies that 

11 provide specific support for benzo[a]pyrene immunotoxicity; however, immunosuppression has 

12 been observed in studies following occupational exposure to PAH mixtures. However, these 

13 findings are limited by co-exposures to other constituents of PAH mixtures. 

14 Each of the immune effects reported in animal bioassays for benzo[a]pyrene provide 

15 equivocal evidence of immunotoxicity; functional assays provide greater support for 

16 immunotoxicity than observational findings such as organ weight, hematological, or 

17 histopathological measures (WHO, 2012). Although the overall database for benzo[a]pyrene 

18 immunotoxicity in experimental animals is limited and the key events underlying the mode of 

19 action are not established, there is evidence of physical alterations to tissues/organs of the immune 

20 system, as well as decreases in immune function. Evidence of benzo[a]pyrene-associated 

21 immunotoxicity is supported by consistent thymic effects observed in two oral studies, as well as 

22 splenic effects, and varying immunosuppressive responses observed in short-term or in vitro tests. 

23 Overall, the weight of evidence in animals indicates that immunotoxicity may be a potential hazard 

24 associated with benzo[a]pyrene exposure, but the immune system does not appear to be a sensitive 

25 target for benzo[a]pyrene-induced toxicity.

26 Susceptible Populations and Lifestages 

27 The severity and persistence of immune effects observed during in utero studies suggests 

28 that immunotoxicity may be greater during gestation than adulthood (Dietert and Pieperbrink, 

29 2006; Holladay and Smialowicz, 2000). Urso and Gengozian (1982) provide experimental support 

30 demonstrating immunosuppression from benzo[a]pyrene exposure during gestation was greater 

31 than for mice exposed after birth to a 25-fold higher dose. There is also substantial literature 

32 indicating that disruption of the immune system during certain critical periods of development 

33 (e.g., initiation of hematopoiesis; migration of stem cells; expansion of progenitor cells) may have 

34 significant and lasting impacts on lifetime immune function (e.g., Burns-Naas et al., 2008; Dietert, 

35 2008; Landreth et al., 2002; Dietert et al., 2000), as well as more specific studies showing increased 
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1 dose sensitivity and disease persistence from developmental versus adult chemical exposure 

2 (reviewed in Luebke et al., 2006).

3 Thymus toxicity is a sensitive and specific effect of benzo[a]pyrene and has been observed 

4 in both prenatal and adult exposure studies. The thymus serves as a major site of thymocyte 

5 proliferation and selection for maturation, and impairment can lead to cell-mediated immune 

6 suppression (Kuper, 2002, 1992; De Waal et al., 1997). The thymus is believed to be critical for T 

7 lymphocyte production during early life and not in adulthood (Hakim et al., 2005; Schonland et al., 

8 2003; Petrie et al., 2002; Mackall et al., 1995). Therefore, the decreases in thymus weight observed 

9 in studies of adult animals exposed to benzo[a]pyrene suggest that immunosuppression may be a 

10 heightened concern for individuals developmentally exposed to benzo[a]pyrene.

11 1.1.4. Other Toxicological Effects

12 There is some evidence that benzo[a]pyrene can produce systemic effects in several organ 

13 systems including the forestomach, liver, kidney, and cardiovascular system, as well as alter 

14 hematological parameters. However, there is less evidence for these effects compared to organ 

15 systems described earlier in Section 1.1.

16 Forestomach effects

17 Lesions have been observed in the forestomach following subchronic and chronic oral 

18 exposure to benzo[a]pyrene (Table 1-9). Increases in the incidence of forestomach hyperplasia 

19 have been observed in Wistar rats following shorter-term, subchronic, and chronic gavage exposure 

20 (Kroese et al., 200; De Jong et al., 1999) and in B6C3F1 mice following chronic dietary exposure 

21 (Beland and Culp, 1998; Culp et al., 1998). 

22 Following chronic gavage exposure, increased incidences of forestomach hyperplasia were 

23 observed in male and female rats at 3 and 10 mg/kg-day; at the highest dose, limited hyperplasia 

24 was reported (Kroese et al., 2001). Only the highest level lesion (hyperplasia, papilloma, or 

25 carcinoma) observed in each organ was scored, such that hyperplasia observed in the forestomach 

26 in which tumors were also observed was not scored. The majority of animals in the high dose 

27 group exhibited forestomach tumors; therefore, the hyperplasia was not scored and the incidence 

28 of forestomach hyperplasia in the study is largely uncharacterized at the highest dose. Shorter term 

29 studies (Kroese et al., 2001; De Jong et al. 1999) showed dose-related increases in forestomach 

30 hyperplasia at doses ≥10 mg/kg-day in Wistar rats. In addition, following chronic dietary exposure, 

31 a dose-dependent increase in the incidence of forestomach hyperplasia and hyperkeratosis was 

32 observed in female mice at ≥0.7 mg/kg-day (Beland and Culp, 1998; Culp et al., 1998). Forestomach 

33 tumors were also observed at ≥0.7 mg/kg-day by Beland and Culp (1998) and Culp et al. (1998).
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1 Table 1-9. Evidence pertaining to the forestomach effects of benzo[a]pyrene in 

2 animals

Study design and reference Results

Kroese et al. (2001)

Wistar (Riv:TOX) rats: male and female

(52/sex/dose group)

0, 3, 10, or 30 mg/kg-day by gavage 5 days/week

104 weeks (chronic)

Wistar (Riv:TOX) rats: male and female

(10/sex/dose group)

0, 3, 10, or 30 mg/kg-day by gavage 5 days/week

90 days (subchronic)

Wistar (specific pathogen-free Riv:TOX) rats

(10/sex/dose group)

0, 1.5, 5, 15, or 50 mg/kg bw by gavage 

5 days/week

5 weeks (shorter-term)

Chronic:

forestomach hyperplasia – (basal cell hyperplasia)

Incidences:

M: 2/50; 8/52; 8/52; 0/52

F: 1/52; 8/51; 13/51; 2/52

Subchronic:

forestomach hyperplasia – (slight basal cell hyperplasia)

Incidences:

M: 2/10; 0/10; 6/10; 7/10

F: 0/10; 2/10; 3/10; 7/10

Short term:

forestomach hyperplasia – (basal cell hyperplasia)

Incidences:

M: 1/10; 1/10; 4/10; 3/10; 7/10

F: 0/10; 1/10; 1/10; 3/10; 7/10*

Beland and Culp (1998); Culp et al. (1998)

B6C3F1 mice: female (48/dose group)

0, 5, 25, or 100 ppm in the diet (average daily 

dosesa: 0, 0.7, 3.3, and 16.5 mg/kg-day)

2 years 

forestomach hyperplasia

Incidences: 13/48; 23/47; 33/46*; 38/47*

De Jong et al. (1999)

Wistar rats: male (8/ dose group)

0, 3, 10, 30, or 90 mg/kg-day by gavage 5 

days/week

5 weeks

forestomach hyperplasia – (basal cell hyperplasia) 

statistically significantly increased incidences at 30 

and 90 mg/kg-day were reported, but incidence data 

were not provided

* indicates statistical significance as identified in study

a

Based on the assumption that daily benzo[a]pyrene intake at 5 ppm was one-fifth of the 25-ppm intake (about 

21 μg/day) and using time-weighted average (TWA) body weights of 0.032 kg for the control, 5- and 25-ppm 

groups and 0.026 kg for the 100-ppm group
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1 Figure 1-6. Exposure-response array for forestomach hyperplasia following 

2 oral exposure

3

4 Mode of Action Analysis—Forestomach Effects

5 Mechanistic investigations suggest that bioactivation of benzo[a]pyrene leads to reactive 

6 intermediates that can lead to mutagenic events, as well as to cytotoxic and apoptotic events. The 

7 available human, animal, and in vitro evidence best supports a mutagenic mode of action as the 

8 primary mode by which benzo[a]pyrene induces carcinogenesis. Available data indicate that 

9 forestomach hyperplasia may be a histological precursor to neoplasia observed at this site after 

10 chronic exposure to benzo[a]pyrene (Kroese et al., 2001; DeJong et al., 1999). Dose-response data 

11 show that forestomach hyperplasia occurs at shorter durations and at lower doses than tumors in 

12 rats and mice exposed to benzo[a]pyene for up to 2 years (Kroese et al., 2001; Beland and Culp, 

13 1998). Kroese et al. (2001) reported that the forestomach lesions demonstrated a progression over 

14 the course of intercurrent sacrifices; the authors described early lesions as focal or confluent basal 

15 hyperplasia, followed by more advanced hyperplasia with squamous cell papilloma, and 

16 culminating in squamous cell carcinoma. The description of the progression of forestomach lesions 
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1 provided by Kroese et al. (2001), coupled with the observation that hyperplasia occurs before 

2 tumors and at lower doses than tumors suggests that forestomach hyperplasia induced by 

3 benzo[a]pyrene is likely a preneoplastic lesion.

4 Summary of Forestomach Effects

5 A dose-dependent increased incidence of forestomach hyperplasia was observed in both 

6 rats and mice and in multiple studies. Evidence for forestomach hyperplasia after both gavage and 

7 dietary administration demonstrates concordance of the effect across different modes of 

8 administration and illustrates that the hyperplasia is not dependent on bolus dosing or mechanical 

9 irritation associated with gavage administration. The increase in forestomach hyperplasia was also 

10 consistent across multiple exposure regimens, including chronic, subchronic, short-term, and acute 

11 exposures.

12 Hematological Effects

13 Altered hematological parameters, including decreases in RBC count, hemoglobin, and 

14 hematocrit have been observed in laboratory animals following benzo[a]pyrene exposure (Table 1-

15 10). Statistically significant decreases in RBC count, hemoglobin, and hematocrit were observed in 

16 male Wistar rats at doses greater than 10 mg/kg-day for 35 days (De Jong et al. 1999). A minimal, 

17 but statistically significant increase in mean cell volume and a decrease in mean cell hemoglobin 

18 were observed at the highest dose (90 mg/kg-day), which may indicate dose-related toxicity for the 

19 RBC’s and/or RBC precursors in the bone marrow (DeJong et al., 1999). Similarly, male and female 

20 F344 rats also showed statistically significant decreases in RBC counts and hematocrit level, along 

21 with decreased hemoglobin levels in a 90 day dietary study (Knuckles et al. 2001). Statistically 

22 significant decreases were observed in the two highest doses (50 and 100 mg/kg-day) in male rats, 

23 while a maximum decrease of 12% was only observed at the highest dose in female rats at 

24 100mg/kg-day. Small, but not statistically significant decreases in RBC counts and hemoglobin 

25 were observed in both a 35-day and 90-day study in Wistar rats (Kroese et al. 2001). It should be 

26 noted that when observed, the magnitudes of the decreases in RBC, hemoglobin, and hematocrit 

27 were generally small; about 18% at 90 mg/kg-day and <10% at lower doses (De Jong et al., 1999) 

28 and about 10% in F344 rats (Knuckles et al., 2001). A decrease in WBC’s, attributed to reduced 

29 numbers of lymphocytes and eosinophils, was also observed at 90 mg/kg-day following gavage 

30 exposure for 35 days (DeJong et al., 1999). The mode of action by which benzo[a]pyrene exposure 

31 may lead to altered hematological parameters is undetermined.

32 Liver Effects

33 Liver effects other than cancer associated with benzo[a]pyrene exposure primarily include 

34 changes in liver weight and abnormal histopathology (Table 1-10). Increased liver weight was 

35 reported in a 90-day study in both male and female Wistar rats given benzo[a]pyrene by gavage 

36 (Kroese et al., 2001). Both females (17% increase) and males (29% increase) demonstrated 
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1 statistically significant increased liver weights at the highest dose tested (30 mg/kg-day); a 

2 statistically significant increase (15%) was also reported in males at 10 mg/kg-day. Similar to the 

3 findings in the 90-day study by Kroese et al. (2001), increased liver:body weight ratios were 

4 observed at the highest dose in a 90-day dietary study in male F344 rats, although there was no 

5 change observed in female liver weights (Knuckles et al. 2001). Increased liver:body weight ratios 

6 were also observed in both sexes at high doses (600 and 1000 mg/kg) in an accompanying acute 

7 study (Knuckles et al. 2001). A statistically significant increase in liver weight was also observed in 

8 male Wistar rats given 90 mg/kg-day benzo[a]pyrene by gavage for 35 days (De Jong et al., 1999). 

9 Consistent with the findings by De Jong et al. (1999), a statistically significant increased liver weight 

10 (about 18%) was also observed in both male and female Wistar rats at the highest dose (50 mg/kg11 day) given by gavage in a 35-day study (Kroese et al., 2001). 

12 Limited exposure-related differences in clinical chemistry parameters associated with liver 

13 toxicity were observed; no differences in alanine aminotransferase (ALT) or serum aspartate 

14 transaminase (AST) levels were observed, and a small dose-related decrease in γ-glutamyl 

15 transferase (GGT) was observed in males only exposed to benzo[a]pyrene for 90 days (Kroese et al., 

16 2001). 

17 Treatment-related lesions in the liver (oval cell hyperplasia) were reported following 

18 exposure to 90 mg/kg-day benzo[a]pyrene for 35 days, however incidence data were not reported 

19 (De Jong et al., 1999). Other histopathological changes observed in the liver include an increased 

20 incidence of liver clear cell foci of alteration in males and females during a 2-year carcinogenicity 

21 study (Kroese et al., 2001), however organs with tumors were not evaluated. Since many of the 

22 animals in the top two doses developed liver tumors, the dose responsiveness of the increased 

23 incidence of clear cell foci is unclear. 

24 A dose-dependent increase in liver microsomal EROD activity, indicative of CYP1A1 

25 induction, was observed in both sexes at doses ≥1.5 mg/kg-day (Kroese et al. 2001). However, at 

26 the highest dose tested, with the greatest fold induction in EROD activity, there was no evidence of 

27 associated adverse histopathologic findings. The finding of increased liver weight across multiple 

28 studies of varying exposure durations, as well as histopathological changes in the liver provide 

29 evidence of the liver as a target of benzo[a]pyrene-induced toxicity. The mode of action by which 

30 benzo[a]pyrene induces these effects is unknown.

31 Kidney Effects

32 There is minimal evidence of kidney toxicity following exposure to benzo[a]pyrene (Table 

33 1-10). A statistically significant decrease in kidney weight was observed in the highest dose tested 

34 (90 mg/kg-day) in a 35-day gavage study in male Wistar rats (De Jong et al. 1999). Decreases in 

35 kidney weight at 3 and 30 mg/kg-day were also observed, but these changes were not dose36 dependent. Additionally, kidney weights were not affected by following exposure to 

37 benzo[a]pyrene for 35 days (Kroese et al. 2001). Histopathological analysis of kidney lesions 
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1 revealed an apparent dose responsive increase in the incidence of abnormal tubular casts in the 

2 kidney in male F344 rats exposed by diet for 90 days (Knuckles et al 2001). The casts were 

3 described as molds of distal nephrons lumen and were considered by the study authors to be 

4 indicative of renal dysfunction. However, the statistical significance of the kidney lesions is 

5 unclear. Several gaps and inconsistencies in the reporting make interpretation of the kidney effects 

6 difficult, including 1) no reporting of numerical data, 2) no indication of statistical significance in 

7 the accompanying figure for kidney lesions, 3) discrepancies between the apparent incidences and 

8 sample sizes per dose group, and 4) uncertainty in how statistical analysis of histopathological data 

9 was applied. As such, the significance of the abnormal tubular casts is unclear. While there are 

10 some findings to suggest that the kidneys may be affected by benzo[a]pyrene exposure, overall 

11 there is insufficient data to suggest that the kidneys may be a primary target of benzo[a]pyrene12 induced toxicity.

13 Cardiovascular Effects

14 Atherosclerotic vascular disease and increased risk of cardiovascular mortality has been 

15 associated with cigarette smoking (Ramos and Moorthy, 2005; Miller and Ramos, 2001; Thirman et 

16 al., 1994) and, to a more limited degree, occupational exposure to PAH mixtures (Friesen et al., 

17 2010, 2009; Burstyn et al., 2005; Chau et al., 1993). Elevated mortality due to cardiovascular 

18 disease was observed in a PAH-exposed occupational population (coke oven plant workers), but 

19 elevated cardiovascular mortality was also observed in the non-exposed or slightly exposed 

20 populations as well (Chau et al., 1993). Elevated risks of ischemic heart disease (IHD) were 

21 associated with past cumulative benzo[a]pyrene exposure (with a 5-year lag), although the trend 

22 was not statistically significant; there was no observed association with more recent 

23 benzo[a]pyrene exposure (Friesen et al., 2010). Elevated risk of mortality from IHD was also 

24 associated with cumulative benzo[a]pyrene exposure in a cohort of male asphalt workers (although 

25 not statistically significant); the trend in average benzo[a]pyrene exposure and association with 

26 IHD was statistically significant, with an approximately 60% increase in risk between the lowest 

27 and highest exposure groups (Burstyn et al., 2005). The two studies which associate 

28 benzo[a]pyrene exposure with cardiovascular effects (Friesen et al., 2010; Burstyn et al., 2005) rely 

29 on statistical models to create exposure groups based on previously gathered benzo[a]pyrene air 

30 samples that may or may not have included members of the cohort under examination. 

31 Additionally, while these studies used benzo[a]pyrene exposure groupings for analysis, they cannot 

32 address co-exposures that may have occurred in the occupational setting (asphalt or aluminum 

33 smelters) or exposures that occurred outside the workplace. 

34 Increased systolic and diastolic blood pressure has been observed in the offspring of dams 

35 exposed to increasing concentrations of benzo[a]pyrene (Jules et al., 2012) (Table 1-10). At the 

36 highest dose tested (1200 μg/kg BW by gavage to the dams), systolic pressures were elevated 

37 approximately 50% and diastolic pressures were elevated approximately 80% above controls. 
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1 Reduced endothelial integrity and increased smooth muscle cell mass, both related to 

2 atherosclerosis, have been observed in Sprague-Dawley rats exposed to 10mg/kg benzo[a]pyrene 

3 by i.p. injection (once/week for 8 weeks) (Zhang and Ramos, 1997). The molecular mechanisms 

4 underlying PAH-induced vascular injury, and the development of atherosclerosis are not well 

5 established, but current hypotheses include cell proliferative responses to injury of endothelial cells 

6 from reactive metabolites (including ROS) and genomic alterations in smooth muscle cells from 

7 reactive metabolites leading to transformed vasculature cells and eventual plaque formation 

8 (Ramos and Moorthy, 2005). However, while the link between PAHs and atherosclerotic disease 

9 have been studied, experiments specifically looking at the relationship between levels of exposure 

10 to benzo[a]pyrene (via environmentally relevant routes), and the development of aortic wall 

11 lesions related to atherosclerosis have not generally been performed. 

12 One exception to this observation comes from a series of experiments on Apolipoprotein E 

13 knock-out (ApoE -/-) mice exposed orally to benzo[a]pyrene (Knaapen et al., 2007; Curfs et al., 

14 2005, 2004; Godschalk et al., 2003). ApoE -/- mice develop spontaneous atherosclerosis, which is 

15 thought to be due to enhanced oxidative stress from the lack of ApoE (Godschalk et al., 2003). 

16 Overall, these studies suggest that benzo[a]pyrene exposure in ApoE-/- mice enhances the 

17 progression of atherosclerosis through a general local inflammatory process. 

18 Neurological Effects

19 Impaired learning and memory, as well as neurochemical alterations, have been observed in 

20 humans following occupational exposure to PAH mixtures (Niu et al, 2009). Male coke oven 

21 workers were analyzed for alterations in neurobehavioral function using the World Health 

22 Organization Neurobehavioral Core Test Battery (WHO-NCTB), as well as changes in 

23 neurotransmitter concentrations in blood. Urinary levels of the PAH metabolite 1-hydroxypyrene 

24 were used as markers of PAH exposure. In the WHO-NCTB, coke workers had lower scores in the 

25 digit span and forward digit span tests than matched control subjects, suggesting that short-term 

26 memory was impaired. The authors also reported that the digit span and forward digit span scores 

27 cores significantly decreased with increasing 1-hydroxypyrene levels in urine. PAH exposure also 

28 altered the blood levels of several neurotransmitters. As in the functional assays, the authors 

29 reported that alterations in neurochemical measures were associated with urinary levels of 1-

30 hydroxypyrene. 

31 Reductions in neuromuscular, autonomic, sensorimotor, and electrophysiological endpoints 

32 have been reported in rats and mice following acute or short-term exposure to benzo[a]pyrene 

33 (Saunders et al., 2006, 2002, 2001; Liu et al., 2002; Grova et al., 2007, 2008; Bouayed et al., 2009a).

34 Impaired Morris water maze performance was observed following subchronic oral gavage in adult 

35 rats (Chen et al. 2011b); however, the study was conducted with one dose group (Table 1-10). 

36 These data suggest that longer durations of benzo[a]pyrene exposure could be neurotoxic; 

37 however, only limited data are available to inform the neurotoxic potential of repeated sub-chronic 
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1 or chronic exposure to benzo[a]pyrene. The available data from the human and chronic and 

2 subchronic animal studies (Kroese et al., 2001; Beland and Culp 1998) provide limited support for 

3 benzo[a]pyrene as a neurotoxicant in adults.

4 Table 1-10. Evidence pertaining to other systemic effects of benzo[a]pyrene in 

5 animals 

Study Design and Reference Results

Hematological Parameters

Kroese et al., 2001

Wistar rats, 10/sex/dose

0, 3, 10, 30 mg/kg-d by gavage 5 

days/week

90 days

RBC count and hemoglobin changes not statistically significant in males or 

females at any dose (numerical data not reported)

Knuckles et al., 2001

F344 rats, 20/sex/dose

0, 5, 50, 100 mg/kg-d by diet

90 days

↓ RBC count

Females (% change from controls): statistically significant at 100mg/kg-d 

(numerical data not reported)

Males (% change from controls): statistically significant at 50 and 

100mg/kg-d (numerical data not reported)

↓ hematocrit 

Females (% change from controls): statistically significant at 100mg/kg-d 

(numerical data not reported) 

Males (% change from controls): statistically significant at 50 and 

100mg/kg-d (numerical data not reported)

↓ hemoglobin

Females: statistically significant at 100mg/kg-d (numerical data 

not reported)Males: statistically significant at 100mg/kg-d 

(numerical data not reported)

De Jong et al., 1999

Wistar rats, 8 males/dose 

0, 3, 10, 30, 90 mg/kg-d by gavage 5 

days/week

35 days

↓ RBC count

% change from controls: 0, -1,- 5*, -10*, -18*

↓ hemoglobin

% change from controls: 0, -1, -7*, -10*, -18*

↓ hematocrit

% change from controls: 0, 0, -6*, -8*, -14*

↓ WBC count

% change from controls: 0, -8, -9, -9, -43* 

↑ mean cell volume

% change from controls: 0, 0, -3, 0, +3* 

↓ mean corpuscular hemoglobin concentration

% change from controls: 0, -1, -1, -1, -3* 
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Study Design and Reference Results

Kroese et al., 2001

Wistar rats, 10/sex/dose 

0, 1.5, 5, 15, 50 mg/kg-d by gavage 5 

days/week

35 days

RBC count: not statistically significant (numerical data not reported)

hemoglobin: not statistically significant (numerical data not reported)

Liver Effects

Kroese et al., 2001

Wistar rats, 10/sex/dose

0, 3, 10, 30 mg/kg-d by gavage 5 

days/week

90 days

↑ liver weight

Females (% change from controls): 0, -2, +4, +17*

Males (% change from controls): 0, +7, +15*, +29*

Liver histopathology: no effects reported

Knuckles et al., 2001

F344 rats, 20/sex/dose

0, 5, 50, 100 mg/kg-d by diet

90 days

↑ liver:body weight ratio

Females: no change (numerical data not reported) 

Males (% change from controls): 23% change reported at 100mg/kg-d 

(numerical data not reported)

De Jong et al., 1999

Wistar rats, 8 males/dose 

0, 3, 10, 30, 90 mg/kg-d by gavage 5 

days/week

35 days

↑ liver weight

% change from controls: 0, -9, +7, +5, +15*

↑ liver oval cell hyperplasia (numerical data not reported)

Kroese et al., 2001

0, 1.5, 5, 15, 50 mg/kg-d by gavage 

5d/wk for 35d

Wistar rats, 10/sex/dose

↑ liver weight

Females (% change from controls): 0, +3, +2, +9, +18*

Males (% change from controls): 0, +2, +1, +3, +18*

Liver histopathology: no effects reported

Kidney Effects

Knuckles et al., 2001

F344 rats, 20/sex/dose

0, 5, 50, 100 mg/kg-d by diet

90 days

↑ abnormal tubular casts

Females: not statistically significant (numerical data not reported)

Males: apparent dose-dependent increase (numerical data not 

reported)

De Jong et al., 1999

Wistar rats, 8 males/dose

0, 3, 10, 30, 90 mg/kg-d by gavage 5 

days/week

35 days

↓ kidney weight

% change from controls: 0, -11*, -4, -10*, -18*

Kroese et al., 2001

Wistar rats, 10/sex/dose 

0, 1.5, 5, 15, 50 mg/kg-d by gavage 5 

days/week

35 days

kidney weight: no change (data not reported)

Neurological Effects
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Study Design and Reference Results

Chen et al. (2011b)

Sprague-Dawley rats, male, 32/dose 

0 or 2 mg/kg-day by gavage

90 days

↑ time required for treated rats to locate platform in water maze (data 

reported graphically)

Bouayed et al. (2009a)

Male Swiss albino mice, 9/group

0, 0.02, 0.2 mg/kg by gavage

28 days

Significant decrease in latency to attack and increase in the number of 

attacks in the resident-intruder test at 0.02 mg/kg-day (but not at high 

dose)

Significant increase in mount number in the copulatory behavior test at 

0.02 mg/kg-day (but not at high dose)

1 *Statistically significant.

2 1.1.5. Carcinogenicity

3 Evidence in Humans

4 There are many epidemiologic studies involving exposure to PAH mixtures that contain 

5 benzo[a]pyrene (e.g., studies of coke oven workers, asphalt workers). This discussion primarily 

focuses on epidemiologic studies that included a direct measure of benzo[a]pyrene exposure.

4 6 The 

7 identified studies were separated into tiers according to extent and quality of the exposure analysis 

8 and other study design features:

9 Tier 1: Detailed exposure assessment conducted, large sample size (greater than ~50 

10 exposed cases), and adequate follow-up period to account for expected latency (e.g., 

11 greater than 20 years for lung cancer).

12 Tier 2: More limited exposure assessment, or sample size or follow-up period did not meet 

13 the criteria for Tier 1, or only a single-estimate exposure analysis was conducted.

14 For lung cancer, each of the Tier 1 studies observed increasing risks of lung cancer with 

increasing cumulative exposure to benzo[a]pyrene (measured in μg/m3 15 -years), and each of these 

16 studies addressed in the analysis the potential for confounding by smoking (Armstrong et al., 2009; 

17 Spinelli et al., 2006; Xu et al., 1996) (Table 1-12). These three studies represent different 

18 geographic locations and two different industries. The pattern of results in the Tier 2 studies was 

19 mixed, as would be expected for studies with less precise exposure assessments or smaller sample 

20 sizes: one of the SMR estimates was < 1.0, with the other 8 estimates ranging from 1.2 to 2.9 (Table 

21 1-13). In considering all of the available studies, particularly those with the strongest methodology, 

22 there is considerable support for an association between benzo[a]pyrene exposure and lung cancer, 

23 although the relative contribution benzo[a]pyrene and of other PAHs cannot be established. 

24 For bladder cancer, the cohort and nested case-control studies observed a much smaller 

25 number of cases compared with lung cancer; this limits their ability to examine exposure-response 

 

4

These studies were identified from PubMed searches (last search conducted May 15, 2012 using terms benzo(a)pyrene 

AND (cohort OR case-control) AND cancer) and two recent meta-analyses (Armstrong et al, 2004; Bosetti et al, 2007).
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1 relationships. Two cohort studies with detailed exposure data, however, had sample sizes of 48 

2 and 78 cases (Burstyn et al., 2009; Gibbs et al. 2007a,b,c) (Tier 1 studies, Table 1-14). Although 

cumulative exposure (up to approximately 2 μg/m3 3 -years) was not related to increasing risk in the 

4 study of asphalt workers by Bursyn et al. (2010), an exposure-response was seen with the wider 

exposure range (i.e., above 100 μg/m3 5 ) examined in a study in an aluminum smelter workers by 

6 Gibbs et al (2007a,b,c). This difference in response is not surprising, given that the highest 

7 exposure group in the asphalt worker study corresponded to the exposures seen in the lowest 

8 exposure category in the study of aluminum smelter workers. The 6 studies with more limited 

9 exposure information or analyses each included between 2 and 16 bladder cancer cases, with 

10 relative risk estimates ranging from 0.6 to 2.9. None of these individual effect estimates was 

11 statistically significant (Table 1-14). 

12 Two of the identified studies contained information on risk of mortality from melanoma. 

13 Neither of these studies observed increased risks of this type of cancer, with an SMR of 0.91 (95% 

14 CI 0.26, 2.48) [22 cases) in Spinelli et al., 2006 and 0.58 (95% CI 0.12, 1.7) in Gibbs et a. (2007a) [3 

15 cases]. 

16 Of additional interest is non-melanoma skin cancer, particularly with respect to dermal 

17 exposures. The literature pertaining to this kind of cancer and PAH exposure goes back to the 18th

18 century work of Sir Percival Pott describing scrotal cancer, a squamous cell skin cancer, in chimney 

19 sweeps (Brown and Thornton, 1957). One of the identified studies reported an increased risk of 

20 mortality from non-melanoma skin cancer among asphalt workers (roofers), with an SMR of 4.0 

21 (95% CI: 1.0, 10.9) among workers with 20 or more years. In addition to this study, 3 studies in 

22 Scandinavian countries examined non-melanoma skin cancer risk in relation to occupations with 

23 likely dermal exposure to creosote (i.e., timber workers, brickmakers, power linesmen) using 

24 incidence data from population registries (Tornqvist et al., 1986; Karlehagen et al., 1992; Pukkala et 

25 al., 1995). The SIR estimates were 1.5 (95% CI: 0.7, 2.6) based on 5 exposed cases, 2.37 (95% CI: 

26 1.08, 4.50) [9 cases], and 4.64 (95% CI: 1.51, 10.8) [5 cases], respectively, in Tornqvist et al. (1986), 

27 Karlehagen et al. (1992), and Pukkala et al. (1995). These studies provide additional support for 

28 the association between dermal PAH exposure, including benzo[a]pyrene exposure, and skin 

29 cancer. 

30 Lung, bladder, and skin cancers are the cancers that have been observed in occupational 

31 studies of PAH mixtures (Benbrahim-Tallaa et al., 2012; IARC, 2010a,b, Secretan et al., 2009; Baan 

32 et al., 2009) (see Table 1-11). The reproducibility of these three sites in different populations and 

33 exposure settings adds plausibility to the hypothesis that common etiologic factors may be 

34 operating. The potential role that benzo[a]pyrene may play as a causal agent is further supported 

35 by the observation that these same three sites are also increased in the studies that included a 

36 direct measure of benzo[a]pyrene.
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1 Table 1-11. Cancer sites for PAH-related agents reviewed by IARC

PAH-related mixture or 

occupation

Sites with sufficient 

evidence in humans

Sites with limited 

evidence in humans Reference

Aluminum production Lung

Urinary bladder

Baan et al (2009)

Carbon electrode manufacture Lung IARC (2010)

Coal gasification Lung Baan et al (2009)

Coal tar distillation Skin Baan et al (2009)

Coal tar pitch (paving and roofing) Lung Urinary bladder Baan et al (2009)

Coke production Lung Baan et al (2009)

Creosotes Skin IARC (2010)

Diesel exhaust Lung Urinary bladder Benbrahim-Tallaa et 

al. (2012)

Indoor emissions from household 

combustion of biomass fuel 

(primarily wood)

Lung Secretan et al (2009)

Indoor emissions from household 

combustion of coal

Lung Secretan et al (2009)

Mineral oils, untreated or mildly 

treated

Skin Baan et al (2009)

Shale oils Skin Baan et al (2009)

Soot (chimney sweeping) Lung

Skin

Urinary bladder Baan et al (2009)

Adapted from IARC, 2010.

Table 1-12. Summary of epidemiologic studies of benzo[a]pyrene (direct 

measures) in relation to lung cancer risk: Tier 1 studies

Reference, design Results

Armstrong et al., 2009 (Quebec, Canada)

Cohort, aluminum smelter workers, 7 plants

16,431 (15,703 men; 728 women)

Duration: minimum 1 year, began work 1966–1990

Follow-up: through 1999 (mean ~ 30 years)

Smoking information collected from medical records

Exposure: Job exposure matrix ~5,000 personal B[a]P 

measures from the 1970s to 1999

Related references: Lavoué et al., 2007 (exposure data); 

Gibbs et al., 2007a,b,c; Armstrong et al., 1994 

SMR 1.32 (1.22, 1.42) [677 cases]

Lung cancer risk by cumulative B[a]P exposure

Median 

B[a]P

μg/m3-years

n 

cases SMR (95% CI) RR (95% CI)

0 35 0.62 (0.44, 0.87) 1.0 (referent)

10 266 1.09 (0.96, 1.23) 1.75 (1.23, 2.48)

30 70 1.88 (1.47, 2.38) 3.02 (2.01, 4.52)

60 53 1.21 (0.91, 1.59) 1.94 (1.27, 2.97)

120 114 1.93 (1.59, 2.32) 3.09 (2.12, 4.51)

240 116 1.79 (1.48, 2.15) 2.86 (1.96, 4.18)

480 23 2.36 (1.49, 3.54) 3.77 (2.23, 6.38)

No evidence of confounding by smoking

Additional modeling as continuous variable: RR 1.35 (95% CI 1.22, 

1.51) at 100 μg/m3-years (0.0035 per μg/m3-years increase); other 

shapes of exposure-response curve examined.
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Adapted from IARC, 2010.

Table 1-12. Summary of epidemiologic studies of benzo[a]pyrene (direct 

measures) in relation to lung cancer risk: Tier 1 studies

Reference, design Results

Spinelli et al., 2006 (British Columbia, Canada)

Cohort, aluminum smelter workers

6,423 (all men)

Duration: minimum ≥ 3 years; began work 1954–1997

Follow-up: through 1999 (14% loss to follow-up; mean ~ 

24years)

Smoking information from self-administered questionnaire

Exposure: Job exposure matrix using 1,275 personal B[a]P 

measures from 1977 to 2000 (69% for compliance 

monitoring)

Related references: Friesen et al., 2006 (exposure data); 

Spinelli et al., 1991

SMR 1.07 (0.89, 1.28) [120 cases]

SIR 1.10 (0.93, 1.30) [147 cases]

Lung cancer risk by cumulative B[a]P exposure

B[a]P

μg/m3-years

n 

cases RR (95% CI)a

0 to 0.5 25 1.0 (referent)

0.5 to 20 42 1.23 (0.74, 2.03)

20 to 40 23 1.35 (0.76, 2.40)

40 to 80 25 1.36 (0.78, 2.39)

≥ 80 32 1.79 (1.04, 3.01)

a

Adjusting for smoking category; trend p < 0.001

Xu et al.,1996 (China)

Nested case-control in iron – steel worker cohort

610 incident cases (96% participation); 959 controls (94% 

participation) (all men)

Duration: data not reported

Smoking information collected from interviews; next-of-kin 

interviews with 30% of lung cancer cases and 5% of controls 

Exposure: Job exposure matrix 82,867 historical monitoring 

records, 1956 to 1992 (not sure what proportion were 

B[a]P)

Lung cancer risk by cumulative B[a]P exposure

B[a]P

(μg/m3-years)

n 

cases RR (95% CI)a

< 0.84 72 1.1 (0.8, 1.7)

0.85 to 1.96 117 1.6 (1.2, 2.3)

1.97 to 3.2 96 1.6 (1.1, 2.3)

≥ 3.2 b105 1.8 (1.2, 2.5)

a

Adjusting for birth year and smoking category; trend p < 

0.004. Referent group is “nonexposed” (employed in 

administrative or low-exposure occupations)

b Table IV of Xu et al. (1996) report unclear; could be >= 3.0 for 

this category

1

2
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Table 1-13. Summary of epidemiologic studies of benzo[a]pyrene (direct 

measures) in relation to lung cancer risk: Tier 2 studies

Reference, design Results

Limited Follow-up Period (≤ 20 years)

Friesen et al., 2009 (Australia)

Cohort, aluminum smelter workers

4,316 (all men)

Duration: minimum 90 days; began work after 1962

Follow-up: through 2002, mean 16 years (maximum 20)

Smoking information from company records if employed 

before 1995 and study interviews if employed after 1994 

Exposure: Job / task exposure matrix using TWA B[a]P 

measures (n=655), 1977–2004 (79% from 1990 – 2004)

RR 1.2 (0.7, 2.3) [19 cases in exposed; 20 in unexposed]

Lung cancer risk by cumulative B[a]P exposure

B[a]P

μg/m3-years

n 

cases RR (95% CI)a

0 20 1.0 (referent)

> 0 to 0.41 6 0.7 (0.3, 1.8)

0.41 to 10.9 6 1.4 (0.6, 3.5)

> 10.9 7 1.7 (0.7, 4.2)

a

Poisson regression, adjusting for smoking; trend p = 0.22

Proxy Measure

Olsson et al., 2010 (Denmark, Norway, Finland, Israel)

Nested case-control, asphalt workers

433 lung cancer cases (65% participation); 1,253 controls (58% 

participation), matched by year of birth, country (all men)

Duration: minimum ≥ 2 seasons, median 8 seasons; began 

work 1913 – 1999

Follow-up: 1980 to 2002 – 2005 (varied by country)

Smoking information from interviews

Exposure: compilation of coal tar exposure measures, 

production characteristics, and repeat measures in asphalt 

industry in each country used to develop exposure matrix

Related references: Burstyn et al., 2000; Boffetta et al., 2003a

Lung cancer risk by cumulative coal tar exposure a

Coal tar

unit-years

a

n 

cases RR (95% CI)

0.39 – 4.29 43 1.31 (0.87, 2.0)

4.30 – 9.42 32 0.98 (0.62, 1.6)

9.43 – 16.88 30 0.97 (0.61, 1.6)

16.89 – 196.48 54 1.60 (1.09, 2.4)

a Adjusting for set, age, country, tobacco pack-years

b

trend p = 0.07

Costantino et al., 1995 (United States – Pennsylvania)

Cohort, coke oven workers 

5,321 and 10,497 unexposed controls (non-oven steel workers; 

matched by age, race, date of first employment) (all men)

Duration: data not reported; worked in 1953

Follow-up through 1982 (length data not reported)

Exposure: average daily exposure coal tar pitch volatiles: 3.15 

mg/m3top-side full-time jobs, 0.88 mg/m3side jobs; used to 

calculate weighted cumulative exposure index

Related reference: Mazumdar et al., 1991 (exposure data)

SMR 1.95 (1.59, 2.33) [255 cases]

Lung cancer risk by cumulative exposure

Coal tar pitch volatiles

(mg/m3-months)

n

cases RR (95% CI) a

0 203 1.0 (referent)

1 – 49 34 1.2 (0.85, 1.8)

50 – 199 43 1.6 (1.1, 2.3)

200 – 349 59 2.0 (1.5, 2.8)

350 – 499 39 2.0 (1.6, 3.2)

500 – 649 27 2.7 (2.0, 4.6)

≥ 650 56 3.1 (2.4, 4.6)

a

adjusted for age, race, coke plant, period of follow-up; 

trend p < 0.001 

Limited Exposure Information

Liu et al., 1997 (China)

Cohort, various carbon plants and aluminum smelter workers

6,635 (all men)

Duration: minimum 15 years; began work before 1971

Follow-up: through 1985 (mean ~ 14 years)

Smoking information from questionnaire

SMR 2.2 (1.1, 2.8) [50 cases]

Lung cancer risk by exposure category

Exposure 

Category

Mean B[a]P

μg/m3

n 

cases SMR (95% CI)a

None -- 13 1.49 (0.83, 2.5)

Low -- 6 1.19 (0.48, 2.5)
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Exposure: Area samples from one carbon plant, 1986 - 1987 Moderate 0.30 5 1.52 (0.55, 3.4)

High 1.19 26 4.30 (2.9, 6.2)

a

Calculated by EPA from data in paper

Berger et al., 1992 (Germany)

Cohort, coke oven workers

789 (all men)

Duration: minimum 10 years (mean 27 years); began work 

1900 to 1989

Follow-up through 1989 (length data not reported)

Smoking information from plant records and interviews 

Exposure: mean B[a]P: 28 (range 0.9 – 89) μg/m3

Related reference: Manz et al., 1983 (exposure information)

SMR 2.88 (2.28–3.59) [78 cases]

Hansen et al., 1991; 1989 (Denmark)

Cohort, asphalt workers

679 workers (applicators) (all men)

Duration: data not reported; employed 1959 to 1980

Follow-up to 1986 (mean ~ 11 years) 

Smoking information from 1982 surveys of industry and 

general population

Exposure: asphalt fume condensate, 35 personal samples 

during flooring: median 19.7 (range 0.5 – 260) mg/m3

SMR 2.90 (1.88, 4.3) [25 cases] (ages 40 to 89)

SMR 2.46 (1.59, 3.6) [25 cases] (with smoking adjustment)

Gustavsson et al., 1990 (Sweden)

Cohort, gas production (coke oven) workers

295 (all men)

Duration: minimum 1 year, median 15 years; employed 1965 

to 1972 Follow-up 1966 to 1986 (mortality); 1966 to 1983 

(incidence; mean ~ 15 years)

Smoking information from interviews with older workers 

Exposure: area sampling - top of ovens. B[a]P - 1964 mean 4.3 

(range 0.007 to 33); 1965 mean 0.52, (0.021 to 1.29) μg/m3

SMR 0.82 (0.22, 2.1) [4 cases] (referent group = employed men)

SIR 1.35 (0.36, 3.5)[4 cases] 

Moulin et al., 1989 (France)

Cohort and nested case-control, two carbon electrode plants 

1,302 in Plant A (all men), employed in 1975; follow-up 1975–

1985 (incidence); smoking information from plant records

1,115 in Plant B (all men); employed in 1957; follow-up 1957–

1984 (mortality)

Duration of employment and follow-up: data not reported 

Exposure: B[a]P - 19 area samples and 16 personal samples in 

Plant A (personal sample mean 2.7; range 0.59 – 6.2 μg/m3); 

10 area samples and 7 personal samples in Plant B; personal 

sample mean 0.17, range 0.02 – 0.57 μg/m3

Plant A: SMR 0.79 (0.32, 1.6) [7 cases]

Plant B: SMR 1.18 (0.63, 2.0) [13 cases]

Internal Comparison (case-control), ≥ 1 year duration:

Plant A: OR 3.42 (0.35, 33.7) [7cases, 21 controls]

Plant B: OR 0.49 (0.12. 2.0) [13 cases, 33 controls]

Hammond et al., 1976 (United States)

Cohort, asphalt – roofers

5,939 (all men)

Duration: minimum 9 years, began before 1960

Follow-up: through 1971 

Exposure: 52 personal samples (masks with filters) during 

specific jobs and tasks. Mean B[a]P 16.7 μg per 7-hour day

SMR 1.6 (1.3, 1.9)a[99 cases] (≥ 20 years since joining union)

a

confidence intervals calculated by EPA from data in paper
Toxicological Review of Benzo[a]pyrene

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Table 1-14. Summary of epidemiologic studies of benzo[a]pyrene (direct 

measures) in relation to bladder cancer risk 

Reference, design Results

Tier 1 Studies

Burstyn et al. (2010) (Denmark, Norway, Finland, Israel)

Cohort, asphalt workers

7,298 all men)

Duration: minimum ≥ 2 seasons, median 8 seasons; began 

work 1913 – 1999

Follow-up: began around 1960, ended around 2000 (years 

varied by country); median 21 years

Smoking information not collected

Exposure: compilation of B[a]P measures, production 

characteristics, and repeat measures in asphalt industry in 

each country used to develop exposure matrix

Related references: Burstyn et al., 2000; Boffetta et al., 

2003a

48 incident bladder cancer cases (39 cases in analyses with 15 

year lag)

Bladder cancer risk by cumulative B[a]P exposure a

B[a]P

μg/m3-years

a

n 

cases

RR (95% CI)

(no lag)b

RR (95% CI)

(15 year lag)c

0 to 0.253 12 1.0 (referent) 1.0 (referent)

0.253 to 0.895 12 0.69 (0.29, 1.6) 1.1 (0.44, 2.9)

0.895 to 1.665 12 1.21 (0.45, 3.3) 1.7 (0.62, 4.5)

≥ 1.665 12 0.84 (0.24, 2.9) 1.1 (0.30, 4.0)

a Adjusting for age, calendar period, total duration of 

employment, country; btrend p = 0.9; c trend p = 0.63

Stronger pattern seen with average exposure in 15-year lag (RR 

1.5, 2.7, 1.9 in second through fourth quartile; trend p = 0.15)

Gibbs et al. (2007a,b,c) (Quebec, Canada)

Cohort, aluminum smelter workers, 7 plants

16,431 (15,703 men; 728 women)

Duration: minimum 1 year, began work 1966 – 1990 

Follow-up: through 1999 (mean ~ 30 years)

Smoking information collected from medical records

Exposure: Job exposure matrix using ~5,000 personal B[a]P 

measures from the 1970s to 1999

Related references: Lavoué et al., 2007 (exposure data); 

Armstrong et al., 1994 ; Gibbs et al. 1979; 1985

Hired before 1950: SMR 2.24 (1.77, 2.79) [78 cases]

Bladder cancer risk by cumulative B[a]P exposure 

B[a]P

μg/m3-years

a

n 

cases SMR (95% CI)

Smoking-adjusted

RRb

0 3 0.73 (0.15, 2.1) 1.0 (referent)

10 14 0.93 (0.45, 1.4) 1.11

30 3 1.37 (0.28, 4.0) 1.97

60 1 0.35 (0.9, 1.9) 0.49

120 15 4.2 (2.4, 6.9) 8.49

240 30 6.4 (4.3, 9.2)

480 12 23.9 (12.2, 41.7)

a Category midpoint

b Confidence intervals not reported; highest category is ≥ 80 

μg/m3-years (n observed = 57).

Mortality risk reduced in cohort hired in 1950 – 1959, SMR=1.23.

Similar patterns seen in analysis of bladder cancer incidence.

Tier 2 Studies 

Friesen et al., 2009 (Australia)

See Table 1-Y for study details

RR 0.6 (0.2, 2.0) [5 cases in exposed; 8 in unexposed]

Bladder cancer risk by cumulative B[a]P exposure

B[a]P

μg/m3-years

n 

cases RR (95% CI)a

0 8 1.0 (referent)

> 0 to 0.41 1 0.2 (0.03, 1.9)

0.41 to 10.9 2 0.7 (0.2, 3.7)

> 10.9 2 1.2 (0.2, 5.6)

a

Poisson regression, adjusting for smoking category; trend p = 

0.22

Spinelli et al., 2006 (British Columbia, Canada)

See Table 1-Y for study details

SMR 1.39 (0.72, 2.43) [12 cases]

Costantino et al., 1995 (United States – Pennsylvania)

See Table 1-Y for study details

SMR 1.14 (1.59, 2.33) [16 cases]
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Hammond et al., 1976 (United States)

See Table 1-Y for study details

SMR 1.7 (0.94, 2.8)a[13cases] (≥ 20 years since joining union)

a

confidence intervals calculated by EPA from data in paper

Moulin et al., 1989 (France)

See Table 1-Y for study details

Plant A: [0 observed cases; expected < 1.0]

Plant B: SMR 1.94 (0.40, 5.0) [3 cases]

Gustavsson et al., 1990 (Sweden)

See Table 1-Y for study details

SMR 2.85 (0.30, 10.3) [2 cases] (referent group = employed men)

1 Evidence in Animals

2 Oral Exposure

3 Evidence of tumorigenicity following oral exposure to benzo[a]pyrene has been 

4 demonstrated in rats and mice. Oral exposure to benzo[a]pyrene has resulted in an increased 

5 incidence of tumors in the alimentary tract in male and female rats (Kroese et al., 2001; Brune et al., 

6 1981) and female mice (Beland and Culp, 1998; Culp et al., 1998), and liver carcinomas in male and 

7 female rats, and kidney adenomas in male rats (Kroese et al., 2001).

8 Forestomach tumors have been observed in several lifetime cancer bioassays in rats and 

9 mice following both gavage and dietary exposure to benzo[a]pyrene at doses ranging from 0.016 

10 mg/kg-day in Sprague-Dawley rats to 3.3 and 10 mg/kg-day in B6C3F1 mice and Wistar rats, 

11 respectively (Kroese et al., 2001; Beland and Culp, 1998; Culp et al., 1998; Brune et al., 1981). In 

12 addition, multiple less-than-lifetime oral exposure cancer bioassays in mice provide supporting 

13 evidence that oral exposure to benzo[a]pyrene is associated with an increased incidence of 

14 forestomach tumors (Weyand et al., 1995; Benjamin et al., 1988; Robinson et al., 1987; El Bayoumy, 

15 1985; Triolo et al., 1977; Wattenberg, 1974; Field and Roe, 1970; Roe et al., 1970; Biancifiori et al., 

16 1967; Chouroulinkov et al., 1967; Fedorenko and Yansheva, 1967; Neal and Rigdon, 1967; 

17 Berenblum and Haran, 1955). 

18 Elsewhere in the alimentary tract, dose-related increases of benign and malignant tumors 

19 were observed. In rats, oral cavity tumors were induced in both sexes and adenocarcinomas of the 

20 jejunum were induced in males (Kroese et al., 2001). In mice, tumors were induced in the tongue, 

21 esophagus, and larynx of females (males were not tested) (Beland and Culp, 1998; Culp et al., 1998).

22 Chronic oral exposure to benzo[a]pyrene resulted in a dose-dependent increased incidence 

23 of liver carcinomas in both sexes of Wistar rats (Table 1-15), with the first liver tumors detected in 

24 week 35 in high-dose male rats; liver tumors were described as complex, with a considerable 

25 proportion (59/150 tumors) metastasizing to the lungs (Kroese et al., 2001). Hepatocellular 

26 tumors were not observed in mice (Beland and Culp, 1998; Culp et al., 1998).

27 A statistically significantly increased incidence of kidney tumors (cortical adenomas) was 

28 observed in male Wistar rats following chronic gavage exposure (Kroese et al., 2001) (Table 1-15). 

29 The kidney tumors were observed at the mid- and high-dose groups. Kidney tumors were not 

30 observed at a lower dose in another study (Brune et al., 1981).

31 Lung tumors were also observed following oral exposure in female AJ mice (Weyand et al., 

32 1995).
Toxicological Review of Benzo[a]pyrene

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1 Table 1-15. Tumors observed in chronic oral animal bioassays.

Study design and reference Results

Kroese et al. (2001)

Wistar (Riv:TOX) rats (52/sex/dose 

group)

0, 3, 10, or 30 mg/kg-day by gavage 

5 days/week

104 weeks 

forestomach –

Incidences:

M: 0/52; 7/52*; 18/52*; 17/52* (papilloma)

M: 0/52; 1/52; 25/52*; 35/52* (squamous cell carcinoma)

F: 1/52; 3/51; 20/51*; 25/52* (papilloma)

F: 0/52; 3/51; 10/51*; 25/52* (squamous cell carcinoma)

oral cavity – 

Incidences:

M: 0/24; 0/24; 2/37; 10/38 (papilloma)

M: 1/24; 0/24; 5/37; 11/38 (squamous cell carcinoma)

F: 0/19; 0/21; 0/9; 9/31(papilloma)

F: 1/19; 0/21; 0/9; 9/31 (squamous cell carcinoma)

jejunum – (adenocarcinomas) 

Incidences:

M: 0/51; 0/50; 1/51; 8/49

F: 0/50; 0/48; 0/50; 2/51

duodenum – (adenocarcinomas)

Incidences:

M: 0/51; 0/50; 0/51; 1/49

F: 0/49; 0/48; 0/50; 2/51

liver – (adenomas and carcinomas) 

Incidences:

M: 0/52; 3/52; 15/52*; 4/52 (adenoma)

M: 0/52; 1/52; 23/52*; 45/52* (carcinoma)

F: 0/52; 2/52; 7/52*; 1/52 (adenoma)

F: 0/52; 0/52; 32/52*; 50/52* (carcinoma)

kidney – (cortical adenoma)

Incidences: 

M: 0/52; 0/52; 7/52; 8/52

F: increase not observed

auditory canalb(Zymbal’s gland) – (carcinomas)

Incidences: 

M: 0/1; 0/0; 2/7; 19/33*

F: 0/0; 0/1; 0/0; 13/20*

Beland and Culp (1998); Culp et al. 

(1998)

B6C3F1 mice: female (48/dose group)

0, 5, 25, or 100 ppm (average daily 

dosesa: 0, 0.7, 3.3, and 16.5 mg/kg-day) 

in the diet

2 years

forestomach – (papillomas and squamous cell carcinomas)

Incidences:

1/48; 3/47; 36/46*; 46/47*

esophagus – (papillomas and carcinomas)

Incidences: 0/48; 0/48; 2/45; 27/46*

tongue – (papillomas and carcinomas)

Incidences: 0/49; 0/48; 2/46; 23/46*

larynx – (papillomas and carcinomas)

Incidences: 0/35; 0/35; 3/34; 5/38
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Study design and reference Results

Brune et al. (1981)

Sprague-Dawley rats: male and female

(32/sex/dose)

Gavage: 0, 6, 18, 39 mg/kg-yr (0, 0.016, 

0.049, 0.107 mg/kg-day)

Diet: 0, 6, 39 mg/kg-yr (0, 0.016, 0.107 

mg/kg-day)

treated until moribund or dead

forestomach – (papillomas and carcinomasb) – gavage

Incidences:

3/64; 12/64*; 26/64*; 14/64*

forestomach – (papillomas) – diet

Incidences:

2/64; 1/64; 9/64*

larynx and esophagus – (papillomas) – gavage

Incidences: 3/64; 1/64; 0/64; 0/64

larynx and esophagus – (papillomas) – diet

Incidences: 1/64; 2/64; 1/64

1 * indicates statistical significance as identified in study

2 a Based on the assumption that daily benzo[a]pyrene intake at 5 ppm was one-fifth of the 25-ppm intake 

3 (about 21 μg/day) and using time-weighted average (TWA) body weights of 0.032 kg for the control, 5- and 

4 25-ppm groups and 0.026 kg for the 100-ppm group

5 b Auditory canal tissue was not histologically examined in the lower dose groups and the controls 

6 c Two malignant forestomach tumors were observed (one each in the mid- and high-dose groups)

7 Inhalation Exposure

8 Chronic inhalation exposure to benzo[a]pyrene resulted in the development of tumors in 

9 the respiratory tract and pharynx in Syrian golden hamsters (Table 1-16). A dose-dependent 

10 increased incidence of tumors in the upper respiratory tract, including the larynx and trachea, were 

11 reported by Thyssen et al. (1981) at ≥10 mg/kg-day. In addition, a decrease in tumor latency was 

12 observed in the larynx and trachea, and nasal cavity tumors were observed at the mid- and high13 dose but the incidences were not dose-dependently increased. A dose-related increase in tumors in 

14 the upper digestive tract (pharynx and esophagus) was also reported. In addition, a single 

15 forestomach tumor was observed at both the mid- and high-doses, and forestomach tumors were 

16 not observed in control animals. The study authors presumed that the pharyngeal and esophageal 

17 tumors were a consequence of mucociliary particle clearance.

18 Under contract to the U.S. EPA, Clement International Corporation (1990) obtained the 

19 individual animal data (including individual animal pathology reports, time-to-death data, and 

20 exposure chamber monitoring data) collected by Thyssen et al. (1981). A re-analysis of the 

21 individual animal pathology reports from the original study supports the dose-dependent increased 

22 incidence of tumors in the larynx and pharynx (Clement International Corporation, 1990; EPA, 

23 1990). The exposure measurements and individual animal data from Thyssen et al. (1981) were 

24 used to calculate average continuous lifetime exposures for each individual hamster. Group 

25 averages of individual average continuous lifetime exposure concentrations were 0, 0.25, 1.01, and 

4.29 mg/m3 26 for the control through high-exposure groups, as described in Clement International 

27 Corporation (1990).
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1 Table 1-16. Tumors observed in chronic inhalation animal bioassays.

Study design and reference Results

Thyssen et al. (1981)

Syrian golden hamsters: male

(20–30 animals/group)

Target exposure concentrations: 0, 2, 10, or 50 mg/m3

Average exposure concentrations b: 0, 0.25, 1.01, and 

4.29 mg/m3

Inhalation: for 3–4.5 hours/5-7 days per week until 

hamsters died or became moribund

larynx – Incidences: 0/27; 0/27; 8/26; 13/25

tumor latency a: 107and 67.6 weeks

trachea – Incidences: 0/27; 0/27; 1/26; 3/25

tumor latency: 115 and 63 weeks

nasal cavity – incidences: 0/27; 0/27; 3/26; 1/25

tumor latency: 116 and 79 weeks

Revised tumor incidence data c

larynx – Incidences: 0/27; 0/27; 11/26; 12/34

pharynx – Incidences: 0/27; 0/27; 9/26; 18/34

larynx and pharynx (combined)d – Incidences: 0/27; 

0/27; 16/26; 18/34

pharynx – Incidences: 0/27; 0/27; 6/26; 14/25

esophagus – (papillomas and squamous cell 

carcinomas)

Incidences: 0/27; 0/27; 0/27; 2/25

Forestomach-(papillomas and squamous cell 

carcinomas)

Incidences: 0/27; 0/27; 1/26; 1/25

a 2 tumor latency provided for 10 and 50 mg/m3 dose groups

3 b Duration adjusted inhalation concentrations calculated from exposure chamber monitoring data and 

4 exposure treatment times obtained by Clement International Corporation (1990). Daily exposure times: 4.5 

5 hours/day, 5 days/week on weeks 1–12; 3 hours/day, 5 days/week on weeks 13–29; 3.7 hours/day, 

6 5 days/week on week 30; 3 hours/day, 5 days/week on weeks 31–41; and 3 hours/day, 7 days/week for 

7 reminder of the experiment.

8 c Revised tumor incidence data based on original study pathology reports obtained by Clement International 

9 Corporation (1990). 

10 dNasal, forestomach, esophageal, and tracheal tumors occurred in hamsters that also had tumors in the larynx 

11 or pharynx, except for two animals in the mid-dose group that displayed nasal tumors (one malignant and 

12 one benign) without displaying tumors in the pharynx or larynx.

13 Dermal Exposure

14 Repeated application of benzo[a]pyene to skin (in the absence of exogenous promoters) has 

15 been demonstrated to induce skin tumors in mice, rats, rabbits, and guinea pigs. These studies have

16 been reviewed by multiple national and international health agencies (IARC, 2010, 1983, 1973; 

17 WHO, 1998; ATSDR, 1995). The analysis in this document focuses on chronic carcinogenicity 

18 bioassays in several strains of mice following repeated dermal exposure to benzo[a]pyrene for the 

19 animals’ lifetime (Table 1-17). These studies involved 2- or 3-times/week exposure protocols, at 

20 least two exposure levels plus controls, and histopathological examinations of the skin and other 

21 tissues (Sivak et al., 1997; Grimmer et al., 1984, 1983; Habs et al., 1984, 1980; Schmähl et al., 1977; 

22 Schmidt et al., 1973; Roe et al., 1970; Poel, 1960, 1959). 

23 Mice have been the most extensively studied species in dermal carcinogenesis studies of 

24 benzo[a]pyrene because of evidence that they may be more sensitive than other animal species; 

25 however, comprehensive comparisons of species differences in sensitivity to lifetime dermal 
Toxicological Review of Benzo[a]pyrene

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1 exposure are not available. Systemic tumors in benzo[a]pyrene-treated mice were not observed in 

2 lifetime dermal bioassays in which macroscopic examination of internal organs was included 

3 (Higginbotham et al., 1993; Habs et al., 1980; Schmahl et al., 1977; Schmidt et al., 1973; Roe et al., 

4 1970; Poel, 1959). 

5 Table 1-17. Tumor observations in dermal animal bioassays.

Study design and reference Results

Poel (1959)

C57L mice: male (13–56/dose)

0, 0.15, 0.38, 0.75, 3.8, 19, 94, 188, 376, or 752 μg

Dermal - 3 times/week for up to 103 weeks or 

until the appearance of a tumor by gross 

examination

skin tumors – (gross skin tumors and epidermoid carcinoma) 

– dose-dependent decreased time of tumor appearance

Incidences:

Gross skin tumors: 0/33; 5/55; 11/55; 7/56; 41/49; 38/38; 

35/35; 12/14; 14/14; 13/13

Epidermoid carcinoma: 0/33; 0/55; 2/55; 4/56; 32/49; 37/38; 

35/35; 10/14; 12/14; 13/13

Poel (1960)

SWR, C3HeB, or A/He mice: male (14-25/dose)

0, 0.15, 0.38, 0.75, 3.8, 19.0, 94.0, or 470 μg

Dermal - 3 times/week until mice died or a skin 

tumor was observed

skin tumors and dose-dependent decreased time of first 

tumor appearance

Incidences:

SWR: 0/20; 0/25; 2/22; 15/18; 12/17; 16/16; 16/17; 14/14

C3HeB: 0/17; 0/19; 3/17; 4/17; 11/18;/ 17/17; 18/18; 17/17

A/He mice: 0/17; 0/18; 0/19; 0/17; 0/17; 21/23; 11/16; 17/17

Roe et al. (1970)

Swiss mice (50/dose)

0, vehicle, 0.1, 0.3, 1, 3, or 9 μg

Dermal - 3 times/week for up to 93 weeks

skin tumors – malignant skin tumors were observed in 4/41 

and 31/40 mice in the two high dose groups, respectively

Incidences:

0/43; 0/47; 1/42; 0/42; 1/43; 8/41; 34/46

Schmidt et al., 1973

NMRI mice: female (100/group)

Swiss mice: female (100/group)

0, 0.05, 0.2, 0.8, or 2 μg

Dermal - 2/week until spontaneous death 

occurred or until an advanced carcinoma was 

observed

skin tumors – (carcinomas)

Incidences:

NMRI: 

2/100 at 2 μg (papillomas); 

2/100 at 0.8 μg and 30/100 at 2 μg (carcinomas)

Swiss: 

3/80 at 2 μg (papillomas); 

5/80 at 0.8 μg and 45/80 at 2 μg (carcinomas)

Schmähl et al., 1977

NMRI mice: female (100/group)

0, 1, 1.7, or 3 μg

Dermal - 2 times/week

skin tumors – (papillomas and carcinomas)

Incidences:

0/81; 1/77; 0/88; 2/81 (papillomas)

0/81; 10/77; 25/88; 43/81 (carcinomas)

Habs et al., 1980

NMRI mice: female (40/group)

0, 1.7, 2.8, or 4.6 μg

Dermal - 2 times/week until natural death or gross 

observation of infiltrative tumor growth

skin tumors and dose-dependent increase in agestandardized tumor incidence

Incidences:

0/35; 8/34; 24/35; 22/36

Age-standarized tumor incidence:

0, 24.8, 89.3, 91.7%

Grimmer et al., 1984, 1983

CFLP mice: female (65–80/group)

0, 3.9, 7.7, and 15.4 μg (1983 study)

0, 3.4, 6.7, and 13.5 μg (1984 study)

skin tumors – (papillomas and carcinomas) – with a decrease 

in tumor latency

Incidences:

1983: 0/80; 7/65; 5/64; 2/64 (papillomas)
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Study design and reference Results

Dermal - 2 times/wk for 104 weeks 0/80; 15/65; 34/64; 54/64 (carcinomas)

1984: 0/65; 6/64; 8/65; 4/65 (papillomas)

 0/65; 37/64; 45/65; 53/64 (carcinomas)

Habs et al., 1984

NMRI mice: female (20/group)

0, 2, or 4 μg

Dermal - 2 times/week for life

skin tumors – (papillomas and carcinomas) – with a decrease 

in mean survival time

Incidences:

0/20; 2/20; 0/20 (papillomas)

0/20; 7/20; 17/20 (carcinomas)

Higginbotham et al., 1993

Ah-receptor-responsive Swiss mice: female

(23-23/group)

0, 0.25, 1 or 2 μg

Dermal - 2 times/wk for 40 weeks

Skin tumors were not observed

Sivak et al., 1997

C3H/HeJ mice: male (30/group)

0, 0.05, 0.5, or 5 μg

Dermal - 2 times/wk for 104 weeks

skin tumors – (papillomas and carcinomas)

Incidences:

0/30; 0/30; 5/30 (2 papillomas, 3 carcinomas); 27/30 (1 

papilloma, 28 carcinomas)

1 Mode of action analysis—Carcinogenicity

2 The carcinogenicity of benzo[a]pyrene, the most studied PAH, is well documented (IARC, 

3 2010; Xu et al., 2009; Jiang et al., 2007, 2005; Xue and Warshawsky, 2005; Ramesh et al., 2004; 

4 Bostrom et al., 2002; Penning et al., 1999; WHO, 1998; Harvey, 1996; ATSDR, 1995; Cavalieri and 

5 Rogan, 1995; U.S. EPA, 1991b). The primary mode of action by which benzo[a]pyrene induces 

6 carcinogenicity is via a mutagenic mode of action. This mode of action is presumed to apply to all 

7 tumor types and is relevant for all routes of exposure. The general sequence of key events 

8 associated with a mutagenic mode of action for benzo[a]pyrene are: (1) bioactivation of 

9 benzo[a]pyrene to DNA-reactive metabolites via three possible metabolic activation pathways: a 

10 diol epoxide pathway, a radical cation pathway, and an o-quinone and ROS pathway (2) direct DNA 

11 damage by reactive metabolites, including the formation of DNA adducts and ROS-mediated 

12 damage; (3) formation and fixation of DNA mutations, particularly in tumor suppressor genes or 

13 oncogenes associated with tumor initiation; and (4) clonal expansion of mutated cells during the 

14 promotion and progression phases of cancer development. These events are depicted in Figure 1-7. 

15 Also included in the figure are other processes that may contribute to the carcinogenicity of 

16 benzo[a]pyrene via the promotion and progression phases of cancer development, including 

17 inflammation, cytotoxicity, and sustained regenerative cell proliferation. The available human, 

18 animal, and in vitro evidence supports a mutagenic mode of action as the primary mode by which 

19 benzo[a]pyrene induces carcinogenesis. 

20
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1

2 Figure 1-7. Proposed metabolic activation pathways and key events in the 

3 carcinogenic mode of action for benzo[a]pyrene

4 Data in Support of the MOA

5 Summary of Metabolic Activation Pathways

6 Diol epoxide pathway

7 Benzo[a]pyrene diol epoxide metabolites, believed to be the most potent DNA-binding 

8 metabolites of benzo[a]pyrene, are formed through a series of Phase I metabolic reactions (see 

9 Appendix B of the Supplemental Information). The initial metabolism is carried out primarily by 

10 the inducible activities of CYP enzymes including CYP1A1, CYP1B1, and CYP1A2. Further 

11 metabolism by epoxide hydrolase and the mixed function oxidase system yields (+)-anti12 benzo[a]pyrene-7,8-diol-9,10-epoxide (BPDE), one of the most potent DNA-binding metabolites of 

13 benzo[a]pyrene. Benzo[a]pyrene diol epoxide metabolites interact preferentially with the exocyclic 

14 amino groups of deoxyguanine and deoxyadenine (Geacintov et al., 1997; Jerina et al., 1991). 

15 Adducts may give rise to mutations unless these adducts are removed by DNA repair processes 

16 prior to replication. The stereochemical nature of the diol epoxide metabolite (i.e., anti- versus syn17 diol epoxides) affects the number and type of adducts and mutation that occurs (Geacintov et al., 

benzo[a] 

pyrene

diol epoxide

radical cation

o-quinoneand 

ROS

DNA Adducts

DNA Adducts

DNA Adducts 

and oxidative 

base damage

Mutation 

(Transversion) 

K-Ras, H-Ras and 

p53 targets

Mutation 

(Depurination): 

H-ras target

Mutation 

(Depurination): 

Oxidative damage 

and strand scission

Proliferation of 

initiated cells

Neoplasm

Inflammatory 

response

Cytotoxicity 

HO

OH

O

+

Exposure Metabolism

O

O

Initiation Promotion Progression

Key events in the mode of action for benzo[a]pyrene carcinogenicity

Binding to AhR

Upregulation of genes 

related to 

biotransformation, 

growth, and 

differentiation
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1 1997). Transversion mutations (e.g., GC→TA or AT→TA) are the most common type of mutation 

2 found in mammalian cells following diol epoxide exposure (Bostrom et al., 2002).

3 Radical cation pathway

4 Radical cation formation involves a one-electron oxidation by CYP or peroxidase enzymes 

5 (i.e., horseradish peroxidase, prostaglandin H synthetase) that produce electrophilic radical cation 

6 intermediates (Cavalieri and Rogan, 1995, 1992). Radical cations can be further metabolized to 

7 phenols and quinones (Cavalieri et al., 1988d, e), or they can form unstable adducts with DNA that 

8 ultimately result in depurination (Cavalieri et al., 2005, 1993; Rogan et al., 1993). The predominant 

9 depurinating adducts occur at the N-3 and N-7 positions of adenine and the C-8 and N-7 positions of 

10 guanine (Cavalieri and Rogan, 1995).

11 o-Quinone/ROS pathway

12 The o-quinone metabolites of PAHs are formed by enzymatic dehydrogenation of 

13 dihydrodiols (Bolton et al., 2000; Penning et al., 1999; Harvey, 1996; ATSDR, 1995) (see Appendix B 

14 of the Supplemental Information). DHH (dihydrodiol dehydrogenase) enzymes are members of the 

15 α-keto reductase gene superfamily. o-Quinone metabolites are potent cytotoxins, are weakly 

16 mutagenic, and are capable of producing a broad spectrum of DNA damage. These metabolites can 

17 interact directly with DNA as well as result in the production of ROS (i.e., hydroxyl and superoxide 

18 radicals) that may produce further cytotoxicity and DNA damage. The o-quinone/ROS pathway 

19 also can produce depurinated DNA adducts from benzo[a]pyrene metabolites (Jiang et al., 2007, 

2005; McCoull et al., 1999). In this pathway, and in the presence of NAD(P)+ 20 , AKR oxidizes 

21 benzo[a]pyrene-7,8-diol to a ketol, which subsequently forms benzo[a]pyrene-7-8-dione. This and 

22 other PAH o-quinones react with DNA to form unstable, depurinating DNA adducts. In the presence 

23 of cellular reducing equivalents, o-quinones can also activate redox cycles, which produce ROS 

24 (Penning et al., 1996). DNA damage in in vitro systems following exposure to benzo[a]pyrene-7,8-

25 dione or other o-quinone PAH derivatives occurs through the AKR pathway and can involve the 

26 formation of stable DNA adducts (Balu et al., 2004), N-7 depurinated DNA adducts (McCoull et al., 

27 1999), DNA damage from ROS (8-oxo-dG) (Park et al., 2006), and strand scission (Flowers et al., 

28 1997, 1996). 

29 Summary of Genotoxicity and Mutagenicity

30 The ability of metabolites of benzo[a]pyrene to cause mutations and other forms of DNA 

31 damage in both in vivo and in vitro studies is well documented (see genotoxicity tables in Appendix 

32 B in Supplemental Information). With metabolic activation (e.g., the inclusion of S9), 

33 benzo[a]pyrene is consistently mutagenic in the prokaryotic Salmonella/Ames and Escherichia coli

34 assays. In mammalian in vitro studies, benzo[a]pyrene is consistently mutagenic, clastogenic, and 

35 induces cell transformation both with and without metabolic activation. Cytogenetic damage in the 

36 form of chromosomal aberrations (CA), micronuclei (MN), sister chromatid exchanges (SCE), and 

37 aneuploidy are commonplace following benzo[a]pyrene exposure as are DNA adduct formation, 
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1 single strand breaks (SSB), and induction of DNA repair and unscheduled DNA synthesis. In vitro 

2 mammalian cell assays have been conducted in various test systems, including human cell lines.

3 In the majority of in vivo studies, benzo[a]pyrene has tested positive in multiple species and 

4 strains and under various test conditions for cell transformation, CAs, DNA adducts, DNA strand 

5 breaks, MN formation, germline mutations, somatic mutations (H-ras, K-ras, p53, lacZ, hprt), and 

6 SCEs. Human studies are available following exposures to PAH mixtures through cigarette smoke 

7 or occupational exposure in which benzo[a]pyrene-specific DNA adducts have been detected, and it 

8 has been demonstrated qualitatively that benzo[a]pyrene metabolites damage DNA in exposed 

9 humans.

10 Experimental Support for the Hypothesized Mode of Action

11 EPA’s Cancer Guidelines (Section 2.4; 2005) describe a procedure for evaluating mode of 

12 action data for cancer. A framework for analysis of mode of action information is provided and 

13 followed below.

14 Strength, consistency, and specificity of association

15 Strong evidence links the benzo[a]pyrene diol epoxide metabolic activation pathway with 

16 key mutational events in genes that are associated with tumor initiation (mutations in the p53

17 tumor suppressor gene and H-ras or K-ras oncogenes) (Table 1-18). Results in support of a 

18 mutagenic MOA via benzo[a]pyrene diol epoxide include observations of frequent G→T 

19 transversion mutations in p53 and ras genes in lung tumors of human cancer patients exposed to 

20 coal smoke (Keohavong et al., 2003; DeMarini et al., 2001) or tobacco smoke (Pfeifer and Hainaut, 

21 2003; Pfeifer et al., 2002; Hainaut and Pfeifer, 2001, Bennett et al., 1999). These results are 

22 consistent with evidence that benzo[a]pyrene diol epoxide is reactive with guanine bases in DNA 

23 (Koreeda et al., 1978; Jeffrey et al., 1976); that G→T transversions, displaying strand bias, are the 

24 predominant type of mutations caused by benzo[a]pyrene in several biological systems (Liu et al., 

25 2005; Hainaut and Pfeifer, 2001; Marshall et al., 1984); and that sites of DNA adduction at guanine 

26 positions in cultured human HeLa or bronchial epithelial cells exposed to benzo[a]pyrene diol 

27 epoxide correspond to p53 mutational hotspots observed in human lung cancers (Denissenko et al., 

28 1996; Puisieux et al., 1991). In addition, mice exposed to benzo[a]pyrene in the diet (Culp et al., 

29 2000) or by i.p. injection (Nesnow et al., 1998a, b, 1996, 1995; Mass et al., 1993) had forestomach 

30 or lung tumors, respectively, showing frequent G→T or C transversions in the K-ras gene. 

31 Supporting evidence includes observations that benzo[a]pyrene diol epoxide [specifically (+)-anti32 BPDE] is more potent than benzo[a]pyrene itself, benzo[a]pyrene phenols, or benzo[a]pyrene diols 

33 in mutagenicity assays in bacterial and in vitro mammalian systems (Malveille et al., 1977; Newbold 

34 and Brookes, 1976) and in producing lung tumors in newborn mice following i.p. administration 

35 (Chang et al., 1987; Buening et al., 1978; Kapitulnik et al., 1978). Other supporting evidence 

36 includes observations of elevated BPDE-DNA adduct levels in respiratory tissue of lung cancer 

37 patients (Li et al., 2001; Xu et al., 1997) and white blood cells (WBC) of groups of coke oven 
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1 workers and chimney sweeps, occupations with known elevated risks of cancer (Vineis and Perera, 

2 2007; Pavanello et al., 1999), and in lung tissue from tobacco smokers with lung cancer (Rojas et al., 

3 2004; 1998; Andreassen et al., 2002; Godschalk et al., 2002; Alexandrov et al., 1992). Several 

4 epidemiological studies have indicated that PAH-exposed individuals who are homozygous for a 

5 CYP1A1 polymorphism, which increases the inducibility of this enzyme (thus increasing the 

6 capacity to produce benzo[a]pyrene diol epoxide), have increased levels of PAH or BPDE-DNA 

7 adducts (Bartsch et al., 2000; Aklillu et al., 2005; Alexandrov et al., 2002; Perera and Weinstein, 

8 2000).

9 Additional supporting evidence of a mutagenic mode of action for benzo[a]pyrene 

10 carcinogenicity is the extensive database of in vitro and in vivo studies demonstrating the 

11 genotoxicity and mutagenicity of benzo[a]pyrene following metabolic activation (Table 1-18). In 

12 vitro studies overwhelmingly support the formation of DNA adducts, mutagenesis in bacteria, yeast, 

13 and mammalian cells, several measures of cytogenetic damage (CA, SCE, MN), and DNA damage. In 

14 vivo systems in animal models are predominantly positive for somatic mutations following 

15 benzo[a]pyrene exposure.

16 Support for the radical cation activation pathway contributing to tumor initiation through 

17 mutagenic events includes observations that depurinated DNA adducts (expected products from 

18 reactions of benzo[a]pyrene radical cations with DNA) accounted for 74% of identified DNA 

19 adducts in mouse skin exposed to benzo[a]pyrene (Rogan et al., 1993) and 9/13 tumors examined 

20 from mice exposed to dermal applications of benzo[a]pyrene had H-ras oncogene mutations 

21 attributed to depurinated DNA adducts from benzo[a]pyrene radical cations (Chakravarti et al., 

22 1995).

23 Support for the o-quinone/ROS pathway contributing to tumor initiation via mutagenic 

24 events includes in vitro demonstration that several types of DNA damage can occur from o25 quinones and ROS (Park et al., 2006a; Balu et al., 2004; McCoull et al., 1999; Flowers et al., 1997, 

26 1996). In addition, benzo[a]pyrene-7,8-dione can induce mutations in the p53 tumor suppressor 

27 gene using an in vitro yeast reporter gene assay (Park et al., 2008; Shen et al., 2006; Yu et al., 2002), 

28 and dominant p53 mutations induced by benzo[a]pyrene-7,8-dione in this system corresponded to 

29 p53 mutational hotspots observed in human lung cancer tissue (Park et al., 2008).

30 Dose-response concordance and temporal relationship

31 In vivo demonstrations showing that benzo[a]pyrene-induced mutational events in p53 or 

32 ras oncogenes precede tumor formation are not available. In vitro exposure of human p53 knock-in 

33 murine fibroblasts to 1 μM benzo[a]pyrene for 4 to 6 days induced p53 mutations with similar 

34 features to those identified in p53 mutations in human lung cancer; i.e., predominance of G→T 

35 transversions with strand bias and mutational hotspots at codons 157-158 (Liu et al., 2005).

36 Bennett et al. (1999) demonstrated a dose-response relationship between smoking 

37 history/intensity and the types of p53 mutations associated with benzo[a]pyrene (G→T 

38 transversions) in human lung cancer patients (Table 1-18). In lung tumors of nonsmokers, 10% of 
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1 p53 mutations were G→T transversions, versus 40% in lung tumors from smokers with >60 pack2 years of exposure. A dose-response relationship also has been demonstrated between BPDE-DNA 

3 adduct levels in lung tissue and odds ratios for cancer in a case control study of 221 lung cancer 

4 cases and 229 healthy controls (Li et al., 2001).

5 In mice, dose-response and temporal relationships have been described between the 

6 formation of BPDE-DNA adducts and skin and forestomach tumors (Table 1-18). In a study using 

7 mice treated dermally with benzo[a]pyrene once or twice per week for up to 15 weeks (10, 25 or 

8 50 nmol benzo[a]pyrene per application), levels of benzo[a]pyrene-DNA adducts in the skin, lung, 

9 and liver increased with increasing time of exposure and increasing dose levels (Talaska et al., 

10 1996). Levels at the end of the exposure period were highest in the skin; levels in the lung and liver 

11 at the same time were 10- and 20-fold lower, respectively. Levels of benzo[a]pyrene-DNA adducts 

12 in skin and lung increased in an apparent biphasic manner showing a lower linear slope between 

13 the two lowest dose levels, compared with the slope from the middle to the highest dose.

14 Another study examined the dose-response relationship and the time course of 

15 benzo[a]pyrene-induced skin damage (Table 1-18), DNA adduct formation, and tumor formation in 

16 female mice. Mice were treated dermally with 0, 16, 32, or 64 μg of benzo[a]pyrene once per week 

17 for 29 weeks (Albert et al., 1991). Indices of skin damage and levels of BPDE-DNA adducts in skin 

18 reached plateau levels in exposed groups by 2–4 weeks of exposure. With increasing dose level, 

19 levels of BPDE-DNA adducts (fmol/μg DNA) initially increased in a linear manner and began to 

20 plateau at doses ≥32 μg/week. Tumors began appearing after 12–14 weeks of exposure for the 

21 mid- and high-dose groups and at 18 weeks for the low-dose group. At study termination 

22 (35 weeks after start of exposure), the mean number of tumors per mouse was approximately one 

23 per mouse in the low- and mid-dose groups and eight per mouse in the high-dose group. The time24 course data indicate that benzo[a]pyrene-induced increases in BPDE-DNA adducts preceded the 

25 appearance of skin tumors, consistent with the formation of DNA adducts as a precursor event in 

26 benzo[a]pyrene-induced skin tumors.

27 Culp et al. (1996a) compared dose-response relationships for BPDE-DNA adducts and 

28 tumors in female B6C3F1 mice exposed to benzo[a]pyrene in the diet at 0, 18.5, 90, or 350 μg/day 

29 for 28 days (to examine adducts) or 2 years (to examine tumors) (Table 1-18). The benzo[a]pyrene 

30 dose-tumor response data showed a sharp increase in forestomach tumor incidence between the 

31 18.5 μg/day group (6% incidence) and the 90 μg/day group (78% incidence). The BPDE-DNA 

32 adduct levels in forestomach showed a relatively linear dose-response throughout the 

33 benzo[a]pyrene dose range tested. The appearance of increased levels of BPDE-DNA adducts in the 

34 target tissue at 28 days is temporally consistent with the contribution of these adducts to the 

35 initiation of forestomach tumors. Furthermore, about 60% of the examined tumors had mutations 

36 in the K-ras oncogene at codons 12 and 13, which were G→T or G→C transversions indicative of 

37 BPDE reactions with DNA (Culp et al., 1996a).
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1 Biological plausibility and coherence

2 The evidence for a mutagenic mode of action for benzo[a]pyrene is consistent with the 

3 current understanding that mutations in p53 and ras oncogenes are associated with increased risk 

4 of tumor initiation (Table 1-18). The benzo[a]pyrene database is internally consistent in providing 

5 evidence for BPDE-induced mutations associated with tumor initiation in cancer tissue from

6 humans exposed to complex mixtures containing benzo[a]pyrene (Keohavong et al., 2003; Pfeifer 

7 and Hainaut, 2003; Pfeifer et al., 2002; DeMarini et al., 2001; Hainaut and Pfeifer, 2001, Bennett et 

8 al., 1999), in animals exposed to benzo[a]pyrene (Culp et al., 2000; Nesnow et al., 1998a, b, 1996, 

9 1995; Mass et al., 1993), and in in vitro systems (Denissenko et al., 1996; Puisieux et al., 1991). 

10 Consistent supporting evidence includes: (1) elevated BPDE-DNA adduct levels in respiratory tissue 

11 of lung cancer patients (Li et al., 2001; Xu et al., 1997) or tobacco smokers with lung cancer (Rojas 

12 et al., 2004; 1998; Andreassen et al., 2002; Godschalk et al., 2002; Alexandrov et al., 1992); (2) 

13 demonstration of dose-response relationships between G→T transversions in p53 mutations in 

14 lung tumors and smoking intensity (Bennett et al.,1999) and between odds ratios for lung cancer 

15 and BPDE-DNA adduct levels in lung tissue (Li et al., 2001); (3) the extensive database of in vitro 

16 and in vivo studies demonstrating the genotoxicity and mutagenicity of benzo[a]pyrene following 

17 metabolic activation; and (4) general concordance between temporal and dose-response 

18 relationships for BPDE-DNA adduct levels and tumor incidence in studies of animals exposed to 

19 benzo[a]pyrene (Culp et al., 1996a; Albert et al., 1991). There is also supporting evidence that 

20 contributions to tumor initiation through mutagenic events can be made by the radical cation 

21 (Chakravarti et al., 1995; Rogan et al., 1993) and o-quinone/ROS metabolic activation pathways 

22 (Park et al., 2008, 2006a; Shen et al., 2006; Balu et al., 2004; Yu et al., 2002; McCoull et al., 1999; 

23 Flowers et al., 1997, 1996).
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1 Table 1-18. Experimental support for the postulated key events for mutagenic 

2 MOA

1. Bioactivation of benzo[a]pyrene to DNA-reactive metabolites via three possible metabolic 

activation pathways: a diol epoxide pathway, a radical cation pathway, and an o-quinone and ROS 

pathway.

Evidence that benzo[a]pyrene metabolites induce key events:

Metabolism of benzo[a]pyrene via all three pathways has been demonstrated in multiple in vitro 

and in vivo studies in humans and animals (see Metabolic Activation Pathways section)

Multiple in vivo studies in humans and animals have demonstrated distribution of reactive 

metabolites to target tissues 

Human evidence that key events are necessary for carcinogenesis:

Humans with CYP polymorphisms or lacking a functional GSTM1 gene form higher levels of 

benzo[a]pyrene diol epoxides, leading to increased BPDE-DNA adduct formation and increased risk 

of cancer (Vineis et al., 2007; Pavanello et al., 2005; 2004; Alexandrov et al., 2002; Perera and 

Weinstein, 2000)

2. Direct DNA damage by the reactive metabolites, including the formation of DNA adducts and ROSmediated damage.

Evidence that benzo[a]pyrene metabolites induce key events:

Reactive benzo[a]pyrene metabolites have demonstrated genotoxicity in most in vivo and in vitro 

systems in which they have been tested, including the bacterial mutation assay, transgenic mouse 

assay, dominant lethal mutations in mice, BPDE-DNA adduct detection in humans and animals, and 

DNA damage, chromosomal aberrations, micronucleus formation, and sister chromatid exchange in 

animals (Appendix B in Supplemental Information)

Multiple in vivo benzo[a]pyrene animal exposure studies have demonstrated DNA adduct formation 

in target tissues that precede tumor formation and increase in frequency with dose (Culp et al., 

1996a; Talaska et al., 1996; Albert et al., 1991)

Benzo[a]pyrene diol epoxide metabolites interact preferentially with the exocyclic amino groups of 

deoxyguanine and deoxyadenine in DNA (Geacintov et al., 1997; Jerina et al., 1991; Koreeda et al., 

1978; Jeffrey et al., 1976)

Benzo[a]pyrene o-quinone metabolites are capable of activating redox cycles and producing ROS 

that cause oxidative base damage (Park et al., 2006a; Balu et al., 2004; McCoull et al., 1999; Flowers 

et al., 1997, 1996)

Human evidence that key events are necessary for carcinogenesis:

Detection of benzo[a]pyrene diol epoxide-specific DNA adducts is strongly associated with increased 

cancer risk in humans that are occupationally exposed (see Evidence in Humans section)

These benzo[a]pyrene diol epoxides formed BPDE-DNA adducts preferentially at guanine residues 

that have been detected in tissues of humans with cancer that were exposed to PAHs (Li et al., 2001; 

Xu et al., 1997; Vineis and Perera, 2007; Pavanello et al., 1999; Rojas et al., 2004; 1998; Andreassen 

et al., 2002; Godschalk et al., 2002; Alexandrov et al., 1992)

3. Formation and fixation of DNA mutations, particularly in tumor suppressor genes or oncogenes 

associated with tumor initiation.

Evidence that benzo[a]pyrene metabolites induce key events:

Several in vivo exposure studies have observed benzo[a]pyrene diol epoxide-specific mutational 

spectra (e.g., GT transversion mutations) in K-ras, H-ras, and p53 in forestomach or lung tumors 
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(Culp et al., 2000; Nesnow et al., 1998a, b, 1996, 1995; Mass et al., 1993)

Multiple animal exposure studies have identified benzo[a]pyrene-specific mutations in H-ras, K-ras, 

and p53 in target tissues preceding tumor formation (Liu et al., 2005; Wei et al., 1999; Culp et al., 

1996a)

Human evidence that key events are necessary for carcinogenesis:

DNA adducts formed by the benzo[a]pyrene diol epoxide reacting with guanine bases lead 

predominantly to GT transversion mutations; these specific mutational spectra have been 

identified in PAH-associated tumors in humans at mutational hotspots, including oncogenes (K-ras) 

and tumor suppressor genes (p53) (Denissenko et al., 1996; Puisieux et al., 1991; Liu et al., 2005; 

Marshall et al., 1984; Koreeda et al., 1978; Jeffrey et al., 1976; Keohavong et al., 2003; DeMarini et 

al., 2001; Pfeifer and Hainaut, 2003; Pfeifer et al., 2002; Hainaut and Pfeifer, 2001; Bennett et al., 

1999)

4. Clonal expansion of mutated cells during the promotion and progression phases of cancer 

development.

Evidence that benzo[a]pyrene metabolites induce key events:

Numerous studies in animals have observed carcinogenesis following exposure to benzo[a]pyrene 

via all exposure routes (see Evidence in Animals section)

Mice exposed dermally to benzo[a]pyrene for 26 weeks were found to have increased frequencies 

of H-ras mutations in exposure-induced hyperplastic lesions that were further increased in tumors 

(Wei et al., 1999)

Human evidence that key events are necessary for carcinogenesis:

The frequency of GT transversions in p53 was found to increase with increasing smoking duration 

and intensity in human lung tumors (Bennett et al., 1999)

Benzo[a]pyrene-specific DNA adducts have been detected in preneoplastic tissues in humans 

exposed to PAH mixtures, e.g., in lung tissue of smokers with lung cancer, and in human skin treated 

with coal tar containing ointment (Godschalk et al., 2002; Bartsch et al., 1999; Alexandrov et al., 

1992)

Detection of benzo[a]pyrene-specific DNA adducts is strongly associated with increased cancer risk 

in coke oven workers and chimney sweeps (Pavanello et al., 1999)

1 Other Possible Modes of Action

2 The carcinogenic process for benzo[a]pyrene is likely to be related to some combination of 

3 molecular events resulting from the formation of several reactive metabolites which interact with 

4 DNA to form adducts and produce DNA damage resulting in mutations in cancer-related genes, such 

5 as tumor suppressor genes or oncogenes. These events may reflect the initiation potency of 

6 benzo[a]pyrene. However, benzo[a]pyrene possesses promotional capabilities that may be related 

7 to AhR affinity, cytotoxicity and the formation of reactive oxygen species, as well as the inhibition of 

8 gap junctional intercellular communication.

9 The ability of certain PAHs to act as initiators and promoters may increase their 

10 carcinogenic potency (Andrews et al., 1978). The promotional effects of PAHs appear to be related 

11 to AhR affinity and the upregulation of genes related to growth and differentiation (Bostrom et al., 

12 2002). The genes regulated by this receptor belong to two major functional groups (i.e., induction 

13 of metabolism or regulation cell differentiation and proliferation). PAHs bind to the cytosolic AhR 
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1 in complex with heat shock protein 90. The ligand-bound receptor is then transported to nucleus in 

2 complex with the AhR nuclear translocator protein. The AhR complex interacts with AhR elements 

3 of DNA to increase the transcription of proteins associated with induction of metabolism and 

4 regulation of cell differentiation and proliferation. Following benzo[a]pyrene exposure, disparities 

5 have been observed in the tumor pattern and toxicity of Ah-responsive and Ah-nonresponsive mice, 

6 as Ah-responsive mice were more susceptible to tumorigenicity in target tissues such as liver, lung, 

7 and skin (Ma and Lu, 2007; Talaska et al., 2006; Shimizu et al., 2000).

8 Inflammatory responses to cytotoxicity may contribute to the tumor promotion process; for 

9 example, benzo[a]pyrene quinones (1,6-, 3,6-, and 6,12-benzo[a]pyrene-quinone) generated 

10 reactive oxygen species and increased cell proliferation by enhancing the epidermal growth factor 

11 receptor pathway in cultured breast epithelial cells (Burdick et al., 2003). In addition, several 

12 studies have demonstrated that exposure to benzo[a]pyrene increases the production of 

13 inflammatory cytokines which may contribute to cancer progression (N’Diaye et al., 2006; Tamaki 

14 et al., 2004; Garçon et al., 2001a, b)

15 Gap junctions are channels between cells that are crucial for differentiation, proliferation, 

16 apoptosis, and cell death and consequently for the two epigenetic steps of tumor formation, 

17 promotion, and progression. Inhibition of gap junctional intercellular communication by 

18 benzo[a]pyrene has been observed in vitro (Sharovskaya et al., 2006; Blaha et al., 2002).

19 Conclusions About the Hypothesized Mode of Action

20 There is sufficient evidence to conclude that the major mode of action for benzo[a]pyrene 

21 carcinogenicity involves mutagenicity mediated by DNA reactive metabolites. The evidence for a 

22 mutagenic mode of action for benzo[a]pyrene is consistent with the current understanding that 

23 mutations in p53 and ras oncogenes are associated with increased risk of tumor initiation. The 

24 benzo[a]pyrene database provides strong and consistent evidence for BPDE-induced mutations 

25 associated with tumor initiation in cancer tissue from humans exposed to complex mixtures 

26 containing benzo[a]pyrene, in animals exposed to benzo[a]pyrene, and in in vitro systems. 

27 Supporting evidence suggests that contributions to tumor initiation through potential mutagenic 

28 events can be made by the radical cation and o-quinone/ROS metabolic activation pathways. Other 

29 processes may contribute to the carcinogenicity of benzo[a]pyrene via the promotion and 

30 progression phases of cancer development (e.g., inflammation, cytotoxicity, sustained regenerative 

31 cell proliferation).

32 Support for the hypothesized mode of action in test animals

33 Benzo[a]pyrene induces gene mutations in a variety of in vivo and in vitro systems and 

34 produces tumors in all animal species tested and all routes of exposure (see Appendix B in 

35 Supplemental Information). Strong, consistent evidence in animal models supports the postulated 

36 key events: the metabolism of benzo[a]pyrene to DNA-reactive intermediates, the formation of 
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1 DNA adducts, and the subsequent occurrence of mutations in oncogenes and tumor suppressor 

2 genes.

3 Relevance of the hypothesized mode of action to humans

4 A substantial database indicates that the postulated key events for a mutagenic mode of 

5 action all occur in human tissues. Strong evidence is available from studies of humans exposed to 

6 PAH mixtures (including coal smoke and tobacco smoke) indicating a contributing role for 

7 benzo[a]pyrene diol epoxide in inducing key mutational events in genes that are associated with

8 tumor initiation (mutations in the p53 tumor suppressor gene and H-ras or K-ras oncogenes). The 

9 evidence includes observations of a spectrum of mutations in ras oncogenes and the p53 gene in 

10 lung tumors of human patients exposed to coal smoke or tobacco smoke) that are similar to the 

11 spectrum of mutations caused by benzo[a]pyrene diol epoxide in several biological systems, 

12 including tumors from mice exposed to benzo[a]pyrene. Additional supporting evidence includes 

13 correspondence between hotspots of p53 mutations in human lung cancers and sites of DNA 

14 adduction by benzo[a]pyrene diol epoxide in experimental systems, and elevated BPDE-DNA 

15 adduct levels in respiratory tissue of lung cancer patients or tobacco smokers with lung cancer.

16 Populations or lifestages particularly susceptible to the hypothesized mode of action

17 A mutagenic mode of action for benzo[a]pyrene-induced carcinogenicity is considered 

18 relevant to all populations and lifestages. The current understanding of biology of cancer indicates 

19 that mutagenic chemicals, such as benzo[a]pyrene, are expected to exhibit a greater effect in early 

20 life exposure versus later life exposure (U.S. EPA, 2005b; Vesselinovitch et al., 1979). Although the 

21 developing fetus and infants may have lower levels of some bioactivating enzymes than adults (e.g., 

22 CYP1A1/1B1), infants or children are expected to be more susceptible. Newborn or infant mice 

23 developed liver and lung tumors more readily than young adult mice following acute i.p. exposures 

24 to benzo[a]pyrene (Vesselinovitch et al., 1975). These results indicate that exposure to 

25 benzo[a]pyrene during early life stages presents additional risk for cancer, compared with 

26 exposure during adulthood. TheEPA’s Supplemental Guidance for Assessing Susceptibility from 

27 Early-Life Exposure to Carcinogens (U.S. EPA, 2005b) recommends the application of age-dependent 

28 adjustment factors (ADAFs) for carcinogens that act through a mutagenic mode of action. Given 

29 that a determination benzo[a]pyrene acts through a mutagenic mode of carcinogenic action has 

30 been made, ADAFs should be applied along with exposure information to estimate cancer risks for 

31 early-life exposure.

32 Population variability in metabolism and detoxification of benzo[a]pyrene, in addition to 

33 DNA repair capability, may affect cancer risk. Polymorphic variations in the human population in 

34 CYP1A1, CYP1B1, and other CYPs have been implicated as determinants of increased individual 

35 lung cancer risk in some studies (Aklillu et al., 2005; Alexandrov et al., 2002; Perera and Weinstein, 

36 2000). Some evidence suggests that humans lacking a functional GSTM1 gene have higher BPDE37 DNA adduct levels and are thus at greater risk for cancer (Vineis et al., 2007; Pavanello et al., 2005; 

38 2004; Alexandrov et al., 2002; Perera and Weinstein, 2000). In addition, acquired deficiencies or 
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1 inherited gene polymorphisms that affect the efficiency or fidelity of DNA repair may also influence 

2 individual susceptibility to cancer from environmental mutagens (Matullo et al., 2003; Shen et al., 

3 2003; Cheng et al., 2000; Perera and Weinstein, 2000; Wei et al., 2000; Amos et al., 1999). In 

4 general, however, available support for the role of single polymorphisms in significantly 

5 modulating human PAH cancer risk from benzo[a]pyrene or other PAHs is relatively weak or 

6 inconsistent. Combinations of metabolic polymorphisms, on the other hand, may be critical 

7 determinants of a cumulative DNA-damaging dose, and thus indicate greater susceptibility to 

8 cancer from benzo[a]pyrene exposure (Vineis et al., 2007).

9 1.2. Summary and Evaluation

10 1.2.1. Weight of Evidence for Effects Other Than Cancer

11 The weight of the evidence from human and animal studies indicate that the strongest 

12 evidence for potential hazard following benzo[a]pyrene exposure are for developmental and 

13 reproductive toxicity and Immunotoxicity. In humans, exposure to PAH mixtures has been shown 

14 to result in developmental and reproductive toxicity and immunotoxicity. Most of the available 

15 human data report associations between particular health endpoints and concentrations of 

16 benzo[a]pyrene-DNA adducts, with fewer studies correlating health effects with external measures 

17 of exposure. The available human studies report effects that are generally analogous to the effects 

18 observed in animal toxicological studies, and provide qualitative, supportive evidence for the effect19 specific hazards identified in Section 1.1..

20 In animals, evidence of developmental and reproductive toxicity and immunotoxicity has

21 been observed across species and dosing regimens (Figure 1-8). The available evidence from mice 

22 and rats treated by gavage during gestation or in the early postnatal period demonstrate 

23 developmental effects including decreased body weight, decreased fetal survival, decreased 

24 fertility, atrophy of reproductive organs, and altered neurobehavioral outcomes (Chen et al., 2012; 

25 Jules et al., 2012; Bouayed et al., 2009b; Kristensen et al., 1995; MacKenzie and Angevine 1981). 

26 Male and female reproductive toxicity, as evidenced by effects on sperm parameters, decreased 

27 reproductive organ weights, histological changes, and hormone alterations, have been observed 

28 after oral exposure in rats and mice (Chen et al., 2011; Chung et al., 2011; Mohamed et al., 2010; 

29 Zheng et al., 2010; MacKenzie and Angevine, 1981). Benzo[a]pyrene exposure has also been shown 

30 to lead to altered immune cell populations and histopathological changes in immune system organs

31 (Kroese et al., 2001; De Jong et al., 1999), as well as thymic and splenic effects following subchronic 

32 oral exposure. Varying immunosuppressive responses are also observed in short term oral and 

33 injection studies. The weight of the evidence indicates that developmental toxicity, reproductive 

34 toxicity and immunotoxicity are hazards following oral exposure to benzo[a]pyrene.

35 Following inhalation exposure to benzo[a]pyrene in animals, evidence of developmental 

36 and reproductive toxicity has been observed. Decreased fetal survival has been observed in rats 

37 exposed to benzo[a]pyrene via inhalation during gestation (Wormley et al., 2004; Archibong et al., 
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1 2002). Male reproductive toxicity, as evidenced by effects on sperm parameters, decreased testes 

2 weight, and hormone alterations, has also been observed in rats following subchronic inhalation

3 exposure to benzo[a]pyrene (Archibong et al., 2008; Ramesh et al., 2008). Female reproductive 

4 toxicity, as evidenced by modified hormone levels in dams, has been observed following inhalation 

5 exposure to benzo[a]pyrene during gestation (Archibong et al., 2002). The weight of the evidence 

6 indicates that developmental toxicity and reproductive toxicity are hazards following inhalation 

7 exposure to benzo[a]pyrene.

8 Forestomach hyperplasia was observed following oral and inhalation exposure; however, 

9 this endpoint most likely reflects early events in the neoplastic progression of forestomach tumors 

10 following benzo[a]pyrene exposure (see Section 1.1.4), and was not carried forward as an effect 

11 other than cancer for the derivation of reference values.

12 1.2.2. Weight of Evidence for Carcinogenicity

13 Under EPA’s Guidelines for Carcinogen Risk Assessment (U.S. EPA, 2005a), benzo[a]pyrene 

14 is “carcinogenic to humans.” EPA’s Cancer Guidelines (U.S. EPA, 2005a) emphasize the importance 

15 of weighing all of the evidence in reaching conclusions about human carcinogenic potential. The 

16 descriptor of “carcinogenic to humans” can be used when the following conditions are met: (a)

17 there is strong evidence of an association between human exposure and either cancer or the key 

18 precursor events of the agent's mode of action but not enough for a causal association, (b) there is 

19 extensive evidence of carcinogenicity in animals, (c) the mode or modes of carcinogenic action and 

20 associated key precursor events have been identified in animals, and (d) there is strong evidence 

21 that the key precursor events that precede the cancer response in animals are anticipated to occur 

22 in humans and progress to tumors, based on available biological information. The data supporting 

23 these four conditions for benzo[a]pyrene are presented below and in Table 1-19.

24 a) Strong human evidence of cancer or its precursors

25 There is a large body of evidence for human carcinogenicity for complex PAH mixtures 

26 containing benzo[a]pyrene, including soot, coal tars, coal-tar pitch, mineral oils, shale oils, and 

27 smoke from domestic coal burning (IARC, 2010a,b; Baan et al., 2009). There is also evidence of 

28 carcinogenicity, primarily of the lung and skin, in occupations involving exposure to PAH mixtures 

29 containing benzo[a]pyrene, such as chimney sweeping, coal gasification, coal-tar distillation, coke 

30 production, iron and steel founding, aluminum production, and paving and roofing with coal tar 

31 pitch (IARC, 2010a; Baan et al., 2009; Straif et al., 2005). Increased cancer risks have been reported 

32 among other occupations involving exposure to PAH mixtures such as carbon black and diesel 

33 exhaust (Bosetti et al., 2007). Benzo[a]pyrene is also a notable carcinogenic constituent of tobacco 

34 smoke (IARC, 2004). 

35 Evidence for a contributing role of benzo[a]pyrene to human carcinogenic responses to 

36 complex PAH mixtures is available. Elevated BPDE-DNA adducts have been reported in smokers 
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1 compared to nonsmokers, and the increased adduct levels in smokers are typically increased two2 fold compared with nonsmokers (Philips, 2002). Elevated BPDE-DNA adduct levels have been 

3 observed in WBCs of groups of coke oven workers and chimney sweeps, occupations with known 

4 elevated risks of cancer (Rojas et al., 2000, 1998; Bartsch et al., 1999, 1998; Pavanello et al, 1999), 

5 and in lung tissue from tobacco smokers with lung cancer (Rojas et al., 2004; 1998; Godschalk et al., 

6 2002; Bartsch et al., 1999; Andreassen et al., 1996; Alexandrov et al., 1992). 

7 Although it is likely that multiple carcinogens present in PAH mixtures contribute to the 

8 carcinogenic responses, strong evidence is available from several studies of humans exposed to 

9 PAH mixtures supporting a contributing role for benzo[a]pyrene diol epoxide in inducing key 

10 mutagenic precursor cancer events in target tissues. Distinctive mutation spectra have been 

11 observed in the tumor suppressor gene p53 and the K-ras oncogene in tumor tissues taken from 

12 lung cancer patients that were chronically exposed to two significant sources of PAH mixtures: coal 

13 smoke and tobacco smoke. Hackman et al. (2000) reported an increase of GCTA transversions 

14 and a decrease of GCAT transitions at the hprt locus in T-lymphocytes of humans with lung 

15 cancer that were smokers compared to nonsmokers. Lung tumors from cancer patients exposed to 

16 emissions from burning smoky coal showed mutations in p53 and K-ras that were primarily G→T 

17 transversions (76 and 86%, respectively) (DeMarini et al., 2001). Keohavong et al. (2003) 

18 investigated the K-ras mutational spectra from nonsmoking women and smoking men chronically 

19 exposed to emissions from burning smoky coal, and smoking men who resided in homes using 

20 natural gas; among those with K-ras mutations, 67, 86, and 67%, respectively, were G→T 

21 transversions. Lung tumors from tobacco smokers showed a higher frequency of p53 mutations 

22 that were G→T transversions compared with lung tumors in nonsmokers (Pfeifer and Hainaut, 

23 2003; Pfeifer et al., 2002; Hainaut and Pfeifer, 2001), and the frequency of these types of p53

24 mutations in lung tumors from smokers increased with increasing smoking intensity (Bennett et al., 

25 1999). 

26 Similarly, investigations of mutagenesis following specific exposures to benzo[a]pyrene (as 

27 opposed to PAH mixtures) have consistently observed that the benzo[a]pyrene diol epoxide is very 

28 reactive with guanine bases in DNA, and that G→T transversions are the predominant type of 

29 mutations caused by benzo[a]pyrene diol epoxide in several biological test systems (Pfeifer and 

30 Hainaut, 2003; Hainaut and Pfeifer, 2001). Following treatment of human HeLa cells with 

31 benzo[a]pyrene diol epoxide, Denissenko et al. (1996) reported that the distribution of BPDE-DNA 

32 adducts within p53 corresponded to mutational hotspots observed in p53 in human lung cancers. 

33 Benzo[a]pyrene exposure induced mutations in embryonic fibroblasts from human p53 “knock-in” 

34 mice that were similar to those found in smoking related human cancers, with a predominance of 

35 G→T transversions that displayed strand bias and were also located in the same mutational 

36 hotspots found in p53 in human lung tumors (Liu et al., 2005). These results, combined with a 

37 mechanistic understanding that mutations in p53 (which encodes a key transcription factor in DNA 

38 repair and regulation of cell cycle and apoptosis) may be involved in the initiation phase of many 
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1 types of cancer, are consistent with a common mechanism for mutagenesis following exposures to 

2 PAH mixtures and provide evidence of a contributing role of benzo[a]pyrene diol epoxide in the 

3 carcinogenic response of humans to coal smoke and tobacco smoke.

4 Therefore, while the epidemiological evidence alone does not establish a causal association 

5 between human exposure and cancer, there is strong evidence that the key precursor events of 

6 benzo[a]pyrene’s mode of action are likely to be associated with tumor formation in humans.

7 b) Extensive animal evidence

8 In laboratory animals (rats, mice, and hamsters), exposure to benzo[a]pyrene via the oral, 

9 inhalation, and dermal routes have been associated with carcinogenic responses both systemically 

10 and at the site of administration. Three 2-year oral bioassays are available that associate lifetime 

11 benzo[a]pyrene exposure with carcinogenicity at multiple sites. These bioassays observed 

12 forestomach, liver, oral cavity, jejunum, kidney, auditory canal (Zymbal’s gland), and skin or 

13 mammary gland tumors in male and female Wistar rats (Kroese et al., 2001); forestomach tumors 

14 in male and female Sprague-Dawley rats (Brune et al., 1981); and forestomach, esophagus, tongue, 

15 and larynx tumors in female B6C3F1 mice (Beland and Culp, 1998; Culp et al., 1998). Repeated or 

16 short-term oral exposure to benzo[a]pyrene was associated with forestomach tumors in additional 

17 bioassays with several strains of mice (Weyand et al., 1995; Benjamin et al., 1988; Robinson et al., 

18 1987; El Bayoumy, 1985; Triolo et al., 1977; Wattenberg, 1974; Field and Roe, 1970; Roe et al., 

19 1970; Biancifiori et al., 1967; Chouroulinkov et al., 1967; Fedorenko and Yansheva, 1967; Neal and 

20 Rigdon, 1967; Berenblum and Haran, 1955). Chronic inhalation exposure to benzo[a]pyrene was 

21 associated with tumors in the larynx and pharynx of male Syrian golden hamsters exposed to

22 benzo[a]pyrene:NaCl aerosols (Thyssen et al., 1981). Additionally, less-than-lifetime oral exposure 

23 cancer bioassays in mice provide supporting evidence that exposure to benzo[a]pyrene is 

24 associated with an increased incidence of lung tumors in mice (Weyand et al., 1995; Robinson et al., 

25 1987; Wattenberg, 1974). Intratracheal instillation of benzo[a]pyrene was associated with 

26 respiratory tract tumors in additional studies with hamsters (Feron and Kruysse, 1978; Ketkar et 

27 al., 1978; Feron et al., 1973; Henry et al., 1973; Saffiotti et al., 1972). Chronic dermal application of 

28 benzo[a]pyrene (2–3 times/week) has been associated with mouse skin tumors in numerous 

29 bioassays (Sivak et al., 1997; Habs et al., 1984, 1980; Grimmer et al., 1984, 1983; Schmähl et al., 

30 1977; Schmidt et al., 1973; Roe et al., 1970; Poel, 1960, 1959). Skin tumors in rats, rabbits, and 

31 guinea pigs have also been associated with repeated application of benzo[a]pyrene to skin in the 

32 absence of exogenous promoters (WHO, 1998; ATSDR, 1995; IARC, 1983, 1973). When followed by 

33 repeated exposure to a potent tumor promoter, acute dermal exposure to benzo[a]pyrene induced 

34 skin tumors in numerous studies of mice, indicating that benzo[a]pyrene is a strong tumor35 initiating agent in the mouse skin model (Weyand et al., 1992; Cavalieri et al., 1991, 1981; Rice et 

36 al., 1985; El-Bayoumy et al., 1982; LaVoie et al., 1982; Raveh et al., 1982; Slaga et al., 1980, 1978; 

37 Wood et al., 1980; Hoffmann et al., 1972).
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1 Carcinogenic responses in animals exposed to benzo[a]pyrene by other routes of 

2 administration include: (1) liver or lung tumors in newborn mice given acute postnatal i.p. 

3 injections (LaVoie et al., 1994, 1987; Busby et al., 1989, 1984; Weyand and LaVoie, 1988; Wislocki 

4 et al., 1986; Buening et al., 1978; Kapitulnik et al., 1978); (2) increased lung tumor multiplicity in 

5 A/J adult mice given single i.p. injections (Mass et al., 1993); (3) injection site tumors in mice 

6 following s.c. injection (Nikonova, 1977; Pfeiffer, 1977; Homburger et al., 1972; Roe and Walters, 

7 1967; Grant and Roe, 1963; Steiner, 1955; Rask-Nielson, 1950; Pfeiffer and Allen, 1948; Bryan and 

8 Shimkin, 1943; Barry et al., 1935); (4) injection site sarcomas in mice following intramuscular 

9 injection(Sugiyama, 1973); (5) mammary tumors in rats with intramammilary administration 

10 (Cavalieri et al., 1991, 1988a, b, c); (6) cervical tumors in mice with intravaginal application 

11 (Naslund et al., 1987); and (7) tracheal tumors in rats with intratracheal implantation (Topping et 

12 al., 1981, Nettesheim et al., 1977).

13 Therefore, the animal database provides extensive evidence of carcinogenicity in animals.

14 c) Identification of key precursor events have been identified in animals

15 There is sufficient evidence to conclude that benzo[a]pyrene carcinogenicity involves a 

16 mutagenic mode of action mediated by DNA reactive metabolites. The benzo[a]pyrene database 

17 provides strong and consistent evidence for BPDE-induced mutations associated with tumor 

18 initiation in cancer tissue from humans exposed to complex mixtures containing benzo[a]pyrene, in 

19 animals exposed to benzo[a]pyrene, and in in vitro systems. Other processes may contribute to the 

20 carcinogenicity of benzo[a]pyrene via the promotion and progression phases of cancer 

21 development (e.g., inflammation, cytotoxicity, sustained regenerative cell proliferation, anti22 apoptotic signaling), but the available evidence best supports a mutagenic mode of action as the 

23 primary mode by which benzo[a]pyrene acts.

24 d) Strong evidence that the key precursor events are anticipated to occur in humans 

25 Mutations in p53 and ras oncogenes have been observed in tumors from mice exposed to 

26 benzo[a]pyrene in the diet (Culp et al., 2000) or by intraperitoneal injection (Nesnow et al., 

27 1998a,b, 1996, 1995; Mass et al., 1993). Mutations in these same genes have also been reported in 

28 lung tumors of human cancer patients, bearing distinctive mutation spectra (G→T transversions) 

29 that correlate with exposures to coal smoke (Keohavong et al., 2003; DeMarini et al., 2001) or 

30 tobacco smoke (Pfeifer and Hainaut, 2003; Pfeifer et al., 2002; Hainaut and Pfeifer, 2001, Bennett et 

31 al., 1999).
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1 Table 1-19. Supporting evidence for the carcinogenic to humans cancer

2 descriptor for benzo[a]pyrene 

Evidence Reference

a) Strong human evidence of cancer or its precursors

Increased cancer risks in humans exposed to 

complex PAH mixtures containing benzo[a]pyrene IARC, 2010a,b, 2004

Benzo[a]pyrene-specific biomarkers detected in 

humans exposed to PAH mixtures

 

– BPDE-DNA adducts in WBCs of coke oven 

workers and chimney sweeps

Rojas et al., 2000, 1998; Bartsch et al., 1999, 1998; 

Pavanello et al, 1999

– BPDE-DNA adducts in smokers Philips, 2002 

Benzo[a]pyrene-specific DNA adducts have been 

detected in preneoplastic target tissues in humans 

exposed to PAH mixtures

– BPDE-DNA adducts in non-tumor lung tissues 

of cigarette smokers with lung cancer and in 

skin eczema patients treated with coal tar

Rojas et al., 2004; 1998; Godschalk et al., 2002, 1998; 

Bartsch et al., 1999; Andreassen et al., 1996; Alexandrov 

et al., 1992 

– BPDE-DNA adduct formation in p53 in human 

cells in vitro corresponds to mutational 

hotspots at guanine residues in human lung 

tumors

Denissenko et al., 1996; Puisieux et al., 1991

Benzo[a]pyrene-specific mutational spectra 

identified in PAH-associated tumors in humans

– GCTA transversions and GCAT transitions 

at hprt locus in T-lymphocytes of humans with 

lung cancer

Hackman et al., 2000

– GT transversions at the same mutational 

hotspot in p53 from smoking-related lung 

tumors in humans

Liu et al., 2005; Pfeifer and Hainaut, 2003; Pfeifer et al., 

2002; Hainaut and Pfeifer, 2001

– GT transversions at the same mutational 

hotspot in p53 and K-ras in human lung tumors 

associated with smoky coal exposures

Keohavong et al., 2003; DeMarini et al., 2001

– Increased percentage of GT transversions in 

p53 in smokers vs. nonsmokers Bennett et al., 1999

b) Extensive animal evidence

Oral exposures

Forestomach tumors in male and female rats and 

in female mice following chronic exposure

Kroese et al., 2001; Brune et al., 1981; Beland and Culp, 

1998; Culp et al., 1998

Forestomach tumors in mice following less-thanlifetime exposures

Weyand et al., 1995; Benjamin et al., 1988; Robinson et 

al., 1987; El Bayoumy, 1985; Triolo et al., 1977; 

Wattenberg, 1974; Field and Roe, 1970; Roe et al., 1970; 

Biancifiori et al., 1967; Chouroulinkov et al., 1967; 

Fedorenko and Yansheva, 1967; Neal and Rigdon, 1967; 

Berenblum and Haran, 1955
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Evidence Reference

Alimentary tract and Liver tumors in male and 

female rats following chronic exposure Kroese et al., 2001

Auditory canal tumors in male and female rats 

following chronic exposure Kroese et al., 2001

Esophageal, tongue, and laryngeal tumors in 

female mice following chronic exposure Beland and Culp, 1998; Culp et al., 1998

Lung tumors in mice following less-than-lifetime 

exposure

Weyand et al., 1995 ; Robinson et al., 1987; Wattenberg, 

1974

Inhalation exposures

Laryngeal and pharyngeal tumors in male 

hamsters following chronic exposure Thyssen et al., 1981

Dermal exposures

Skin tumors in mice following chronic exposures 

without a promoter or acute exposures with a 

promoter

Sivak et al., 1997; Habs et al., 1984, 1980; Grimmer et 

al., 1984, 1983; Schmähl et al., 1977; Schmidt et al., 

1973; Roe et al., 1970; Poel, 1960, 1959

Skin tumors in rats, rabbits, and guinea pigs 

following subchronic exposures WHO, 1998; ATSDR, 1995; IARC, 1983, 1973

Other routes of exposure

Respiratory tract tumors in hamsters following 

intratracheal instillation

Feron and Kruysse, 1978; Ketkar et al., 1978; Feron et 

al., 1973; Henry et al., 1973; Saffiotti et al., 1972

Liver or lung tumors in newborn mice given acute 

postnatal i.p. injections

LaVoie et al., 1994, 1987; Busby et al., 1989, 1984; 

Weyand and LaVoie, 1988; Wislocki et al., 1986; Buening 

et al., 1978; Kapitulnik et al., 1978

Lung tumor multiplicity in A/J adult mice given 

single i.p. injections Mass et al., 1993

c) Identification of key precursor events have been identified in animals

Bioactivation of benzo[a]pyrene to DNA-reactive 

metabolites has been shown to occur in multiple 

species and tissues by all routes of exposure

See ‘Experimental Support for Hypothesized Mode of 

Action’ section

Direct DNA damage by the reactive metabolites, 

including the formation of DNA adducts and ROSmediated damage

Formation and fixation of DNA mutations, 

particularly in tumor suppressor genes or 

oncogenes associated with tumor initiation

d) Strong evidence that the key precursor events are anticipated to occur in humans

Mutations in p53 or ras oncogenes have been 

observed in forestomach or lung tumors from mice 

exposed to benzo[a]pyrene

Culp et al., 2000; Nesnow et al., 1998a,b, 1996, 1995; 

Mass et al., 1993

– G→T transversions in ras oncogenes or the p53

gene have been observed in lung tumors of 

human cancer patients exposed to coal smoke

Keohavong et al., 2003; DeMarini et al., 2001

– Higher frequency of G→T transversions in lung 

tumors from smokers vs. nonsmokers

Pfeifer and Hainaut, 2003; Pfeifer et al., 2002; Hainaut 

and Pfeifer, 2001; Bennett et al., 1999
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2. DOSE-RESPONSE ANALYSIS

1 2.1. Oral Reference Dose for Effects Other Than Cancer

2 The RfD (expressed in units of mg/kg-day) is defined as an estimate (with uncertainty 

3 spanning perhaps an order of magnitude) of a daily exposure to the human population (including 

4 sensitive subgroups) that is likely to be without an appreciable risk of deleterious effects during a 

5 lifetime. It can be derived from a no-observed-adverse-effect level (NOAEL), lowest-observed6 adverse-effect level (LOAEL), or the 95 percent lower bound on the benchmark dose (BMDL), with 

7 uncertainty factors (UFs) generally applied to reflect limitations of the data used.

8 2.1.1. Identification of Studies and Effects for Dose-Response Analysis

9 In Section 1.2.1, developmental, reproductive, and immunological toxicities were 

10 highlighted as hazards of benzo[a]pyrene exposure by the oral route. Studies within each effect 

11 category were evaluated using general study quality characteristics (as discussed in Section 6 of the 

12 Preamble) to help inform the selection of studies from which to derive toxicity values. Rationales 

13 for selecting the studies and effects to represent each of these hazards are summarized below.

14 Human studies are preferred over animal studies when quantitative measures of exposure 

15 are reported and the reported effects are determined to be associated with exposure. For 

16 benzo[a]pyrene, human studies of environmental PAH mixtures across multiple cohorts have 

17 observed effects following exposure to complex mixtures of PAHs. The available human studies 

18 that utilized benzo[a]pyrene-DNA adducts as the exposure metric do not provide external exposure 

19 levels of benzo[a]pyrene from which to derive a value. Thus, these studies were not considered 

20 because of the contribution to the observed hazard of multiple PAHS across multiple routes of 

21 exposure. Animal studies were evaluated to determine which provided the most relevant routes 

22 and durations of exposure; multiple exposure levels to provide information about the shape of the 

23 dose response curve; and power to detect effects at low exposure levels.

24 Developmental toxicity

25 Numerous animal studies observed endpoints of developmental toxicity following oral 

26 exposure during gestational or early post-natal development (Chen et al., 2012; Jules et al., 2012; 

27 Bouayed, 2009b; Kristensen et al., 1995; MacKenzie and Angevine 1981) and were considered for 

28 dose response analysis based on the above criteria. Kristensen et al. (1995), with only one dose 

29 group, was not considered further given its concordance with MacKenzie and Angevine (1981) 

30 which had multiple groups. From the remaining studies demonstrating developmental toxicity, the 

31 studies conducted by Chen et al. (2012) and Jules et al. (2012) were identified as the most 
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1 informative studies for dose-response analysis. The neurodevelopmental study by Chen et al. 

2 (2012) was a well conducted study that evaluated multiple neurobehavioral endpoints and 

3 measures of neurotoxicity in adolescent and adult animals. The study was designed with sufficient 

4 statistical power and randomized pups with a total of 10 males and 10 females per treatment group 

5 and no more than one male and one female from each litter (i.e., pups from 40 litters were used) for 

6 behavioral testing. In addition, the pups were cross-fostered with dams being rotated among litters 

7 every 2-3 days to distribute any maternal caretaking differences randomly across litters and 

8 treatment groups. 

9 Chen et al. (2012) observed increased latency in negative geotaxis and increased motor 

10 activity or anxiety in the open field test at 0.02 mg/kg-day, and decreased anxiety in the elevated 

11 plus maze test and impaired spatial learning as measured by an increase in latency time to find a 

12 hidden platform in the Morris water maze test at 0.2 mg/kg-day. Altered behaviors and locomotion 

13 in open field tests can be attributed to anxiety responses due to open spaces and bright light, as 

14 well as changes to motor system function. Chen et al. (2012) reported increased quadrants crossed, 

15 which could indicate either increased motor activity or decreased anxiety (less fear of the open 

16 spaces/ bright lights). Similarly, this study reported the number of open arm entries in the elevated 

17 plus maze test and not time spent in the closed arms and open arms which would provide a more 

18 sensitive measurement of the effect. The number of open arm entries does serve as an indicator of 

19 decreased anxiety and is unlikely to be confounded by changes in motor activity (as total arm 

20 entries appears to be unchanged with treatment), but the magnitude of response in the elevated 

21 plus maze test was not as robust as the response observed in the Morris water maze response. 

Chen et al. (2012) also observed effects in the surface righting test that were on the order of 22 ~0.2-

23 0.3 seconds and in the negative geotaxis test of approximately 3-4 seconds with no automated 

24 recording of latency (such as use of video recordings). Additionally, male and female rats (which 

25 often show differences in the maturation of these developmental landmarks) were pooled for these 

26 measures. Due to these uncertainties altered behaviors and locomotion in open field tests, number 

27 of open arm entries in the elevated plus maze test, effects in the surface righting test, and effects in 

28 the negative geotaxis test were not considered.

29 Impaired spatial learning as measured by an increase in latency time to find a hidden 

30 platform in the Morris water maze test was selected for dose response analysis. This endpoint is 

31 supported by probe trials from the same study in which the platform was removed and the 

32 duration of time spent in the target quadrant and the number of times the animal crosses the 

33 location of the previous platform was measured. Similar impairments were observed in both the 

34 probe tests and water maze escape measures in rats treated with benzo[a]pyrene. Thus, these tests 

35 confirmed that the animals were using learned spatial cues to identify the location of the platform 

36 and not relying on random, non-spatial strategies such as circular swimming. In addition, there 

37 were no changes in measured swim speed indicating an effect on long term memory. 
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1 Jules et al. (2012) was also identified for dose-response analysis. This study was of 

2 sufficient duration, utilized multiple doses, did not observe maternal toxicity, and evaluated 

3 multiple cardiovascular endpoints. The study authors reported increases in both systolic

4 (approximately 20-50%) and diastolic (approximately 33-83%) pressure and heart rate in adult 

5 rats that were exposed gestationally to benzo[a]pyrene. A limitation of this study is that the 

6 authors only reported effects at the two highest doses. However, given the magnitude of the 

7 response and the appearance of these effects in adulthood following gestational exposure, these 

8 endpoints were selected for dose-response analysis because of their sensitivity and biological 

9 plausibility.

10 Bouayed et al. (2009b) and MacKenzie and Angevine (1981) were not selected for dose 

11 response analysis. Bouayed et al. (2009b) used the same tests as Chen et al. (2012), but at higher 

12 doses (2 and 20 mg/kg-day compared to 0.02, 0.2, and 2 mg/kg-day, respectively). Because Chen et 

13 al. (2012) reported adverse effects at doses lower than Bouayed (2009b), the later study was not 

14 selected for dose-response analysis. Similarly, MacKenzie and Angevine (1981) demonstrated 

15 developmental effects in a multi-dose study with relevant routes and durations of exposure;

16 however, the doses studied (10-160 mg/kg-day) were much higher than those evaluated in other 

17 developmental toxicity studies (Chen et al., 2012; Jules et al., 2012).

18 Reproductive toxicity

19 Male reproductive toxicity was demonstrated in numerous subchronic studies (Chen et al., 

20 2011a; Chung et al., 2011; Mohamed et al., 2010; Zheng et al., 2010). Chung et al. (2011) was not 

21 included in the dose-response analysis because of study reporting limitations (i.e., only reported 

22 significant observations at the mid-dose). Chen et al. (2011a) is a subchronic study that applied 

23 only a single dose level. Because the study corroborated available multi-dose studies it was not 

24 considered for dose-response analysis. The studies conducted by Mohamed et al. (2010) and Zheng 

25 et al (2010) were identified as the most informative male reproductive toxicity studies for dose26 response analysis. Decreased sperm count observed by Mohammed et al. (2010) and decreased 

27 intratesticular testosterone levels observed by Zheng et al. (2010) were selected for dose-response 

28 analysis as both represent sensitive endpoints of male reproductive toxicity and are indicators of 

29 potentially decreased fertility. These effects are also consistent with human studies in PAH 

30 exposed populations as effects on male fertility and semen quality have been demonstrated in 

31 epidemiological studies of smokers (reviewed by Soares and Melo, 2008). 

32 Female reproductive toxicity was demonstrated in two subchronic studies (Gao et al., 2011;

33 Xu et al., 2010). Specifically, these studies demonstrated altered ovarian weights and follicle 

34 numbers and cervical epithelial cell hyperplasia following oral exposure to benzo[a]pyrene. These

35 studies were identified as the most informative studies for dose-response analysis. Gao et al. 

36 (2011) identified statistically-significant dose-related increases in the incidence of cervical 

37 inflammatory cells in mice exposed to low doses of benzo[a]pyrene for 98 days (Gao et al., 2011, 
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1 2010). Cervical effects of increasing severity (including epithelial hyperplasia, atypical hyperplasia, 

2 apoptosis, and necrosis) were also observed at higher doses (Gao et al., 2011, 2010). There are no 

3 data on cervical effects in other species or in other mouse strains. However, Gao et al. (2011) also 

4 evaluated cervical effects in separate groups of mice exposed via i.p. injection, and observed similar 

5 responses in these groups of mice, providing support for the association between effects in this 

6 target organ and benzo[a]pyrene exposure. Epidemiological studies have demonstrated an 

7 association between cigarette smoking and increased risk of cervical cancer (Pate Capps et al., 

8 2009). In addition, benzo[a]pyrene metabolites and benzo[a]pyrene-DNA adducts have been 

9 detected in human cervical mucus and cervical tissues obtained from smokers (Melikian et al., 

10 1999; Phillips et al., 2002). 

11 Xu et al. (2010) identified biologically and statistically significant decreases in ovary weight, 

12 estrogen, and primordial follicles, and altered estrus cycling in treated animals. These reductions in 

13 female reproductive parameters are supported by a large database of animal studies indicating that 

14 benzo[a]pyrene is ovotoxic with effects including decreased ovary weight, decreased primordial 

15 follicles, and reduced fertility (Mattison et al., 1980; MacKenzie and Angevine 1981; Swartz and 

16 Mattison 1985; Miller et al., 1992; Kristensen et al., 1995; Borman et al, 2000). Additionally, 

17 epidemiology studies indicate that exposure to complex mixtures of PAHs, such as through cigarette 

18 smoke, is associated with measures of decreased fertility in humans (El Nemr et al., 1998; Neal et 

19 al., 2005). Specific associations have also been made between infertility and increased levels of 

20 benzo[a]pyrene in follicular fluid in women undergoing in vitro fertilization (Neal et al., 2008).

21 Immunotoxicity

22 As described in Section 1.1.4, the immune system was identified as a target of 

23 benzo[a]pyrene-induced toxicity based on findings of organ weight and immunoglobulin 

24 alterations, as well as effects on cellularity and functional changes in the immune system in animals. 

25 The studies conducted by Kroese et al. (2001) and De Jong et al. (1999) are subchronic studies with 

26 multiple exposure levels and adequate power to detect effects. In comparing these two studies, the 

27 Kroese et al. (2001) study is preferred for dose-response analysis due to its longer duration (90 

28 days).

29 Decreased thymus weight, observed in Kroese et al. (2001), decreased IgM and IgA levels, 

30 and decreased numbers of B-cells, observed in De Jong et al. (1999), were selected for dose31 response analysis. It is recognized that thymus weight changes on their own have been noted to be 

32 less reliable indicators of immunotoxicity (Luster et al., 1992). However, there are converging lines 

33 of evidence that support the derivation of a candidate RfD for benzo[a]pyrene immunotoxicity, 

34 including: alterations in immunoglobulin levels have been noted in humans after exposure to PAHs, 

35 as well as in animal studies after exposure to benzo[a]pyrene; changes in B cell populations in the 

36 spleen provide additional evidence of immunotoxicity; and functional effects on the immune 

37 system, including dose-related decreases in SRBC-specific IgM levels and dose-dependent decreases 
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1 in resistance to pneumonia or Herpes simplex type 2 following short-term s.c. injection (Temple et 

2 al. 1993; Munson et al. 1985). The observed decreases in thymus weight, IgM and IgA levels, and 

3 number of B cells associated with exposure to benzo[a]pyrene were determined to be 

4 representative of immunotoxicity following benzo[a]pyrene exposure and were selected for dose5 response analysis.

6 2.1.2. Methods of Analysis

7 No biologically based dose-response models are available for benzo[a]pyrene. In this 

8 situation, EPA evaluates a range of dose-response models thought to be consistent with underlying 

9 biological processes to determine how to best empirically model the dose-response relationship in 

10 the range of the observed data. Consistent with this approach, all models available as part of EPA’s 

11 Benchmark Dose Software (BMDS) were evaluated. Consistent with EPA’s draft Benchmark Dose 

12 Technical Guidance Document (U.S. EPA, 2000b), the BMD and the 95% lower confidence limit on 

13 the BMD (BMDL) were estimated using a BMR of 1 SD from the control mean for continuous data or 

14 a BMR of 10% extra risk for dichotomous data in the absence of information regarding what level of 

15 change is considered biologically significant, and also to facilitate a consistent basis of comparison 

16 across endpoints, studies, and assessments. The estimated BMDLs were used as points of 

17 departure (PODs). Further details including the modeling output and graphical results for the best 

18 fit model for each endpoint can be found Appendix C of the Supplemental Information.

19 Among the endpoints identified as representative of the hazards of benzo[a]pyrene

20 exposure, the data for Morris water maze escape latency (Chen et al., 2012), decreased ovary 

21 weight (Xu et al., 2010), and decreased thymus weight (Kroese et al., 2000) were amenable to dose22 response modeling. For the water maze escape latency data, the data for male and female rats were 

23 combined for dose-response analysis because of the strong similarity in responses and the lack of 

24 information available suggesting there would be sex-specific differences in the results of this test 

25 (see Appendix C of the Supplemental Information for details of statistical analyses). 

26 The data for the remaining endpoints identified in Section 2.1 were not modeled. 

27 Specifically, the data for cardiovascular effects observed in Jules et al. (2012) were limited due to 

28 the reporting of results at only the two highest dose groups. The data for epididymal sperm counts 

29 presented in the Mohamed et al. (2010) study were reported graphically only and requests for the 

30 raw data were unsuccessful. The observed decrease in IgM and IgA (De Jong et al., 1999) was 

31 inconsistent and not amenable to dose-response modeling. NOAELs or LOAELs were used as the 

32 POD for these endpoints.

33 Human equivalent doses (i.e. HEDs) for oral exposures were derived from the PODs 

34 estimated from the laboratory animal data as described in EPA’s Recommended Use of Body 

Weight 35 3/4 as the Default Method in Derivation of the Oral Reference Dose (U.S. EPA, 2011b). In this 

36 guidance, EPA advocates a hierarchy of approaches for deriving HEDs from data in laboratory 

37 animals, with the preferred approach being physiologically-based toxicokinetic modeling. Other 
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1 approaches can include using chemical-specific information in the absence of a complete 

2 physiologically-based toxicokinetic model. As discussed in Appendix B of the Supplemental 

3 Information, several animal PBPK models for benzo[a]pyrene have been developed and published, 

4 but a validated human PBPK model for benzo[a]pyrene for extrapolating doses from animals to 

5 humans is not available. In lieu of either chemical-specific models or data to inform the derivation 

of human equivalent oral exposures, a body weight scaling to the 3⁄4 power (i.e., BW3/4 6 ) approach is 

7 applied to extrapolate toxicologically equivalent doses of orally administered agents from adult 

8 laboratory animals to adult humans for the purpose of deriving an oral RfD. 

9 Consistent with EPA guidance (U.S. EPA, 2011b), the PODs estimated based on effects in 

10 adult animals are converted to HEDs employing a standard dosimetric adjustment factor (DAF) 

11 derived as follows:

12

DAF = (BWa

1/4 / BWh1/4 13 ),

14 Where

15 BWa = animal body weight

16 BWh = human body weight 

17

18 Using a BWa of 0.25 kg for rats and 0.035 kg for mice and a BWh of 70 kg for humans (U.S. 

19 EPA, 1988), the resulting DAFs for rats and mice are 0.24 and 0.15, respectively. Applying this DAF 

20 to the POD identified for effects in adult rats or mice yields a PODHED as follows (see Table 2-1): 

21

22 PODHED = Laboratory animal dose (mg/kg-day) x DAF

23

24 Table 2-1 summarizes the sequence of calculations leading to the derivation of a human25 equivalent POD for each data set discussed above.
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1 Table 2-1. Summary of Derivation of Points of Departure

Endpoint and 

Reference

Species/

Sex Modela

BMR BMD

(mg/kg-day)

BMDL

(mg/kg-day)

PODADJ

b

(mg/kg-day)

PODHED

c 

(mg/kg-day)

Developmental

Neurodevelopmental 

impairments 

Chen et al., 2012

Male and 

Female

SpragueDawley Rats

Hilla1SD 0.11 0.06 0.06 0.06

Cardiovascular effects

Jules et al., 2012

Long-Evans 

rats

LOAEL (0.6 mg/kg-day) 0.6 0.15

Reproductive

Decreased ovarian 

weight 

Xu et al., 2010 

Female

SpragueDawley ratsLineara1 SD 2.3 1.5 1.5 0.37

Decreased sperm 

count 

Mohamed et al., 2010

Male C57BL/6 

mice LOAEL (1 mg/kg-day) 1 0.15

Cervical epithelial 

hyperplasia

Gao et al. (2011)

Female ICR 

mice

Loglogistica 10% 0.58 0.37 0.37 0.06

Immunological

Decreased thymus 

weight

Kroese et al., 2001

Female 

Wistar rats Lineara

1SD 10.5 7.6 7.6 1.9

Decreased IgM levels

De Jong et al., 1999

Male 

Wistar rats NOAEL (10 mg/kg-day) 7.1 1.7

Decreased IgA levels 

De Jong et al., 1999

Male 

Wistar rats NOAEL (30 mg/kg-day) 21 5.2

Decreased number of 

B cells 

DeJong et al., 1999

Male 

Wistar rats NOAEL (30 mg/kg-day) 21 5.2

a

2 For modeling details, see Appendix B in Supplemental Information

b

3 For studies in which animals were not dosed daily, administered doses were adjusted to calculate the time4 weighted average daily doses prior to BMD modeling.

c

HED PODs were calculated using BW3/4 5 scaling (US EPA, 2011b) for effects from dosing studies in adult animals 

6 (i.e., Mohamed et al., 2010; Xu et al., 2010; Gao et al., 2011; and De Jong et al., 1999) or for developmental effects 

resulting from in utero exposures. BW3/4 7 scaling was not employed for deriving HEDs from studies in which doses

8 were administered directly to early postnatal animals (i.e., Chen et al., 2012) because of the absence of 

9 information on whether allometric (i.e., body weight) scaling holds when extrapolating doses from neonatal 

10 animals to adult humans due to presumed toxicokinetic and/or toxicodynamic differences between lifestages (U.S. 

11 EPA, 2011b; Hattis et al., 2004).
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1 2.1.3. Derivation of Candidate Reference Doses

2 Table 2-2 is a continuation of Table 2-1 and summarizes the application of uncertainty 

3 factors to each POD to derive a candidate RfD for each data set. The selection of uncertainty factors 

4 was based on EPA’s A Review of the Reference Dose and Reference Concentration Processes (U.S. EPA, 

5 2002; Section 4.4.5) and is described in Section 7.6 of the Preamble. Figure 2-1 presents graphically 

6 these candidate RfDs, uncertainty factors, and points of departure, with each bar corresponding to 

7 one data set described in Tables 2-1 and 2-2. 

8 Table 2-2. Effects and corresponding derivation of candidate RfDs

Endpoint and Reference PODHED

a POD 

type

UFA UFH UFL UFS UFD

Composite

UF

Candidate

RfD (mg/kg-d)

DEVELOPMENTAL

Neurodevelopmental 

impairments in rats 

Chen et al., 2012

0.06 BMDL1SD 10 10 1 1 3 300 2 x 10-4

Cardiovascular effects in rats

Jules et al., 2012

0.15 LOAEL 3 10 10 1 3 1000 2 x 10-4

REPRODUCTIVE

Decreased ovary weight and 

ovarian follicles in rats 

Xu et al., 2010

0.37 BMDL1SD 3 10 1 10 3 1000 4 x 10-4

Decreased sperm count in mice 

Mohamed et al., 2010

0.15 LOAEL 3 10 10 10 3 10000 Not calculated due 

to UF > 3000a

Cervical epithelial hyperplasia

in mice

Gao et al. (2011)

0.06 BMDL10 3 10 1 10 3 1000 6 x 10-5

IMMUNOLOGICAL

Decreased thymus weight in 

rats

Kroese et al., 2001

1.9 BMDL1SD 3 10 1 10 3 1000 2 x 10-3

Decreased serum IgM in rats 

De Jong et al., 1999

1.7 NOAEL 3 10 1 10 3 1000 2 x 10-3

Decreased serum IgA in rats 

De Jong et al., 1999

5.2 NOAEL 3 10 1 10 3 1000 5 x 10-3

Decreased number of B cells in 

rats 

De Jong et al., 1999

5.2 NOAEL 3 10 1 10 3 1000 5 x 10-3
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Endpoint and Reference PODHED

a POD 

type

UFA UFH UFL UFS UFD

Composite

UF

Candidate

RfD (mg/kg-d)

UFA – A value of 3 (100.5 = 3.16, rounded to 3) was applied to account for uncertainty in characterizing 

toxicodynamic differences between rats and humans when an HED was calculated using BW3/4 scaling as 

uncertainty in characterizing toxicokinetic differences was accounted for through calculation of an HED using 

a standard DAF consistent with EPA guidance (U.S. EPA 2011b). A value of 10 was applied when BW3/4 scaling 

was not employed to account for uncertainty in extrapolating from laboratory animals to humans because of 

the absence of information to characterize either the toxicokinetic or toxicodynamic differences between 

animals and humans following oral exposure to BaP. 

UFH – A value of 10 was applied to account for potentially susceptible individuals because adequate information is 

not available to quantitatively estimate variability in human susceptibility. In the case of benzo[a]pyrene, 

Insufficient information is available to quantitatively estimate variability in human susceptibility.

UFL – A value of 1 was applied when the POD is based on dose-response modeling or a NOAEL; 10 when the POD 

is a LOAEL. In the case of benzo[a]pyrene, An UFL of 1 was applied for LOAEL-to-NOAEL extrapolation because 

a BMR of a 1 SD change from the control mean in neurodevelopmental impairments was selected under an 

assumption that it represents a minimal biologically significant response level. 

UFS – A value of 1 was applied when dosing occurred during gestation or the early postnatal period that is 

relevant to developmental effects (U.S. EPA, 1991a); 10 when the POD is based on a subchronic study (studies 

in this table, other than the developmental toxicity studies, were 42-90 days in duration) to account for the 

possibility that longer exposure may induce effects at a lower dose.

UFD – A value of 3 was applied to account for database deficiencies including the lack of a standard 

multigenerational study or extended 1-generation study that includes exposure from premating through 

lactation, considering that benzo[a]pyrene has been shown to affect fertility in adult male and female animals 

by multiple routes of exposure (see Section 1.1.2). Also, the lack of a study examining functional neurological 

endpoints following a more comprehensive period of developmental exposure (i.e., gestation through 

lactation) is a data gap, considering human and animal evidence indicating altered neurological development 

(see Section 1.1.1).

a

As recommended in EPA's A Review of the Reference Dose and Reference Concentration Processes (U.S. EPA, 

2002), the derivation of a reference value that involves application of the full 10-fold uncertainty factor in four 

or more areas of extrapolation should be avoided.
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1 Figure 2-1. Candidate RfDs with corresponding POD and composite UF

2

3 2.1.4. Derivation of Organ/System-specific Reference Doses

4 Table 2-3 distills the candidate reference doses from Table 2-2 into a single value for each 

5 organ or system. These organ or system-specific reference values may be useful for subsequent 

6 cumulative risk assessments that consider the combined effect of multiple agents acting at a 

7 common site. 
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1 Developmental Toxicity

2 The candidate RfD based on neurodevelopmental impairment in rats (Chen et al., 2012) was 

3 selected as the organ/system-specific RfD representing developmental toxicity. This candidate RfD 

4 was selected because it is associated with the application of the smaller composite uncertainty 

5 factor and because similar effects were replicated across other studies.

6 Reproductive Toxicity

7 The candidate RfD based on decreased ovary weight and ovarian follicle numbers in rats 

8 from the Xu et al. (2010) study was selected as the organ/system-specific RfD representing 

9 reproductive toxicity. The ovarian effects are supported by a large database of animal studies and 

10 human studies of exposure to benzo[a]pyrene and PAH mixtures. The data supporting cervical 

11 effects associated with oral benzo[a]pyrene exposure are limited to a single study; however the 

12 finding is supported by corollary findings after i.p. exposure and by studies in humans. 

13 Immunotoxicity

14 The candidate RfDs based on decreased thymus weight (Kroese et al., 2001) and serum IgM 

15 levels in rats (DeJong et al., 1999) were selected as the organ/system-specific RfD representing 

16 immunotoxicity. The observed decreases in thymus weight, IgM and IgA levels, and number of B 

17 cells associated with exposure to benzo[a]pyrene were determined to be representative of 

18 immunotoxicity. In combination, these effects provide more robust evidence of immunotoxicity. 

19 The candidate RfDs for decreased thymus weight (Kroese et al., 2001) and serum IgM levels in rats 

20 (DeJong et al., 1999) were equal and provided the most senstitive candidate RfDs, thus these 

21 candidate RfDs were selected as the organ/system-specific RfDs representing immunotoxicity. 

22 Table 2-3. Organ/system-specific RfDs and proposed overall RfD for 

23 benzo[a]pyrene

Effect Basis RfD (mg/kg-d) Confidence

Developmental Neurodevelopmental impairments 2 x 10-4 MEDIUM

Reproductive Decreased ovary weight and ovarian follicles 4 x 10-4 MEDIUM

Immunological Decreased thymus weight and serum IgM 2 x 10-3LOW

Proposed Overall RfD Developmental toxicity 2 x 10-4 MEDIUM

24 2.1.5. Selection of the Proposed Overall Reference Dose

25 To estimate an exposure level below which the effects identified as potential hazards from 

26 benzo[a]pyrene exposure are not expected to occur, the lowest organ/system-specific RfD (2 x 10-4

27 mg/kg-day) is proposed as the overall reference dose for benzo[a]pyrene. This value, based on 
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1 induction of neurodevelopmental impairments in rats exposed to benzo[a]pyrene during a 

2 susceptible lifestage is supported by a large number of animal and human studies. 

3 2.1.6. Confidence Statement

4 A confidence level of high, medium, or low is assigned to the study used to derive the RfC, 

5 the overall database, and the RfC itself, as described in Section 4.3.9.2 of EPA’s Methods for 

6 Derivation of Inhalation Reference Concentrations and Application of Inhalation Dosimetry (U.S. EPA, 

7 1994).

8 Confidence in the principal study (Chen et al., 2012) is medium-to-high. The study design 

9 included randomized experimental testing, blinded observations, culling of pups to account for 

10 nutritional availability, treatment-randomization; and controls for litter and nursing bias. Some 

11 informative experimental details were, however, omitted including the sensitivity of some assays at 

12 the indicated developmental ages and lack of reporting gender-specific data for all outcomes. 

13 Notably, these study limitations do not apply to the endpoint chosen to derive the RfD, and the 

14 overall methods and reporting are considered sufficient. Confidence in the database is medium, 

15 primarily due to the lack of a multigenerational reproductive toxicity study given the sensitivity to 

16 benzo[a]pyrene during development. Reflecting medium-to-high confidence in the principal study 

17 and medium confidence in the database, confidence in the RfD is medium.

18 2.1.7. Previous Reference Dose

19 An RfD was not derived in the previous IRIS assessment.
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1 2.2. Inhalation Reference Concentration for Effects Other Than Cancer

The RfC (expressed in units of mg/m3 2 ) is defined as an estimate (with uncertainty spanning 

3 perhaps an order of magnitude) of a continuous inhalation exposure to the human population 

4 (including sensitive subgroups) that is likely to be without an appreciable risk of deleterious effects 

5 during a lifetime. It can be derived from a NOAEL, LOAEL, or the 95 percent lower bound on the 

6 benchmark concentration (BMCL), with UFs generally applied to reflect limitations of the data used.

7 2.2.1. Identification of Studies and Effects for Dose-Response Analysis

8 In Section 1.2.1, developmental and reproductive toxicities were highlighted as hazards of 

9 benzo[a]pyrene exposure by the inhalation route. Studies within each effect category were 

10 evaluated using general study quality characteristics (as discussed in Section 6 of the Preamble) to 

11 help inform the selection of studies from which to derive toxicity values. Rationales for selecting 

12 the studies and effects to represent each of these hazards are summarized below.

13 Human studies of environmental PAH mixtures across multiple cohorts have observed 

14 developmental and reproductive effects following prenatal exposure; however, these studies are 

15 limited by exposure to complex mixtures of PAHs and within individual studies there may have 

16 been more than one route of exposure. In addition, the available human studies that utilized

17 benzo[a]pyrene-DNA adducts as the exposure metric do not provide external exposure levels of 

18 benzo[a]pyrene from which to derive a reference concentration. Although preferred for derivation 

19 of reference values, human studies were not considered because of the contribution to the observed 

20 hazard of multiple PAHs across multiple routes of exposure.

21 Animal studies were evaluated to determine which provided the most relevant routes and 

22 durations of exposure; multiple exposure levels to provide information about the shape of the dose 

23 response curve; and power to detect effects at low exposure levels. The only chronic inhalation 

24 study available for benzo[a]pyrene, Thyssen et al. (1981), was designed as a cancer bioassay and 

25 did not report other effects. However, several reproductive and developmental toxicity studies are 

26 available in which effects on fetal survival and the male reproductive system have been observed.

27 Developmental toxicity

28 Developmental toxicity, as represented by decreased fetal survival and developmental 

29 neurotoxicity, was observed by Archibong et al. (2002), Wu et al. (2003), and Wormley et al. 

30 (2004). Wu et al. (2003) was not considered for dose-response analysis due to lack of study details 

31 related to number of dams and litters per group and lack of reporting of numerical data. From the 

32 remaining studies demonstrating developmental toxicity, the studies conducted by Archibong et al. 

33 (2002) and Wormley et al. (2004) were identified as the most informative studies for dose34 response analysis. Archibong et al. (2002) observed biologically significant effects at the lowest 

dose tested by the inhalation route (i.e., LOAEL of 25 μg/m3 35 ). This study indicates that the 

36 developing fetus is a sensitive target following inhalation exposure to benzo[a]pyrene. The 
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1 observed decrease in fetal survival is supported by the available oral database for benzo[a]pyrene

2 (e.g., decreased survival of litters in mice following in utero exposure to benzo[a]pyrene on GD 7-

3 16) (MacKenzie and Angevine, 1981). In addition, a single exposure inhalation study by Wormley 

4 et al. (2004) demonstrated developmental toxicity, represented by decreased pups/litter and 

5 electrophysiological changes in the hippocampus, as a result of gestational exposure. The single 

6 exposure concentration used by Wormley et al. study was the equivalent of the high-dose exposure 

7 concentration applied in the Archibong et al. (2002) and Wu et al. (2003) studies, thus the endpoint 

8 of fetal survival observed by Wormley et al. (2004) was not considered for dose-response analysis. 

9 However, similar to oral studies of benzo[a]pyrene exposure, Wormley et al. (2004) observed 

10 effects indicative of developmental neurotoxicity and therefore these effects were considered 

11 further for dose-response analysis. 

12 Reproductive toxicity

13 Reproductive toxicity, as represented by reductions in sperm quality, both count and 

14 motility, and testis weights in adults, was observed by Archibong et al. (2008) and Ramesh et al. 

15 (2008) and Archibong et al. (2002). Archibong et al. (2008) and Ramesh et al. (2008) reported the 

16 results of a single exposure, subchronic inhalation exposure study in male rats. This subchronic 

17 study was of sufficient duration and possessed adequate power to detect effects, but utilized a 

18 single exposure concentration which is less informative for dose-response analysis than a design 

19 using multiple exposure concentrations. The endpoints of decreased testes weight and sperm 

20 count and motility reported in Archibong et al. (2008) were selected for dose-response analysis as 

21 both represent sensitive endpoints of male reproductive toxicity and are indicators of potentially 

22 decreased fertility. These effects are also consistent with human studies in PAH exposed 

23 populations as effects on male fertility and semen quality have been demonstrated in 

24 epidemiological studies of smokers (reviewed by Soares and Melo, 2008). 

25 2.2.2. Methods of Analysis

26 Data for decreased fetal survival (Archibong et al., 2002) were not amenable to BMD 

27 modeling due to the pattern of variability (heterogeneous variances) in the data set; the response at 

28 the lowest exposure showed the greatest variability. Therefore, the LOAEL from this study was

29 used as the POD for dose-response analysis. Wormley et al. (2004) and Archibong et al. (2008),

30 using only one exposure level, were judged not to support dose-response modeling due to the lack 

31 of understanding of the underlying dose-response relationship (i.e., limited database). LOAELs 

32 were also used as the PODs for dose-respopnse analysis.

33 By definition, the RfC is intended to apply to continuous lifetime exposures for humans (U.S. 

34 EPA, 1994b). EPA recommends that adjusted continuous exposures be used for inhalation 

35 developmental toxicity studies as well as for studies of longer durations (U.S. EPA, 2002). The 
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1 LOAELs identified from Archibong et al. (2002), Archibong et al. (2008), and Wormley et al. (2004) 

2 were adjusted to account for the discontinuous daily exposure as follows:

3

4 PODADJ = POD × hours exposed per day/24 hours

5 = LOAEL × (duration of exposure/24 hours)

6 = PODADJ

7

8 Next, the human equivalent concentration (HEC) was calculated from the PODADJ by 

9 multiplying by a dosimetric adjustment factor (DAF), which, in this case, was the regional deposited 

10 dose ratio (RDDRER) for extrarespiratory (i.e., systemic) effects as described in Methods for 

11 Derivation of Inhalation Reference Concentrations and Application of Inhalation Dosimetry (U.S. EPA, 

12 1994b). The observed developmental effects are considered systemic in nature (i.e., 

13 extrarespiratory) and the normalizing factor for extrarespiratory effects of particles is body weight. 

14 The RDDRER was calculated as follows:

TOT H

TOT A

E H

E A

A

H

E R (F )

(F )

(V )

(V )

BW

BW RDDR

15

16 where:

17 BW = body weight (kg)

18 VE = ventilation rate (L/minute)

19 FTOT = total fractional deposition

20

21 The total fractional deposition (FTOT) includes particle deposition in the nasal-pharyngeal, 

22 tracheobronchial, and pulmonary regions. FTOT for both animals and humans was calculated using 

23 the Multi-Path Particle Dosimetry model, a computational model used for estimating human and rat 

24 airway particle deposition and clearance [Multi-Path Particle Dosimetry (MPPD); Version 2.0 © 

25 2006, publicly available through the Hamner Institute]. FTOT was based on the average particle size 

26 of 1.7 ± 0.085 (MMAD ± geometric SD) as reported in Wu et al. (2003) for the exposure range 25 -

100 um3 27 . For the model runs, the Yeh-Schum 5-lobe model was used for the human and the 

28 asymmetric multiple path model was used for the rat (see Appendix C for MPPD model output). 

29 Both models were run under nasal breathing scenarios with the inhalability adjustment selected. A 

30 geometric SD of 1 was used as the default by the model because the reported geometric SD of 0.085 

31 was ≤1.05.

32 The human parameters used in the model for calculating FTOT and in the subsequent 

33 calculation of the PODHEC were as follows: human body weight, 70 kg; VE, 13.8 L/minute; breathing 

34 frequency, 16 per minute; tidal volume, 860 mL; functional residual capacity, 3,300 mL; and upper 

35 respiratory tract volume, 50 mL. Although the most sensitive population in Archibong et al. (2002) 

36 and Wormley et al. (2004) studies is the developing fetus, the adult rat dams were directly exposed. 

37 Thus, adult rat parameters were used in the calculation of the HEC. The parameters used for the rat 
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1 were body weight, 0.25 kg (a generic weight for male and female rats) ; VE, 0.18 L/minute; 

2 breathing frequency, 102 per minute; tidal volume, 1.8 mL; functional residual capacity, 4 mL; and 

3 upper respiratory tract volume, 4.42 mL. All other parameters were set to default values (see 

4 Appendix C).

5 Under these conditions, the MPPD model calculated FTOT values of 0.621 for the human and 

6 0.181 for the rat. Using the above equation, the RDDRER was calculated to be 1.1.

7 From this, the PODHEC was calculated as follows:

8 PODHEC = PODADJ × RDDRER

9

10 Table 2-4 summarizes the sequence of calculations leading to the derivation of a human11 equivalent concentration for each data set discussed above.

12 Table 2-4. Summary of Derivation of Points of Departure

Endpoint and 

Reference Study Design Model BMR BMC(μg/m3)

BMCL

(μg/m3)

PODADJ

a

(μg/m3)

PODHEC

b

(μg/m3)

DEVELOPMENTAL

Decreased fetal survival

Archibong et al., 2002

Pregnant rats GD 11-20, 

4hrs/d 7d/wk LOAEL (25 μg/m3

) 4.2 4.6

Decreased long term 

potentiation in 

hippocampus

Wormley et al., 2004

Pregnant rats GD 11-21, 

4 hrs/d, 7d/wk LOAEL (100 μg/m3

) 16.7 18.3

REPRODUCTIVE

Decreased testis weight 

Archibong et al., 2008

Male rats, 60d, 4hrs/d, 

7d/wk LOAEL (75 μg/m3

) 12.5 13.8

Decreased sperm count 

and motility

Archibong et al., 2008

Male rats, 60d, 4hrs/d, 

7d/wk LOAEL (75 μg/m3

) 12.5 13.8

13 a PODs were adjusted for continuous daily exposure: PODADJ= POD × hours exposed per day/24 hours

14 b PODHEC calculated by adjusting the PODADJ by the regional deposited dose ratio calculated using particle size 

15 reported in Hood et al. (2000) using MPPD software as detailed in Section 2.2.2. and Appendix C in the 

16 Supplemental Information.

17 2.2.3. Derivation of Candidate Reference Concentrations

18 Table 2-5 is a continuation of Table 2-4 and summarizes the application of uncertainty 

19 factors to each POD to derive a candidate reference concentration for each data set. The selection 

20 of uncertainty factors was based on EPA’s A Review of the Reference Dose and Reference 

21 Concentration Processes (U.S. EPA, 2002; Section 4.4.5) and is described in the Section 7.6 of the 

22 Preamble. Figure 2-2 presents graphically these candidate reference doses, uncertainty factors, and 

23 points of departure, with each bar corresponding to one data set described in Tables 2-4 and 2-5. 
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1 Table 2-5. Effects and corresponding derivation of candidate RfCs

Endpoint PODHEC

(μg/m3)

POD 

type

UFA UFH UFL UFS UFD

Composite 

UF

Candidate

RfC (mg/m3)

a

DEVELOPMENTAL

Decreased fetal survival

in rats

Archibong et al., 2002

4.6 LOAEL 3 10 10 1 10 3000 2 x 10-6

Decreased long term 

potentiation in the 

hippocampus of rats

Wormley et al., 2004

18.3 LOAEL 3 10 10 1 10 3000 6 x 10-6

REPRODUCTIVE

Decreased testis weight 

in rats

Archibong et al., 2008

13.8 LOAEL 3 10 10 10 10 30000

Not calculated 

due to UF > 

3000b

Decreased sperm count 

and motility in rats

Archibong et al., 2008

13.8 LOAEL 3 10 10 10 10 30000

Not calculated 

due to UF > 

3000b

UFA – 3 to account for residual toxicodynamic uncertainties when an HEC was calculated by inhalation particle 

dosimetry methods (U.S. EPA, 1994b), as is the case with benzo[a]pyrene.

UFH – 10 to account for potentially susceptible individuals when adequate information is not available to 

quantitatively estimate variability in human susceptibility. In the case of benzo[a]pyrene, Insufficient 

information is available to quantitatively estimate variability in human susceptibility to benzo(a)pyrene.

UFL – 1 when the POD is based on BMD modeling or a NOAEL; 10 or 3 when the POD is a LOAEL. In the case of 

benzo[a]pyrene, UFL of 10 was applied to account for the use of a LOAEL. A NOAEL was not identified for 

decreased fetal survival observed by Archibong et al (2002). At the lowest dose, benzo[a]pyrene treated 

dams gave birth to 15% fewer pups compared to dams treated with vehicle alone (carbon black particles). 

Due to the lack of a NOAEL and the inability to model the data set for decreased fetal survival, a UF of 10 was 

applied to extrapolate to a NOAEL. 

UFS – 1 when dosing occurred during gestation or the early postnatal period that is relevant to developmental 

effects (U.S. EPA, 1991a); 10 when the POD is based on a subchronic study (studies in this table were 60 days 

in duration) to account for the possibility that longer exposure may induce effects at a lower dose

UFD – 10 to account for database deficiencies including the lack of a standard multigenerational study or extended 

1-generation study that includes exposure from premating through lactation, considering that benzo[a]pyrene 

has been shown to affect fertility in adult male and female animals by multiple routes of exposure (see 

Section 1.1.2). According to EPA’s A Review of the Reference Dose and Reference Concentration Processes

(U.S. EPA, 2002; Section 4.4.5), the UFD is intended to account for the potential for deriving an 

underprotective RfD/RfC as a result of an incomplete characterization of the chemical’s toxicity, but also 

including a review of existing data that may also suggest that a lower reference value might result if additional 

data were available. In the case of benzo[a]pyrene, oral exposure studies have demonstrated effects at doses 

lower than those where mortality was observed in the inhalation study by Archibong et al. (2002). The lack of 

a study examining functional neurological endpoints following a more comprehensive period of 

developmental exposure (i.e. gestation through lactation) is also a data gap, considering human and animal 

evidence indicating altered neurological development (see Section 1.1.1).

a

Candidate RfCs were converted from μg/m3to mg/m3

b

As recommended in EPA's A Review of the Reference Dose and Reference Concentration Processes (U.S. EPA, 

2002), the derivation of a reference value that involves application of the full 10-fold uncertainty factor in four or 

more areas of extrapolation should be avoided.
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1 Figure 2-2. Candidate RfCs with corresponding POD and composite UF

2

3 2.2.4. Derivation of Organ/System-specific Reference Concentrations

4 Table 2-6 distills the candidate reference concentrations from Table 2-5 into a single value 

5 for each organ or system. These organ or system-specific reference values may be useful for 

6 subsequent cumulative risk assessments that consider the combined effect of multiple agents acting 

7 at a common site. The candidate RfCs for reproductive toxicity derived from Archibong et al. 

8 (2008) were not selected to represent reproductive toxicity because as recommended in EPA's A 

9 Review of the Reference Dose and Reference Concentration Processes (U.S. EPA, 2002), the derivation 

10 of a reference value that involves application of the full 10-fold uncertainty factor in four or more 

11 areas of extrapolation should be avoided.
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1 Table 2-6. Organ/system-specific RfCs and proposed overall RfC for 

2 benzo[a]pyrene

Effect Basis RfC (mg/m3) Confidence

Developmental Decreased fetal survival 2 x 10-6low-to-medium

Reproductive Reductions in sperm parameters Not calculated NA

Proposed Overall RfC Decreased fetal survival 2 x 10-6low-to-medium

3 2.2.5. Selection of the Proposed Overall Reference Concentration

4 The study by Archibong et al. (2002) was selected as the study used for the derivation of the 

5 proposed overall RfC, as it observed biologically significant effects at the lowest dose tested by the 

6 inhalation route. This study indicates that the developing fetus is a sensitive target following 

7 inhalation exposure to benzo[a]pyrene and the observed decreases in number of pups/litter and 

8 fetal survival/litter are the most sensitive noncancer effects observed following inhalation 

9 exposure to benzo[a]pyrene. Additional support for this endpoint is provided by the oral studies of 

10 benzo[a]pyrene. A developmental/reproductive study conducted via the oral route in mice 

11 observed decreased survival of litters, decreased pup weight, and decreased reproductive organ 

12 weight following in utero exposure to benzo[a]pyrene on GD 7–16 (MacKenzie and Angevine, 

13 1981).

14 2.2.6. Confidence Statement

15 A confidence level of high, medium, or low is assigned to the study used to derive the RfC, 

16 the overall database, and the RfC itself, as described in Section 4.3.9.2 of EPA’s Methods for 

17 Derivation of Inhalation Reference Concentrations and Application of Inhalation Dosimetry (U.S. EPA, 

18 1994).

19 The overall confidence in the RfC is low-to-medium. Confidence in the principal study 

20 (Archibong et al., 2002) is medium. The conduct and reporting of this developmental dietary study 

21 were adequate; however, a NOAEL was not identified. Confidence in the database is low due to the 

22 lack of a multigeneration toxicity study, the lack of studies on immune endpoints, and the lack of 

23 information regarding subchronic and chronic inhalation exposure. However, confidence in the RfC 

24 is bolstered by consistent systemic effects observed by the oral route (including reproductive and 

25 developmental effects) and similar effects observed in human populations exposed to PAH 

26 mixtures. Reflecting medium confidence in the principal study and low confidence in the database, 

27 confidence in the RfC is low-to-medium.

28 2.2.7. Previous Reference Concentration

29 An RfC was not derived in the previous IRIS assessment.
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1 2.2.8. Uncertainties in the Derivation of the RfD and RfC

2 The following discussion identifies uncertainties associated with the RfD and RfC for 

3 benzo[a]pyrene. To derive the RfD, the UF approach (U.S. EPA, 2000, 1994b) was applied to a POD 

4 based on neurodevelopmental impairments in rats treated developmentally. To derive the RfC, this 

5 same approach was applied to a POD from a developmental study for the effect of decreased fetal 

6 survival. UFs were applied to the POD to account for extrapolating from an animal bioassay to 

7 human exposure, the likely existence of a diverse population of varying susceptibilities, and 

8 database deficiencies. These extrapolations are carried out with default approaches given the lack 

9 of data to inform individual steps. 

10 The database for benzo[a]pyrene contains limited human data. The observation of effects 

11 associated with benzo[a]pyrene exposure in humans is complicated by several factors including the 

12 existence of benzo[a]pyrene in the environment as one component of complex mixtures of PAHs, 

13 exposure to benzo[a]pyrene by multiple routes of exposure within individual studies, and the 

14 difficulty in obtaining accurate exposure information. Data on the effects of benzo[a]pyrene alone 

15 are derived from a large database of studies in animal models. The database for oral 

16 benzo[a]pyrene exposure includes two chronic bioassays in rats and mice, two developmental 

17 studies in mice, and several subchronic studies in rats.

18 Although the database is adequate for RfD derivation, there is uncertainty associated with 

19 the database including that the principal study for the RfD exposed animals during a relatively short 

20 period of brain development potentially underestimating the magnitude of resulting neurological 

21 effects. Also, the database lacks a comprehensive multi-generation reproductive/developmental 

22 toxicity studies and immune system endpoints were not evaluated in the available chronic-duration 

23 or developmental studies. Additionally, the only available chronic studies of oral or inhalational 

24 exposure to benzo[a]pyrene focused primarily on neoplastic effects leaving non-neoplastic effects 

25 mostly uncharacterized. 

26 The only chronic inhalation study of benzo[a]pyrene was designed as a lifetime 

27 carcinogenicity study and did not examine noncancer endpoints (Thyssen et al., 1981). However, 

28 subchronic and short-term inhalation studies are available, which examine developmental and 

29 reproductive endpoints in rats. Developmental studies by the inhalation route identified 

30 biologically significant reductions in the number of pups/litter and percent fetal survival and 

31 possible neurodevelopmental effects (e.g., diminished electrophysiological responses to stimuli in 

32 the hippocampus) following gestational exposures. Additionally, a 60-day oral study in male rats 

33 reported male reproductive effects (e.g., decreased testes weight and sperm production and 

34 motility), but provides limited information to characterize dose-response relationships with 

35 chronic exposure scenarios. One area of uncertainty pertains to the lack of information regarding 

36 fertility in animals exposed gestationally to benzo[a]pyrene, especially in light of developmental 

37 studies by the oral route indicating reduced fertility in the F1 generation and decreased 

38 reproductive organ weights. The database also lacks a multigenerational reproductive study via the 
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1 inhalation route. Areas of uncertainty include the lack of chronic inhalation studies focusing on 

2 noncancer effects, limited data on dose-response relationships for impaired male or female fertility 

3 with gestational exposure or across several generations, and limited data on immune system 

4 endpoints with chronic exposure to benzo[a]pyrene.

5 The toxicokinetic and toxicodynamic differences for benzo[a]pyrene between the animal 

6 species in which the POD was derived and humans are unknown. PBPK models can be useful for 

7 the evaluation of interspecies toxicokinetics; however, the benzo[a]pyrene database lacks an 

8 adequate model that would inform potential differences. There is some evidence from the oral 

9 toxicity data that mice may be more susceptible than rats to some benzo[a]pyrene effects (such as 

10 ovotocity [Borman et al., 2000]), although the underlying mechanistic basis of this apparent 

11 difference is not understood. Most importantly, it is unknown which animal species may be more 

12 comparable to humans.

13 2.3. Oral Slope Factor for Cancer

14 2.3.1. Analysis of Carcinogenicity Data

15 The database for benzo[a]pyrene contains numerous cancer bioassays that identify tumors, 

16 primarily of the alimentary tract including the forestomach, following oral exposure in rodents. 

17 Three 2-year oral bioassays are available that associate lifetime benzo[a]pyrene exposure with 

18 carcinogenicity at multiple sites: forestomach, liver, oral cavity, jejunum, kidney, auditory canal 

19 (Zymbal’s gland) tumors, and skin or mammary gland tumors in male and female Wistar rats 

20 (Kroese et al., 2001); forestomach tumors in male and female Sprague-Dawley rats (Brune et al., 

21 1981); and forestomach, esophageal, tongue, and larynx tumors in female B6C3F1 mice (Beland and 

22 Culp, 1998; additional results reported by Culp et al., 1998).

23 In addition to these 2-year cancer bioassays, there are studies available that provide 

24 supporting evidence of carcinogenicity but are less suitable for dose-response analysis due to one 

25 or more limitation in study design: (1) no vehicle control group, (2) only one benzo[a]pyrene dose 

26 group, or (3) a one-time exposure to benzo[a]pyrene (Benjamin et al., 1988; Robinson et al., 1987; 

27 El Bayoumy, 1985; Wattenberg, 1974; Field and Roe, 1970; Roe et al., 1970; Biancifiori et al., 1967; 

28 Chouroulinkov et al., 1967; Berenblum and Haran, 1955). Of the controlled, multiple dose-group, 

29 repeat-dosing studies that remain, most treated animals for <1 year, which is less optimal for 

30 extrapolating to a lifetime exposure (Weyand et al., 1995; Triolo et al., 1977; Fedorenko and 

31 Yansheva, 1967; Neal and Rigdon, 1967). 

32 Brune et al. (1981) dosed rats (32/sex/group) with several concentrations of 

33 benzo[a]pyrene dissolved in a 1.5% caffeine solution, sometimes as infrequently as once every 

34 ninth day, for up to 2 years and observed increased forestomach tumors. This study was not 

35 selected for quantitation due to the nonstandard treatment protocol in comparison to the studies 

36 conducted by Kroese et al. (2001) and Beland and Culp (1998).
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1 The Kroese et al. (2001) and Beland and Culp (1998) studies were selected as the best 

2 available studies for dose-response analysis and extrapolation to lifetime cancer risk following oral 

3 exposure to benzo[a]pyrene. The rat bioassay by Kroese et al. (2001) and the mouse bioassay by 

4 Beland and Culp (1998) were conducted in accordance with Good Laboratory Practice principles as 

5 established by the Organization for Economic Co-operation and Development (OECD). These 

6 studies included histological examinations for tumors in many different tissues, contained three 

7 exposure levels and controls, contained adequate numbers of animals per dose group 

8 (~50/sex/group), treated animals for up to 2 years, and included detailed reporting of methods 

9 and results (including individual animal data).

10 Details of the rat (Kroese et al., 2001) and female mouse (Beland and Culp, 1998) study 

11 designs are provided in Appendix B of the Supplemental Information. Dose-related, statistically 

12 significant increasing trends in tumors were noted at the following sites:

13 Squamous cell carcinomas (SCCs) or papillomas of the forestomach or oral cavity in male 

14 and female rats;

15 SCCs or papillomas of the forestomach, tongue, larynx, or esophagus in female mice;

16 Auditory canal carcinomas in male and female rats;

17 Kidney urothelial carcinomas in male rats;

18 Jejunum/duodenum adenocarcinomas in female and male rats;

19 Hepatocellular adenomas or carcinomas in male and female rats; and

20 SCCs or basal cell tumors of the skin or mammary gland in male rats.

21 These tumors were generally observed earlier during the study with increasing exposure 

22 levels, and showed statistically significantly increasing trends in incidence with increasing 

23 exposure level (Cochran-Armitage trend test, p ≤ 0.001). These data are summarized in Appendix C 

24 of the Supplemental Information. As recommended by NTP (McConnell et al., 1986) and as outlined 

25 in EPA’s Cancer Guidelines (U.S. EPA, 2005a), etiologically similar tumor types (i.e., benign and 

26 malignant tumors of the same cell type) were combined for these tabulations when it was judged 

27 that the benign tumors could progress to the malignant form,. In addition, when one tumor type 

28 occurred across several functionally related tissues, as with squamous cell tumors in the tongue, 

29 esophagus, larynx and forestomach, or adenocarcinomas of the jejunum or duodenum, these 

30 incidences were also aggregated as counts of tumor-bearing animals.

31 In the rat study (Kroese et al., 2001), the oral cavity and auditory canal were examined 

32 histologically only if a lesion or tumor was observed grossly at necropsy. Consequently, dose33 response analysis for these sites was not straightforward. Use of the number of tissues examined 

34 histologically as the number at risk would tend to overestimate the incidence, because the 

35 unexamined animals were much less likely to have a tumor. On the other hand, use of all animals in 

36 a group as the number at risk would tend to underestimate if any of the unexamined animals had 

37 tumors that could only be detected microscopically. The oral cavity squamous cell tumors were 
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1 combined with those in the forestomach because both are part of the alimentary tract, recognizing 

2 that there was some potential for underestimating this cancer risk.

3 The auditory canal tumors from the rat study were not considered for dose-response

4 analysis, for several reasons. First, the control and lower dose groups were not thoroughly 

5 examined, similar to the situation described above for oral cavity tumors. Unlike the oral cavity 

6 tumors, the auditory canal tumors were not clearly related to any other site or tumor type, as they 

7 were described as a mixture of squamous and sebaceous cells derived from pilosebaceous units. 

8 The tumors were observed mainly in the high dose groups and were highly coincident with the oral 

9 cavity and forestomach tumors. Because only one mid-dose male had an auditory canal tumor 

10 which did not also have a forestomach or oral cavity squamous cell tumor, and none were observed 

11 in low-dose male or female rats, the data are insufficient to conclude whether the auditory canal 

12 tumors occur independently of other tumors . The investigators did not suggest that these tumors 

13 were metastases from other sites (in which case the auditory canal tumors would be repetitions of 

14 other tumors, or statistically dependent). Therefore dose-response analysis was not pursued for 

15 this site, either separately or in combination with another tumor type.

16 2.3.2. Dose Response Analysis – Adjustments and Extrapolations Methods

17 EPA’s Cancer Guidelines (U.S. EPA, 2005a) recommend that the method used to characterize 

18 and quantify cancer risk from a chemical is determined by what is known about the mode of action 

19 of the carcinogen and the shape of the cancer dose-response curve. The dose response is assumed 

20 to be linear in the low-dose range, when evidence supports a mutagenic mode of action because of 

21 DNA reactivity, or if another mode of action that is anticipated to be linear is applicable. In this 

22 assessment, EPA concluded that benzo[a]pyrene carcinogenicity involves a mutagenic mode of 

23 action (as discussed in Section 1.1.5.). Thus, a linear approach to low-dose extrapolation was used.

24 The high-dose groups of both the rat and mouse studies were dead or moribund by week 79 

25 for female mice, week 72 for female rats, and week 76 for male rats. Due to the occurrence of 

26 multiple tumor types, earlier occurrence with increasing exposure, and early termination of the 

27 high-dose group in each study, methods that can reflect the influence of competing risks and 

28 intercurrent mortality on site-specific tumor incidence rates are preferred. In this case EPA has 

29 used the multistage-Weibull model which incorporates the time at which death-with-tumor 

30 occurred as well as the dose. 

31 Adjustments for approximating human equivalent slope factors applicable for continuous 

32 exposure were applied prior to dose-response modeling. First, continuous daily exposure for the 

33 gavage study in rats (Kroese et al., 2001) was estimated by multiplying each administered dose by 

34 (5 days)/(7 days) = 0.71, under the assumption of equal cumulative exposure yielding equivalent 

35 outcomes. Dosing was continuous in the mouse diet study (Beland and Culp, 1998), so no 

36 continuous adjustment was necessary. Next, consistent with the EPA’s Cancer Guidelines (U.S. EPA, 

37 2005a), an adjustment for cross-species scaling was applied to address toxicological equivalence 
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1 across species. Following EPA’s cross-species scaling methodology, the time-weighted daily 

average doses were converted to human equivalent doses (HEDs) on the basis of (body weight) 2 3/4

3 (U.S. EPA, 1992). This was accomplished by multiplying administered doses by (animal body 

weight (kg)/70 kg)0.25 4 (U.S. EPA, 1992), where the animal body weights were TWAs from each 

5 group, and the U.S. EPA (1988) reference body weight for humans is 70 kg. It was not necessary to 

6 adjust the administered doses for lifetime equivalent exposure prior to modeling for the groups 

7 terminated early, because the multistage-Weibull model characterizes the tumor incidence as a 

8 function of time, from which it provides an extrapolation to lifetime exposure.

9 Details of the modeling can be found in Appendix C of the Supplemental Information. PODs 

10 for estimating low-dose risk were identified at doses at the lower end of the observed data, 

11 generally corresponding to 10% extra risk. The lifetime oral cancer slope factor for humans is 

12 defined as the slope of the line from the lower 95% bound on the exposure at the POD to the control 

13 response (slope factor = 0.1/BMDL10). This slope, a 95% upper confidence limit (UCL) represents a 

14 plausible upper bound on the true risk.

15 2.3.3. Derivation of the Oral Slope Factor 

16 The PODs estimated for each tumor site are summarized in Table 2-7. Details of the model 

17 selection process are provided in Appendix C of the Supplemental Information. Using linear 

18 extrapolation from the BMDL10, human equivalent oral slope factors were derived for each 

19 gender/tumor site combination and are listed in Table 2-7.
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Table 2-7. Summary of the Oral Slope Factor Derivations

Tumor

Species/

Sex

Selected 

Model BMR

BMD

(mg/kg-d)

POD= 

BMDL

(mg/kg-d)

Slope factora

(mg/kg-d)-1

Forestomach, oral cavity: 

squamous cell tumors

Kroese et al., 2001

Male 

Wistar rats

Multistage 

Weibull

10% 0.453 0.281 0.4

0.5b

Hepatocellular adenomas or 

carcinomas

Kroese et al., 2001

Male 

Wistar rats

Multistage 

Weibull

10% 0.651 0.449 0.2

Jejunum/duodenum 

adenocarcinomas

Kroese et al., 2001

Male 

Wistar rats

Multistage 

Weibull

10% 3.03 2.38 0.04

Kidney: urothelial carcinomas

Kroese et al., 2001

Male 

Wistar rats

Multistage 

Weibull

10% 4.65 2.50 0.04

Skin, mammary:

 Basal cell tumors

 Squamous cell tumors

Kroese et al., 2001

Male 

Wistar rats

Multistage 

Weibull

10%

2.86

2.64

2.35

1.77

0.04

0.06

Forestomach, oral cavity: 

squamous cell tumors

Kroese et al., 2001

Female 

Wistar rats

Multistage 

Weibull

10% 0.539 0.328 0.3

0.3b

Hepatocellular adenomas or 

carcinomas

Kroese et al., 2001

Female 

Wistar rats

Multistage 

Weibull

10% 0.575 0.507 0.2

Jejunum/duodenum 

adenocarcinomas

Kroese et al., 2001

Female 

Wistar rats

Multistage 

Weibull

10% 3.43 1.95 0.05

Forestomach, esophagus, 

tongue, larynx: squamous cell 

tumors

Beland and Culp 1998

Female 

B6C3F1

Mice

Multistage 

Weibull

10%

0.127 0.071 1 1

aHuman equivalent slope factor = 0.1/BMDL10HED; see Appendix C of the Supplemental Information for 

details of modeling results.

b Estimates of risk of incurring at least one of the tumor types listed.

1

2 Oral slope factors derived from rat bioassay data varied by gender and tumor site (Table 2-

3 7). Values ranged from 0.04 per mg/kg-day, based on kidney tumors in males, to 0.4 per mg/kg4 day, based on alimentary tract tumors in males. Slope factors based on liver tumors in male and 

5 female rats (0.2 per mg/kg-day) were only slightly lower than slope factors based on alimentary 

6 tract tumors (0.2-0.3 per mg/kg-day). The oral slope factor for alimentary tract tumors in female 

7 mice was highest at 1 per mg/kg-day (Table 2-7), which was approximately fourfold higher than 

8 the oral slope factor derived from the alimentary tract tumors in male rats.
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1 Although the time-to-tumor modeling helps to account for competing risks associated with 

2 decreased survival times and other causes of death including other tumors, considering the tumor 

3 sites individually still does not convey the total amount of risk potentially arising from the 

4 sensitivity of multiple sites—that is, the risk of developing any combination of the increased tumor 

5 types, not just the risk of developing all simultaneously. A method involving the assumption that 

6 the variability in the slope factors could be characterized by a normal distribution is detailed in 

7 Appendix C of the Supplemental Information. The resulting composite slope factor for all tumor 

8 types for male rats was 0.5 per mg/kg-day, about 25% higher than the slope factor based on the 

9 most sensitive tumor site, oral cavity and forestomach, while for female rats, the composite slope 

10 factor was equivalent to that for the most sensitive site (Table 2-7; see Appendix C of Supplemental 

11 Information for composite slope factor estimates).

12 The overall risk estimates from rats and mice spanned about a threefold range. As there are 

13 no data to support any one result as most relevant for extrapolating to humans, the most sensitive 

14 result was used to derive the oral slope factor. The recommended slope factor for assessing human 

15 cancer risk associated with chronic oral exposure to benzo[a]pyrene is 1 per mg/kg-day, based on 

16 the alimentary tract tumor response in female B6C3F1 mice.

17 2.3.4. Uncertainties in the Derivation of the Oral Slope Factor

18 The oral slope factor for benzo[a]pyrene was based on the increased incidence of 

19 alimentary tract tumors, including forestomach tumors, observed in a lifetime dietary study in mice 

20 (Beland and Culp 1998). EPA has considered the uncertainty associated with the relevance of 

21 forestomach tumors for projecting human risk from benzo[a]pyrene exposure. The rodent 

22 forestomach serves to store foods and liquids for several hours before contents continue to the 

23 stomach for further digestion (Clayson et al., 1990; Grice et al., 1986). While humans do not have a 

24 forestomach, squamous epithelial tissue similar to that seen in the rodent forestomach exists in the 

25 oral cavity and upper two-thirds of the esophagus in humans (IARC, 1999). However, due to the 

26 storage function of the forestomach, tissue of the forestomach may be exposed to benzo[a]pyrene 

27 for longer durations than analogous human tissues in the oral cavity and esophagus. This suggests 

28 that the rodent forestomach may be quantitatively more sensitive to development of squamous 

29 epithelial tumors in the forestomach compared to oral or esophageal tumors in humans. 

30 Human studies, specifically associating exposure to benzo[a]pyrene with the alimentary 

31 tract tumors are not currently available. However, benzo[a]pyrene-DNA adducts have been 

32 detected in oral and esophageal tissue obtained from smokers (reviewed by Phillips, 2002) and 

33 several epidemiological studies have identified increased exposure to PAHs as an independent risk 

34 factor for esophageal cancer (Abedi-Ardekani et al., 2010; Szymanska et al., 2010; Wang et al., 2002; 

35 Gustavsson et al., 1998a; Liu et al., 1997). 

36 Uncertainty in the magnitude of the recommended oral slope factor is reflected to some 

37 extent in the range of slope factors among tumors sites and species; the lowest and highest slope 
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1 factors by individual sites, as listed in Table 2-8, show about a 35-fold difference. However, 

2 consideration of individual sites when multiple sites are affected perhaps overemphasizes an 

3 expectation of site concordance. While the highest risk estimates were derived from the incidence 

4 data for forestomach tumors in both rats and mice, the oral slope factor based on the mouse 

5 alimentary tract data was about threefold higher than the overall oral slope factor based on male 

6 rat data (Table 2-8). These comparisons show that the selection of target organ, animal species, 

7 and interspecies extrapolation can impact the oral cancer risk estimate. However, all of the 

8 activation pathways implicated in benzo[a]pyrene carcinogenicity have been observed in human 

9 tissues, and associations have been made between the spectra of mutations in tumor tissues from 

10 benzo[a]pyrene-exposed animals and humans exposed to complex PAH mixtures containing 

11 benzo[a]pyrene (see Section 1.1.5.).

12 Table 2-8. Summary of uncertainties in the benzo[a]pyrene oral cancer slope 

13 factor (OSF) 

Consideration

Impact on unit 

risk Decision Justification

Target organ ↓ OSF, up to 

fivefold, if 

alimentary tract 

tumors not selected

Alimentary tract tumors 

(forestomach, 

esophagus, tongue, 

larynx)

Tumor site is concordant across rats and mice, 

increasing support for its relevance to humans. 

As there are no data to support any one result as 

most relevant for extrapolating to humans, the 

most sensitive result for alimentary tract tumors 

was used to derive the oral slope factor.

Data set ↓ OSF ~threefold if 

rat bioassay were 

selected for OSF 

derivation

Beland and Culp (1998) Beland and Culp (1998) was a well conducted 

study and used the lowest HEDs of the available 

cancer bioassays, reducing low-dose 

extrapolation uncertainty.

Dose metric Alternatives could 

↓ or ↑ slope factor

Administered dose Experimental evidence supports a role for 

metabolism in toxicity, but actual responsible 

metabolites have not been identified. 

Cross-species 

scaling 

Alternatives could 

↓ or ↑ slope factor 

(e.g., 3.5-fold ↓ 

[scaling by body 

weight] or ↑ 2-fold 

[scaling by BW 2/3]) 

BW3/4 scaling (default 

approach)

There are no data to support alternatives. 

Because the dose metric was not an area under 

the curve, BW3/4 scaling was used to calculate 

equivalent cumulative exposures for estimating 

equivalent human risks. While the true human 

correspondence is unknown, this overall 

approach is expected neither to over- or 

underestimate human equivalent risks. 

Dose-response 

modeling 

Alternatives could 

↓ or ↑ slope factor 

Multistage-Weibull 

model 

No biologically based models for benzo[a]pyrene

were available. Because the multistage-Weibull 

model could address additional available data 

(time of death with tumor, and whether a tumor 

caused the death of the animal), this model was 

superior to other available models.

Low-dose 

extrapolation 

↓ cancer risk 

estimate would be 

expected with the 

application of 

Linear extrapolation 

from POD (based on 

mutagenic mode of 

action)

Available mode of action data support linearity 

(mutagenicity is a primary mode of action of 

benzo[a]pyrene); 
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Consideration

Impact on unit 

risk Decision Justification

nonlinear low-dose 

extrapolation

Statistical 

uncertainty at 

POD

↓ OSF 1.8-fold if 

BMD used as the 

POD rather than 

BMDL

BMDL (preferred 

approach for calculating 

plausible upper bound 

slope factor) 

Limited size of bioassay results in sampling 

variability; lower bound is 95% confidence 

interval on administered exposure at 10% extra 

risk of alimentary tract tumors. 

Sensitive 

subpopulations

↑ OSF to unknown 

extent

ADAFs are 

recommended for early 

life exposures

No chemical-specific data are available to 

determine the range of human toxicodynamic 

variability or sensitivity. 

1 2.3.5. Previous Oral Slope Factor

2 The previous cancer assessment for benzo[a]pyrene was posted on the IRIS database in 

3 1987. At that time, benzo[a]pyrene was classified as a probable human carcinogen (Group B2) 

4 based on inadequate data in humans and sufficient data in animals via several routes of exposure. 

5 An oral slope factor was derived from the geometric mean of four slope factor estimates based on 

6 studies in Sprague-Dawley rats (Brune et al., 1981) and CFW-Swiss mice (Neal and Rigdon, 1967). 

7 A single slope factor estimate of 11.7 per mg/kg-day, using a linearized multistage procedure 

8 applied to the combined incidence of forestomach, esophageal, and laryngeal tumors, was derived 

9 from the Brune et al. (1981) study (see Section 1.1.5.2 for study details). Three modeling 

10 procedures were used to derive risk estimates from the Neal and Rigdon (1967) bioassay (see 

11 Section 1.1.5.2). In a report commissioned by EPA, Clement International Corporation (1990) fit a 

12 two-stage response model, based on exposure-dependent changes in both transition rates and 

13 growth rates of preneoplastic cells, to derive a value of 5.9 per mg/kg-day. EPA (1991b) derived a 

14 value of 9.0 per mg/kg-day by linear extrapolation from the 10% response point to the background 

15 response in a re-analysis of the Clement model. Finally, using a Weibull-type model to reflect less16 than-lifetime exposure to benzo[a]pyrene, the same assessment (U.S. EPA, 1991b) derived an 

17 upper-bound slope factor estimate of 4.5 per mg/kg-day. The four slope factor estimates were 

18 within threefold of each other and were judged to be of equal merit. Consequently, the geometric 

19 mean of these four estimates, 7.3 per mg/kg-day, was recommended as the oral slope factor.

20 2.4. Inhalation Unit Risk for Cancer

21 2.4.1. Analysis of Carcinogenicity Data

22 The inhalation database demonstrating carcinogenicity of benzo[a]pyrene consists of a 

23 lifetime inhalation bioassay (Thyssen et al. 1981) and intratracheal instillation studies in hamsters

24 (Feron and Kruysse, 1978; Ketkar et al., 1978; Feron et al., 1973; Henry et al., 1973; Saffiotti et al., 

25 1972). 

26 The bioassay by Thyssen et al. (1981) represents the only lifetime inhalation cancer 

27 bioassay available for describing dose-response relationships for cancer from inhaled 
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1 benzo[a]pyrene. Strengths of the study include the following: 1) exposures was conductd for 2 

2 years; 2) histological examination of organs was extensive and adequate; 3) multiple exposure 

3 groups were used; and 4) individual animal pathology reports with time of death and tumor 

4 detection data were available. Increased incidences of benign and malignant tumors of the larynx

5 and pharynx were seen with increasing exposure concentration. Survival was decreased relative to 

6 control only in the high-dose exposure group; mean survival times in the 0, 2, and 10 mg/m3

7 concentration groups were 96.4, 95.2, and 96.4 weeks, respectively, and 59.5 weeks in the 50 

8 mg/m3 group animals. Overall, tumors occurred earlier in the highest benzo[a]pyrene exposure 

9 group than in the mid-exposure group. 

10 Limitations of the study include the following: (1) only male animals were tested; (2) 

11 particle analysis of aerosols was not reported (i.e., mass median aerodynamic diameter and 

12 geometric SD were not reported); and (3) benzo[a]pyrene exposure occurred through the 

13 inhalation of hygroscopic particles (benzo[a]pyrene was adsorbed onto sodium chloride aerosols), 

14 which may have a different deposition than benzo[a]pyrene adsorbed onto carbonaceous non15 hygroscopic particles as is more typical in the environment. In addition, reported information 

16 regarding the actual exposure levels achieved during the study was incomplete. Analytic 

17 measurements varied more than 20% from the average achieved for each dose group during the 

18 first 79 weeks of the study, but were not available for the remaining 52 weeks that it took to

19 complete the experiment. Despite the limitations, the strengths of the study were judged to support 

20 use of the data to derive an inhalation unit risk for benzo[a]pyrene.

21 The intratracheal instillation studies provide supporting evidence of carcinogenicity of 

22 inhaled benzo[a]pyrene; however, the use of intratracheal dosing alters the deposition, clearance, 

23 and retention of substances, and therefore studies utilizing this exposure technique are not as 

24 useful for the quantitative extrapolation of cancer risk from the inhalation of benzo[a]pyrene in the 

25 environment (Driscoll et al., 2000).

26 2.4.2. Dose Response Analysis—Adjustments and Extrapolations Methods

27 Biologically based dose-response models for benzo[a]pyene are not available. A simplified 

28 version of the two-stage carcinogenesis model proposed by Moolgavkar and Venzon (1979) and 

29 Moolgavkar and Knudson (1981) has been applied to the Thyssen et al. (1981) data (U.S. EPA, 

30 1990). However, the simplifications necessary to fit the tumor incidence data reduced that model 

31 to an empirical model, i.e., there were no biological data to inform estimates of cell proliferation 

32 rates for background or initiated cells. There were sufficient data to apply a time-to-tumor dose33 response model, described in detail in Appendix C of the Supplemental Information.

34 The tumor incidence data used for dose-response modeling comprised the benign and 

35 malignant tumors in the pharynx and larynx. The pharynx and larynx are associated with the upper 

36 digestive tract and the upper respiratory tract, respectively. However, these sites are close 

37 anatomically and in some cases where both tissues were affected, the site of origin could not be 
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1 distinguished (U.S. EPA, 1990). In addition, the benign tumors (e.g., papillomas, polyps and 

2 papillary polyps) were considered early stages of the squamous cell carcinomas in these tissues

3 (U.S. EPA, 1990). Following EPA’s Cancer Guidelines (Section 2.2.2.1.2; U.S. EPA, 2005a), incidence 

4 data for animals with malignant or benign tumors originating from the same cell type were selected 

5 for dose-response modeling based on the assumption that the benign tumors could develop into 

6 malignancies.

7 The availability of the raw chamber air monitoring data and individual times on study 

8 allowed the calculation of TWA continuous exposures for each hamster (U.S. EPA, 1990). Group 

averages of these concentrations were 0, 0.25, 1.01, and 4.29 mg/m3 9 , respectively, for the 0, 2, 10, 

and 50 mg/m3 10 study concentrations.

11 A toxicokinetic model to assist in cross-species scaling of benzo[a]pyrene inhalation 

12 exposure was not available. In addition, default dosimetry adjustments utilized in the 

13 benzo[a]pyrene RfC calculation could not be applied because aerosol particle distribution data 

14 were not available for the hamster inhalation bioassay by Thyssen et al. (1981). The carrier 

15 particle used in Thyssen et al. (1981) was sodium chloride, a soluble hygroscopic particle, and the 

16 approaches presented in the RfC methodology guidelines (U.S. EPA 1994b) were developed for 

17 insoluble and nonhygroscopic particles. Consequently, without data to inform a basis for 

18 extrapolation to humans, it was assumed that equal risk for all species would be associated with 

19 equal concentrations. This is equivalent to assuming that any metabolism of benzo[a]pyene is 

20 directly proportional to breathing rate and that the deposition rate is equal between species. 

21 A time-to-tumor dose-response model was fit to the TWA exposure concentrations and the 

22 individual animal tumor and survival data for tumors in the larynx, pharynx, trachea, esophagus, 

23 and forestomach, using the computer software program MSW (U.S. EPA, 2010) as described in 

24 Appendix C of the Supplemental Information. Unlike in the available oral bioassays, Thyssen et al.

25 did not determine cause of death for any of the animals. Since the investigators for the oral 

26 bioassays considered the same tumors to be fatal at least some of the time, bounding estimates for 

27 the Thyssen et al. (1981) data were developed by treating the tumors alternately as either all 

28 incidental or all fatal. Modeling results are provided in Appendix C of the Supplemental 

29 Information.

30 2.4.3. Derivation of the Inhalation Unit Risk 

31 Because benzo[a]pyrene carcinogenicity involves a mutagenic mode of action, linear low 

32 dose extrapolation from the BMCL10 was used (U.S. EPA, 2005a) to derive the inhalation unit risk. 

33 BMCs and BMCLs associated with an extra risk of 10% calculated using the multistage-Weibull 

34 model, based on the occurrence of upper respiratory and upper digestive tract tumors in male 

35 hamsters exposed to aerosols of benzo[a]pyrene for up to 132 weeks, were estimated under two 

36 bounding assumptions. The results are summarized in Table 2-9. At one extreme, taking the 

37 tumors to have been the cause of death of the experimental animals, the BMC10 and BMCL10 were 
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0.648 and 0.461 mg/m3 1 , respectively. Then taking all of the tumors to have been incidental to the 

cause of death for each animal, the BMC10 and BMCL10 were 0.285 and 0.198 mg/m3 2 , respectively, 

3 about twofold lower than the first case. Because the tumors were unlikely to have all been fatal, the 

4 lower BMDL10 was selected for estimating the inhalation unit risk. Using linear extrapolation from 

the BMCL10 of 0.198 mg/m3, an inhalation unit risk of 0.5 per mg/m3 5 , or 5 × 10-4 per μg/m3

6 (rounding to one significant digit), was calculated.

Table 2-9. Summary of the Inhalation Unit Risk Derivation

Tumor Site and Context

Species/

Sex

Selected 

Model BMR

BMC

(mg/m3)

POD= 

BMCL

(mg/m3)

Unit Riska

(mg/m3)

-1

Upper respiratory and digestive 

tracts; all treated as cause of death

Thyssen et al., 1981

Male 

Hamsters

Multistage 

Weibull 10% 0.648 0.461 0.22

Upper respiratory and digestive 

tracts; all treated as incidental to 

death

Thyssen et al., 1981

Male 

Hamsters

Multistage 

Weibull 10% 0.285 0.198 0.51

a Human equivalent unit risk = 0.10/BMCL10; see Appendix C for details of modeling results.

7 2.4.4. Uncertainties in the Derivation of the Inhalation Unit Risk

8 Only one animal cancer bioassay by the inhalation route is available that describes the dose9 response relationship for respiratory tract tumors with chronic inhalation exposure to 

10 benzo[a]pyrene (Thyssen et al., 1981). Although corroborative information on dose-response 

11 relationships in other animal species is lacking, the findings for upper respiratory tract tumors are 

12 consistent with findings in other hamster studies with intratracheal administration of 

13 benzo[a]pyrene, and with some of the portal of entry effects in oral exposure studies. This study is 

14 adequate for dose-response analysis and derivation of an inhalation unit risk estimate, but some 

15 associated uncertainty includes the inability to apply U.S. EPA (1994b) dosimetry approaches to 

16 extrapolate inhaled doses from animals to humans, due to the use of a soluble hygroscopic carrier 

17 particle (sodium chloride) for the delivery of benzo[a]pyrene. One likely consequence of the use of 

18 hygroscopic carrier particles would be the growth of benzo[a]pyrene-sodium chloride particles in 

19 the humid environment of the respiratory tract resulting in increased particle diameter and 

20 resulting changes in particle deposition, specifically, increased impaction in the upper respiratory 

21 tract (Varghese and Gangamma, 2009; Asgharian, 2004; Ferron, 1994; Xu and Yu, 1985). Exposure 

22 to benzo[a]pyrene in the environment predominantly occurs via non-soluble, non-hygroscopic, 

23 carbonaceous particles (such as soot and diesel exhaust particles). The potential impact of 

24 differences in carrier particle on the magnitude of the inhalation unit risk is unknown.

25 The exposure measures entailed some uncertainty, in that the exposure concentrations for 

26 about half of the study period were documented to vary above and below the average achieved by 
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1 about twofold in all exposed groups. One possibility that cannot be ruled out, due to lack of data, is 

2 that peak exposure above some concentration may be associated with the observed effects. Use of 

3 Haber’s Law (equal cumulative exposures being expected to lead to similar outcomes) to estimate 

4 continuous equivalent exposures may be not justified. There is not enough information available to 

5 estimate an impact on the estimated unit risk due to this uncertainty.

6 Table 2-10. Summary of uncertainties in the benzo[a]pyrene cancer inhalation

7 unit risk (IUR)

Consideration

Impact on unit 

risk Decision Justification

Data set and 

target organ

No IUR if Thyssen 

et al. (1981) not 

used

Respiratory tract 

tumors from Thyssen et 

al. (1981)

The Thyssen et al. (1981) bioassay is the only 

lifetime inhalation cancer bioassay available for 

describing dose-response relationships for 

cancer from inhaled benzo[a]pyrene. 

Intratracheal implantation study supports the 

association of benzo[a]pyrene exposure with 

respiratory tract tumors. Oral exposure 

resulted in similar tumors.

Dose metric Alternatives could 

↓ or ↑ unit risk

Administered dose as 

time-weighted average

Experimental evidence supports a role for 

metabolism in toxicity, but actual responsible 

metabolites are not identified. 

Cross-species 

scaling 

Alternatives could 

↓ or ↑ slope 

factor 

The carcinogenicity 

observed was portal-ofentry, but data to 

implement RfC 

methodology (EPA, 

1994) were lacking. 

Cross-species scaling 

was not applied.

There are no data to support alternatives. 

Equal risk per μg/m3 is assumed.

Dose-response 

modeling 

Alternatives could 

↓ or ↑ slope 

factor 

Multistage-Weibull 

model 

No biologically based models for 

benzo[a]pyrene were available. Because the 

multistage-Weibull model could address 

additional available data (time of death with 

tumor), this model was superior to other 

available empirical models

Low-dose 

extrapolation 

↓ cancer risk 

estimate would 

be expected with 

the application of 

nonlinear lowdose 

extrapolation

Linear extrapolation 

from the point of 

departure (based on 

mutagenic mode of 

action)

Available mode of action data support linearity 

(mutagenicity is a primary mode of action of 

benzo[a]pyrene); 

Statistical 

uncertainty at 

POD

↓ IUR 1.4-fold if 

BMC used as the 

POD rather than 

BMCL

BMCL (preferred 

approach for 

calculating plausible 

upper bound unit risk) 

Limited size of bioassay results in sampling 

variability; lower bound is 95% confidence 

interval on administered exposure at 10% extra 

risk of respiratory tract tumors. 
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Sensitive 

subpopulations

↑ IUR to 

unknown extent

ADAFs are 

recommended for early 

life exposures

No chemical-specific data are available to 

determine the range of human toxicodynamic 

variability or sensitivity. 

1 2.4.5. Previous Inhalation Unit Risk

2 An inhalation unit risk for benzo[a]pyrene was not previously available on IRIS.

3 2.5. Dermal Slope Factor for Cancer

4 Evidence in humans and animal studies demonstrates an increased incidence of skin tumors 

5 with increasing dermal exposure to polycyclic aromatic hydrocarbons (PAHs) mixtures or to 

6 benzo[a]pyrene alone. Thus, this assessment for benzo[a]pyrene derives a dermal slope factor, a 

7 quantitative risk estimate that is a plausible upper bound on the estimate of risk per μg/day of 

8 dermal exposure. This is the first derivation of a dermal slope factor for the IRIS database.

9 2.5.1. Analysis of Carcinogenicity Data

10 Skin cancer in humans has been documented to result from occupational exposure to 

11 complex mixtures of PAHs including benzo[a]pyrene, such as coal tar pitches, non-refined mineral 

12 oils, shale oils, and soot (IARC, 2010; Baan et al., 2009; WHO, 1998; Boffetta et al., 1997; ATSDR, 

13 1995); however, studies of human exposure to benzo[a]pyrene alone are not available. Repeated 

14 application of benzo[a]pyrene to skin (in the absence of exogenous promoters) has been 

15 demonstrated to induce skin tumors in mice, rats, rabbits, and guinea pigs. However, no lifetime 

16 chronic bioassays of dermal benzo[a]pyrene exposure were located in these species. Therefore,

17 this analysis focuses on chronic carcinogenicity bioassays in several strains of mice demonstrating 

18 increasing incidence of benign and malignant skin tumors, as well as earlier occurrence of tumors 

19 with increasing exposure, following repeated dermal exposure to benzo[a]pyrene for the animals’ 

20 lifetime. These studies involved 2- or 3-times/week exposure protocols, at least two exposure 

21 levels plus controls, and histopathological examinations of the skin and other tissues (Sivak et al., 

22 1997; Grimmer et al., 1984, 1983; Habs et al., 1984, 1980; Schmähl et al., 1977; Schmidt et al., 1973; 

23 Roe et al., 1970; Poel, 1960, 1959) (see Tables B-15 to B-23 in the Supplemental Information).

24 Additional carcinogenicity studies in mice were considered, but not used, in the dose25 response analysis. These studies included: (1) early “skin painting” studies of benzo[a]pyrene 

26 carcinogenicity in mouse skin that did not report sufficient information to estimate the doses 

27 applied (e.g., Wynder and Hoffman, 1959; Wynder et al., 1957); (2) initiation-promotion studies 

28 utilizing acute dosing of benzo[a]pyrene followed by repeated exposure to a potent tumor 

29 promoter (sometimes benzo[a]pyrene at a lower dose than the initiation step), because they are 

30 not as relevant for calculating risks from constant benzo[a]pyrene exposure alone; (3) bioassays 

31 with one benzo[a]pyrene dose level or with only dose levels inducing 90–100% incidence of mice 

32 with tumors, because they provide relatively little information about the shape of the dose33 response relationship (e.g., Wilson and Holland, 1988); (4) studies with shorter exposure and 
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1 observation periods (i.e., <1 year) (Higginbotham et al., 1993; Albert et al., 1991; Nesnow et al., 

2 1983; Emmett et al., 1981; Levin et al., 1977) which are less relevant for characterizing lifetime 

3 risk; and (5) studies involving vehicles expected to interact with or enhance benzo[a]pyrene 

4 carcinogenicity (e.g., Bingham and Falk, 1969), which precludes assessment of carcinogenic risks of 

5 benzo[a]pyrene alone.

6 Study designs and the extent of data reported varied across the studies identified above. 

7 Because no particular studies stood out as superior for developing a dermal slope factor, these data 

8 sets were considered as a group in order to assess the overall evidence. Each data set was examined 

9 for limitations that would not support dose-response evaluation. Only the data from Poel (1960) 

10 were not considered further for modeling. The three data sets reported by Poel (1960) 

11 demonstrated high mortality and 100% tumor incidence at doses higher than those in any of the 

12 other studies under consideration, used the smallest group sizes of the available studies, and did 

13 not provide sufficient information to estimate the duration of exposure for the dose groups with 

14 less than 100% tumor incidence (see Table B-16 in the Supplemental Information).

15 The remaining studies varied in complete reporting of: incidence of benign and malignant 

16 tumors or only malignancies; times of first tumor occurrence, which informs whether earlier 

17 mortality impacts the number at risk of tumor development; and duration of exposure by dose 

18 group. With the exception of Poel (1959), summarized next, the rest of the studies considered for 

19 dose-response modeling contained groups which survived the majority of a two-year exposure 

20 period. These studies support dose-response modeling, with some limitations that indicate the 

21 slope factors could be underestimates (e.g., possible over-estimation of animalas at risk or exposure 

22 associated with tumor response). The study by Poel (1959) included nine dose groups and also 

23 demonstrated high mortality and tumor incidence at higher exposure levels. All mice in dose 

24 groups with >3.8 μg/application died by week 44 of the study. Therefore, these five dose groups 

25 were omitted prior to dose-response modeling because of the relatively large uncertainty 

26 associated with characterizing lifetime cancer risk from short exposures (<1 year). Four dose 

27 groups in addition to control remained, all with exposures lasting at least 83 weeks. 

28 2.5.2. Dose Response Analysis – Adjustments and Extrapolations Methods

29 As with the oral and inhalation benzo[a]pyrene carcinogenicity data, benzo[a]pyrene’s 

30 dermal exposure carcinogenicity data were generally characterized by earlier occurrence of tumors 

31 and increased mortality with increasing exposure level. However, individual animal data were not 

32 available for any of the identified studies. Therefore, time-to-tumor modeling was not possible. 

33 Each of the dermal data sets was modeled using the multistage model.

34 For all studies, administered doses were converted to average daily doses using the 

35 equation:

36

37 Average dose/day = (μg/application) × (number of exposures/week ÷ 7 days/week)
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1 Next, lifetime equivalent doses were estimated for study groups that were reported to end 

before 104 weeks by multiplying the relevant average daily doses by (Le/104)3 2 , where Le is the 

3 length of exposure, based on observations that tumor incidence tends to increase with age (Doll, 

4 1971). Note that exposure periods <52 weeks would lead to a relatively large adjustment [i.e., 

(52/104) 5 3 = 0.125, or an eightfold lower dose than administered], reflecting considerable 

6 uncertainty in lifetime equivalent dose estimates generated from relatively short studies. This 

7 adjustment was relevant for Poel (1959), Roe et al. (1970), Habs et al. (1980), and Sivak et al. 

8 (1997).

9 Concerning the incidence data, some of these studies reported incidences of skin tumor10 bearing animals for tumors thought to be malignant (Roe et al., 1970; Poel, 1959) or without clear 

11 designation of the relative percentages of animals with carcinomas and papillomas (Habs et al., 

12 1980). In the other studies, incidences of animals with skin papillomas and skin carcinomas were 

13 clearly reported, showing that skin tumors from life-time exposure to benzo[a]pyrene were 

14 predominately malignant (Sivak et al., 1997; Habs et al.,1984; Grimmer et al., 1984, 1983; Schmähl 

15 et al., 1977; Schmidt et al., 1973). Following EPA’s Cancer Guidelines (Section 2.2.2.1.2; U.S. EPA, 

16 2005a), incidence data for animals with malignant or benign skin tumors were selected for dose17 response modeling based on the evidence that skin papillomas can develop into malignant skin 

18 tumors. The data sets as modeled are presented in Tables C-24 through C-27 in the Supplemental

19 Information. 

20 The multistage-cancer model was then fit to each data set. If there was no adequate fit using 

21 the multistage-cancer model, then other dichotomous models were considered. Because 

22 benzo[a]pyrene is expected to cause cancer via a mutagenic mode of action, a linear approach to 

23 low dose extrapolation from the PODs (i.e., BMDL10)was used (U.S. EPA, 2005a) for candidate 

24 dermal slope factors.

25 2.5.3. Derivation of the Dermal Slope Factor 

26 Adequate model fits were found using the multistage model for all but one of the mouse 

27 skin tumor incidence data sets (see Appendix C of the Supplemental Information). The data from 

28 Grimmer et al. (1984) could not be adequately fit by the multistage model initially, and the other 

29 dichotomous models available in BMDS were considered. Due to the supralinear shape of the dose30 response data, only the log-logistic and dichotomous Hill models provided adequate fits. Also due 

31 to the supralinear dose-response shape, the POD for slope factor derivation was identified near the 

32 lowest response of ~70%, because of the lack of data to inform the dose-response relationship at 

33 lower doses.

34 Dermal slope factors, calculated in units of risk per (μg/day) using linear extrapolation from 

35 the BMDL values, ranged from 0.25 to 1.8 per μg/day, a roughly sevenfold range (see Table 2-11). A 

36 number of differences among studies may contribute to this range, including solvent choice, sex and 

37 strain of mice studied, dose ranges, varying group sizes, and level of detail reported. Mouse strains 
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1 were not repeated across sexes among these studies, so it cannot be established whether Swiss or 

2 NRMI mice are more or less sensitive than other strains. In addition, different solvents were used 

3 in the various studies with varying strain and sex combinations tested. For example, toluene was 

4 used in one male study only, and all of the female studies used acetone. Thus, any possible impact 

5 of the solvents used is not clear. As noted earlier, incomplete mortality information in several of 

6 the female mouse studies (Grimmer et al., 1984, 1983; Habs et al., 1984, 1980; Schmähl et al., 1977; 

7 Schmidt et al., 1973) suggests that the dermal slope factors derived from those studies may 

8 underestimate cancer risk. 

9 The estimates derived from the two studies in male mice (Poel, 1959; Sivak et al. 1997) 

10 were at the higher end of the range of slope factors derived. The available information is too 

11 limited to conclude that males are more sensitive than females, in view of the similar result from 

12 the Habs et al. (1984) data and the limitations of the remaining studies that suggested risk might be 

13 underestimated. Nevertheless, the studies in male mice were among the stronger studies. They 

14 included at least three exposure levels, the lowest doses tested, and better reporting of intercurrent 

15 mortality. Both studies yielded very similar BMDs, with Poel (1959) providing a higher BMDL10, 

16 associated with its larger group sizes (~50/dose group). Thus these two results were combined by 

17 estimating the midpoint between the BMDL10s, 0.068 μg/day.

Table 2-11. Summary of Dermal Slope Factor Derivations

Reference

Sex/ strain of 

mouse

Selected 

Modela BMR BMD(μg/d)

BMDL

(μg/d)

Dermal 

slope 

factorb

(μg/d)-1

Sivak et al., 1997 Male C3H/HeJ Multistage 2° 10% 0.11 0.058 1.7

Poel, 1959 Male C57L Multistage 3° 10% 0.12 0.078 1.3

Habs et al., 1984 Female NMRI Multistage 3° 10% 0.078 0.056 1.8

Grimmer et al., 1984 Female CFLP Log-logistic 70% 1.07 0.48 1.5

Schmahl et al., 1977 Female NMRI Multistage 2° 10% 0.23 0.15 0.67

Schmidt et al., 1973 Female Swiss Multistage 3° 10% 0.28 0.22 0.45

Grimmer et al., 1983 Female CFLP Multistage 1° 10% 0.24 0.21 0.48

Habs et al., 1980 Female NMRI Multistage 3° 10% 0.29 0.22 0.45

Schmidt et al., 1973 Female NMRI Multistage 2° 10% 0.33 0.29 0.34

Roe et al., 1970 Female Swiss Multistage 2° 10% 0.69 0.39 0.25

a See Appendix C for details of modeling results.

b Unadjusted for interspecies differences

18

19 The BMDL10 of 0.068 μg/day, based on the tumor responses in C57L male mice (Poel, 1959) 

20 and in C3H/HeJ male mice (Sivak et al., 1997), was chosen for developing a human dermal slope 

21 factor.
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1 2.5.4. Dermal Slope Factor Cross Species Scaling

2 Different methodologies have been established for interspecies scaling of PODs used to 

3 derive oral slope factors and inhalation unit risks. Cross-species adjustment of oral doses is based 

4 on allometric scaling using the 3⁄4 power of body weight. This adjustment accounts for more rapid 

5 distribution, metabolism, and clearance in small animals (U.S. EPA, 2005a). Cross-species 

6 extrapolation of inhalation exposures is based on standard dosimetry models that consider factors 

7 such as solubility, reactivity, and persistence (U.S. EPA, 1994b). No established methodology exists 

8 to adjust for interspecies differences in dermal toxicity at the point of contact; however, allometric 

9 scaling using body weight to the 3⁄4 power was selected based on known species differences in 

10 dermal metabolism and penetration of benzo[a]pyrene. In vitro skin permeation was highest in the 

11 mouse, compared to rat, rabbit and human, and was enhanced by induction of CYP enzymes (Kao et 

12 al., 1985). Using this approach, rodents and humans exposed to the same daily dose of a 

carcinogen, adjusted for BW3/4 13 , would be expected to have equal lifetime risks of cancer.

14 Alternative approaches were also evaluated. A comparison of these alternatives is provided 

15 in Appendix D of the Supplemental Information. 

16 The PODM derived from the mouse studies of Poel (1959) and Sivak et al. (1997) is adjusted 

17 to a HED as follows:

18

POD HED (μg/day) = PODM (μg/day) × (BWH / BWM) 19 3/4

= 0.068 μg/day × (70 kg / 0.035 kg) 20 3/4

21 = 20.3 μg/day

22

23 The resulting PODHED is used to calculate the dermal slope factor for benzo[a]pyrene:

24

Dermal slope factor = 0.1/PODHED = 0.1/(20.3 μg/day) = 0.005 (μg/day) 25 -1

26

27 Note that the dermal slope factor should only be used with lifetime human exposures 

28 <20 μg/day, the human equivalent of the PODM, because above this level, the dose-response 

29 relationship may not be proportional to the mass of benzo[a]pyrene applied.

30 Several assumptions are made in the use of this scaling method. First, it is assumed that the 

31 toxicokinetic processes in the skin will scale similarly to interspecies differences in whole-body 

32 toxicokinetics. Secondly, it is assumed that the risk at low doses of benzo[a]pyrene is linear. 

33 Although one study indicates that at high doses of benzo[a]pyrene carcinogenic potency is related 

34 to mass applied per unit skin and not to total mass (Davies, 1969), this may be due to promotional 

35 effects, such as inflammation, that are observed at high doses of benzo[a]pyrene.

36 The dermal slope factor has been developed for a local effect and it is not intended to 

37 estimate systemic risk of cancer following dermal absorption of benzo[a]pyrene into the systemic 

38 circulation. Although some information suggests that benzo[a]pyrene metabolites can enter 
Toxicological Review of Benzo[a]pyrene

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1 systemic circulation following dermal exposure in humans (Godschalk et al., 1998), lifetime skin 

2 cancer bioassays that have included pathological examination of other organs have not found 

3 elevated incidences of tumors at distal sites (Higginbotham et al., 1993; Habs et al., 1980; Schmahl 

4 et al., 1977; Schmidt et al., 1973; Roe et al., 1970; Poel, 1959). In addition, benzo[a]pyrene tends to 

5 bind to targets within the skin rather than enter the plasma receptor fluid (a surrogate measure of 

6 systemic absorption) in in vitro human skin experiments. These data are consistent with 

7 benzo[a]pyrene’s metabolism to reactive metabolites within the viable layers of the skin (Wester et 

8 al., 1990). Some studies indicate that the fraction of benzo[a]pyrene left within the viable layers of 

9 the skin is a large portion of the applied dose (Moody et al., 2007, 1995). Taken together, these data 

10 support the conclusion that the risk of skin cancer following dermal exposure likely outweighs 

11 cancer risks at distal organs.

12 2.5.5. Uncertainties in the Derivation of the Dermal Slope Factor

13 Uncertainty in the recommended dermal slope factor is partly reflected in the range of POD 

14 values derived from the modeled mouse skin tumor data sets: the lowest and highest BMDL10

15 values listed in Table 2-11 show a seven-fold difference (0.056–0.39 μg/day) in magnitude. There 

16 is some indication that the recommended dermal slope factor may underestimate cancer risk, due 

17 to inadequate data to take the observed decreasing tumor latency with increasing exposure level 

18 into account. Reliance on studies with the lowest exposure levels where early mortality due to 

19 benzo[a]pyrene exposure was low and exposures continued for approximately 104 weeks tends to

20 minimize this source of uncertainty.

21 Human dermal exposure to benzo[a]pyrene in the environment likely occurs predominantly 

22 through soil contact. The available mouse dermal bioassays of benzo[a]pyrene relied on delivery of 

23 benzo[a]pyrene to the skin in a solvent solution (typically acetone or toluene). The use of volatile 

24 solvent likely results in a larger dose of benzo[a]pyrene available for uptake into the skin 

25 (compared to soil). Consequently, reliance on these studies may overestimate the risk of skin 

26 tumors from benzo[a]pyrene contact through soil; however, cancer bioassays delivering 

27 benzo[a]pyrene through a soil matrix are not available.

28 There is uncertainty in extrapolating from the intermittent exposures in the mouse assays 

29 to daily exposure scenarios. All of the dermal bioassays considered treated animals 2-3 times a 

30 week. This assessment makes the assumption that risk is proportional to total cumulative 

31 exposure. However, this may overestimate risk if duration adjusted doses are below doses which 

32 saturate or slow detoxifying metabolic steps. 

33 The available data were not useful to determine which animal species may be the best 

34 surrogate for human dermal response to benzo[a]pyrene. In extrapolation of the animal dermal 

35 information to humans, the assumption is made that equal lifetime risks of cancer would follow 

from exposure to the same daily dose adjusted for BW3/4 36 . Qualitatively, the toxicokinetics and 

37 toxicodynamics in mouse and human skin appear to be similar (Knafla et al., 2011; Bickers et al., 
Toxicological Review of Benzo[a]pyrene

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1 1984). Specifically, all of the activation pathways implicated in benzo[a]pyrene carcinogenicity 

2 have been observed in mouse and human skin, and associations have been made between the 

3 spectra of mutations in tumor tissues from benzo[a]pyrene-exposed animals and humans exposed 

4 to complex PAH mixtures containing benzo[a]pyrene (see Section 1.1.5).

5 The dermal slope factor for benzo[a]pyrene is based on skin cancer and does not represent 

6 systemic cancer risk from dermal exposure. It is unclear whether dermal exposure to 

7 benzo[a]pyrene would result in elevated risk of systemic tumors. Some studies in humans suggest 

8 that although the skin may be responsible for a “first pass” metabolic effect, benzo[a]pyrene9 specific adduct levels have been detected in WBCs following dermal exposure to benzo[a]pyrene, 

10 indicating that dermally applied benzo[a]pyrene enters systemic circulation (Godschalk et al., 

11 1998). Although none of the lifetime dermal bioassays in mice, which included macroscopic 

12 examination of internal organs, reported an elevation of systemic tumors in benzo[a]pyrene13 treated mice compared to controls (Higginbotham et al., 1993; Habs et al., 1980; Schmahl et al., 

14 1977; Schmidt et al., 1973; Roe et al., 1970; Poel, 1959), most of these studies attempted to remove 

15 animals with grossly observed skin tumors from the study before the death of the animal, possibly 

16 minimizing the development of more distant tumors with longer latency. The risk of 

17 benzo[a]pyrene-induced point-of-contact tumors in the skin possibly competes with systemic risk 

18 of tumors. Currently, the potential contribution of dermally absorbed benzo[a]pyrene to systemic 

19 cancer risk is unclear.

20 Table 2-12. Summary of uncertainties in the benzo[a]pyrene cancer dermal 

21 slope factor (DSF)

Consideration

Impact on unit 

risk Decision Justification

Data set ↓ DSF if 

alternative data 

set were selected 

Poel (1959); Sivak et al.

(1997)

Poel (1959) included lowest doses among 

available studies (where intercurrent mortality 

was less likely to impact the number at risk) and 

used adequate group sizes. 

Target organ No DSF if skin 

tumor studies not 

used

Selection of skin 

tumors 

Skin tumors were replicated in numerous 

studies of male or female mice. No studies 

were available indicating that other tumors 

occur following dermal exposure; 

Dose metric Alternatives could 

↓ or ↑ slope 

factor

Administered dose, as 

time-weighted average

in μg/day.

Experimental evidence supports a role for 

metabolism in toxicity, but actual responsible 

metabolites are not identified. 

Cross-species 

scaling 

Alternatives could 

↓ or ↑ slope 

factor 

Total daily dose scaled 

by BW3/4 

Alternatives discussed in Appendix C. An 

established methodology does not exist to 

adjust for interspecies differences in dermal 

toxicity at the point of contact. Benzo[a]pyrene 

metabolism is known to occur in the dermal 

layer. Viewing the skin as an organ, metabolic 

processes were assumed to scale allometrically 

without evidence to the contrary. 
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Consideration

Impact on unit 

risk Decision Justification

Dose-response 

modeling 

Alternatives could 

↓ or ↑ slope 

factor 

Multistage model No biologically based models for 

benzo[a]pyrene were available. The multistage 

model is consistent with biological processes 

and is the preferred model for IRIS cancer 

assessments (Gehlhaus et al., 2011). 

Low-dose 

extrapolation 

↓ cancer risk 

estimate would 

be expected with 

the application of 

nonlinear lowdose 

extrapolation

Linear extrapolation 

from POD (based on 

mutagenic MOA)

Available mode of action data support linearity 

(mutagenicity is a primary mode of action of 

benzo[a]pyrene); 

Sensitive 

subpopulations

↑ DSF to 

unknown extent

ADAFs are 

recommended for early 

life exposures

No chemical-specific data are available to 

determine the range of human toxicodynamic 

variability or sensitivity. 

1 2.5.6. Previous Dermal Slope Factor

2 A dermal slope factor for benzo[a]pyrene was not previously available on IRIS.

3 2.6. Application of Age-dependent Adjustment Factors (ADAFs)

4 Based on sufficient support in laboratory animals and relevance to humans, benzo[a]pyrene 

5 is determined to be carcinogenic by a mutagenic mode of action. According to the Supplemental 

6 Guidance for Assessing Susceptibility from Early Life Exposure to Carcinogens (“Supplemental 

7 Guidance”) (U.S. EPA, 2005b), individuals exposed during early life to carcinogens with a mutagenic 

8 mode of action are assumed to have increased risk for cancer. The oral slope factor of 1 per mg/kgday, inhalation unit risk of 0.5 per mg/m3 9 , and dermal slope factor of 0.005 per μg/day for 

10 benzo[a]pyrene, calculated from data applicable to adult exposures, do not reflect presumed early 

11 life susceptibility to this chemical. Although chemical-specific data exist for benzo[a]pyrene that 

12 quantitatively demonstrate increased early life susceptibility to cancer (Vesselinovitch et al., 1975), 

13 these data were not considered sufficient to develop separate risk estimates for childhood 

14 exposure, as they used acute i.p. exposures (U.S. EPA, 2005b). In the absence of adequate chemical15 specific data to evaluate differences in age-specific susceptibility, the Supplemental Guidance (U.S. 

16 EPA, 2005b) recommends that ADAFs be applied in estimating cancer risk.

17 The Supplemental Guidance (U.S. EPA, 2005b) establishes ADAFs for three specific age 

18 groups. These ADAFs and their corresponding age groupings are: 10 for individuals exposed at <2 

19 years of age, 3 for exposed individuals at 2–<16 years of age, and 1 for exposed individuals ≥16 

20 years of age. The 10- and 3-fold adjustments are combined with age specific exposure estimates 

21 when estimating cancer risks from early life (<16 years of age) exposures to benzo[a]pyrene. To 

22 illustrate the use of the ADAFs established in the Supplemental Guidance (U.S. EPA, 2005b), sample 

23 calculations are presented for three exposure duration scenarios, including full lifetime, assuming a 

24 constant benzo[a]pyrene exposure of 0.001 mg/kg-day (Table 2-13).
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Table 2-13. Application of ADAFs for the estimation of benzo[a]pyrene cancer 

risk following lifetime (70-year) oral exposure

Age group ADAF

Unit risk Exposure concentration

Duration 

adjustment

Cancer risk for 

specific exposure 

duration 

(per mg/kg-d) (mg/kg-d) scenarios

0–<2 yrs 10 1 0.001 2 yrs/70 yrs 0.0003

2–<16 yrs 3 1 0.001 14 yrs/70 yrs 0.0006

≥16 yrs 1 1 0.001 54 yrs/70 yrs 0.0007

Total risk 0.002

1

2 The example exposure duration scenarios include full lifetime exposure (assuming a 70-

3 year lifespan). Table 2-13 lists the four factors (ADAFs, cancer risk estimate, assumed exposure, 

4 and duration adjustment) that are needed to calculate the partial cancer risk based on the early 

5 age-specific group. The cancer risk for each age group is the product of the four factors in columns 

6 2–5. Therefore, the cancer risk following daily benzo[a]pyrene oral exposure in the age group 0–<2 

years is the product of the values in columns 2–5 or 10 × 1 × 0.001 × 2/70 = 3 × 10-4 7 . The cancer 

8 risk for specific exposure duration scenarios that are listed in the last column are added together to 

9 get the total risk. Thus, a 70-year (lifetime) risk estimate for continuous exposure to 0.001 mg/kgday benzo[a]pyrene is 2 × 10-3 10 , which is adjusted for early-life susceptibility and assumes a 70-year 

11 lifetime and constant exposure across age groups.

12 In calculating the cancer risk for a 30-year constant exposure to benzo[a]pyrene at an 

13 exposure level of 0.001 mg/kg-day for ages 0–30 years, the duration adjustments would be 2/70, 

14/70, and 14/70, and the partial risks for the three age groups would be 3 × 10-4, 6 × 10-4 14 , and 2 × 

10-4, which would result in a total risk estimate of 1 × 10-3 15 .

16 In calculating the cancer risk for a 30-year constant exposure to benzo[a]pyrene at an 

17 exposure level of 0.001 mg/kg-day for ages 20–50 years, the duration adjustments would be 0/70, 

0/70, and 30/70. The partial risks for the three groups are 0, 0, and 4 × 10-4 18 , which would result in 

a total risk estimate of 4 × 10-4 19 .

20 Consistent with the approaches for the oral route of exposure (Table 2-13), the ADAFs 

21 should also be applied when assessing cancer risks for subpopulations with early life exposures to 

22 benzo[a]pyrene via the inhalation and dermal routes (presented in Tables 2-14 and 2-15).
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Table 2-14. Application of ADAFs for the estimation of benzo[a]pyrene cancer 

risk following lifetime (70-year) inhalation exposure

Age group ADAF

Unit risk Exposure concentration

Duration 

adjustment

Cancer risk for 

specific exposure 

duration 

(per μg/m scenarios 3) (μg/m3)

0–<2 yrs 10 5 × 10-41 2 yrs/70 yrs 0.0001

2–<16 yrs 3 5 × 10-41 14 yrs/70 yrs 0.0003

≥16 yrs 1 5 × 10-41 54 yrs/70 yrs 0.0004

Total risk 0.0008

Table 2-15. Application of ADAFs for the estimation of benzo[a]pyrene cancer 

risk following lifetime (70-year) dermal exposure

Age 

group ADAF

Unit risk

Exposure 

concentration

Duration 

adjustment

Cancer risk for 

specific exposure 

duration 

(per μg/d) (μg/d) scenarios

0–<2 yrs 10 0.005 0.001 2 yrs/70 yrs 1 × 10-6

2–<16 yrs 3 0.005 0.001 14 yrs/70 yrs 3 × 10-6

≥16 yrs 1 0.005 0.001 54 yrs/70 yrs 4 × 10-6

Total risk 8 × 10-6
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17

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19 measuring the humoral immune response to SRBC in rats and mice treated with benzo[a]pyrene or 

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25

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28

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55
Toxicological Review of Benzo[a]pyrene

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1 Urso, P; Wirsiy, YG; Zhang, W; et al. (2008) Alterations in CD4+, CD8+, Vgamma3, Vgammadelta, and/or 

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4

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6 pregnancy. Scand J Immunol Suppl 11:203–206.

7

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10 (accessed April 20, 2010).

11

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13 Available online at http://www.epa.gov/iris/backgrd.html (accessed April 20, 2010).

14

15 U.S. EPA. (1988) Recommendations for and documentation of biological values for use in risk assessment. 

16 Prepared by the Environmental Criteria and Assessment Office, Office of Health and Environmental 

17 Assessment, Cincinnati, OH for the Office of Solid Waste and Emergency Response, Washington, DC; EPA 

18 600/6-87/008. Available online at http://www.epa.gov/iris/backgrd.html (accessed April 20, 2010).

19

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21 Environmental Assessment, Washington DC. EPA/600/R-92/135.

22

23 U.S. EPA. (1991a) Dose-response analysis of ingested benzo(a)pyrene (CAS no 50-32-8). Cincinnati, OH: U.S. 

24 Environmental Protection Agency. EPA600R92045.

25

26 U.S. EPA. (1991b) Drinking water criteria document for polycyclic aromatic hydrocarbons (PAHs). Cincinnati, 

27 OH: Agency, Environmental Protection Agency; PB92173459.

28

29 U.S. EPA. (1991c) Guidelines for developmental toxicity risk assessment. Federal Register 56(234):63798–

30 63826. Available online at http://www.epa.gov/iris/backgrd.html (accessed April 20, 2010).

31

32 U.S. EPA. (1992) Draft report: a cross-species scaling factor for carcinogen risk assessment based on 

33 equivalence of mg/kg3/4/day. Federal Register 57(109):24152-24173.

34

35 U.S. EPA. (1994a) Interim policy for particle size and limit concentration issues in inhalation toxicity studies. 

36 Federal Register 59(206):53799. Available online at http://www.epa.gov/iris/backgrd.html (accessed April 

37 20, 2010).

38

39 U.S. EPA. (1994b) Methods for derivation of inhalation reference concentrations and application of inhalation 

40 dosimetry. Office of Research and Development, Washington, DC; EPA/600/8-90/066F. Available online at 

41 http://www.epa.gov/iris/backgrd.html (accessed April 20, 2010).

42

43 U.S. EPA. (1995) Use of the benchmark dose approach in health risk assessment. Risk Assessment Forum, 

44 Washington, DC; EPA/630/R-94/007. Available online at 

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46

47 U.S. EPA. (1996) Guidelines for reproductive toxicity risk assessment. Federal Register 61(212):56274–

48 56322. Available online at http://www.epa.gov/iris/backgrd.html (accessed April 20, 2010).

49

50 U.S. EPA. (1998a) Guidelines for neurotoxicity risk assessment. Federal Register 63(93):26926–26954. 

51 Available online at http://www.epa.gov/iris/backgrd.html (accessed April 20, 2010).

52

53 U.S. EPA. (1998b) Integrated risk information system (IRIS); announcement of 1999 program; request for 

54 information. Federal Register 63(237):68285-68287.

55
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This document is a draft for review purposes only and does not constitute Agency policy.

R-23 DRAFT—DO NOT CITE OR QUOTE

1 U.S. EPA. (2000a) Science policy council handbook: risk characterization. Office of Science Policy, Office of 

2 Research and Development, Washington, DC; EPA 100-B-00-002. Available online at 

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4

5 U.S. EPA. (2000b) Benchmark dose technical guidance document. External review draft. Risk Assessment 

6 Forum, Washington, DC; EPA/630/R-00/001. Available online at http://www.epa.gov/iris/backgrd.html 

7 (accessed April 20, 2010).

8

9 U.S. EPA. (2000c) Supplementary guidance for conducting health risk assessment of chemical mixtures. Risk 

10 Assessment Forum, Washington, DC; EPA/630/R-00/002. Available online at 

11 http://www.epa.gov/iris/backgrd.html (accessed April 20, 2010).

12

13 U.S. EPA. (2002) A review of the reference dose and reference concentration processes. Risk Assessment 

14 Forum, Washington, DC; EPA/630/P-02/0002F. Available online at http://www.epa.gov/iris/backgrd.html 

15 (accessed April 20, 2010).

16

17 U.S. EPA. (2004) Risk assessment guidance for Superfund, volume 1: human health evaluation manual (part E, 

18 supplemental guidance for dermal risk assessment). Office of Solide Waste and Emergency Response, 

19 Washington, DC; EPA/540/R/99/005.

20

21 U.S. EPA. (2005a) Guidelines for carcinogen risk assessment. Risk Assessment Forum, Washington, DC; 

22 EPA/630/P-03/001F. Available online at http://www.epa.gov/iris/backgrd.html (accessed April 20, 2010).

23

24 U.S. EPA. (2005b) Supplemental guidance for assessing susceptibility from early-life exposure to carcinogens. 

25 Risk Assessment Forum, Washington, DC; EPA/630/R-03/003F. Available online at 

26 http://www.epa.gov/iris/backgrd.html (accessed April 20, 2010).

27

28 U.S. EPA. (2006a) Science policy council handbook: peer review. Third edition. Office of Science Policy, Office 

29 of Research and Development, Washington, DC; EPA/100/B-06/002. Available online at 

30 http://www.epa.gov/iris/backgrd.html (accessed April 20, 2010).

31

32 U.S. EPA. (2006b) A framework for assessing health risk of environmental exposures to children. National 

33 Center for Environmental Assessment, Washington, DC, EPA/600/R-05/093F. Available online at 

34 http://cfpub.epa.gov/ncea/cfm/recordisplay.cfm?deid=158363 (accessed March 14, 2011).

35

36 U.S. EPA. (2007) Framework for determining a mutagenic mode of action for carcinogenicity: Using EPA's 

37 2005 cancer guidelines and supplemental guidance for assessing susceptibility from early-life exposure to 

38 carcinogens. External Review Draft. EPA 120/R-07/002-A. Available online at 

39 http://www.epa.gov/osa/mmoaframework/pdfs/MMOA-ERD-FINAL-83007.pdf (accessed April 21, 2009).

40

41 U.S. EPA. (2010) Revised Draft: Multistage Weibull time-to-tumor model in EPA’s Benchmark Dose Software 

42 (BMDS): Methodology Description. Washington, DC: U.S. Environmental Protection Agency.

43

U.S. EPA. (2011) Recommended Use of Body Weight3/4 44 as the Default Method in Derivation of the Oral 

45 Reference Dose. EPA/100/R11/0001.

46

47 van den Berg, FA; Baken, KA; Vermeulen, JP; et al. (2005) Use of the local lymph node assay in assessment of 

48 immune function. Toxicology 211(1-2):107–114.

49

50 Van Lipzig, MMH; Vermeulen, NPE; Gusinu, R; et al. (2005) Formation of estrogenic metabolites of 

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53

54 Varghese, SK; Gangamma, S. (2009) Particle deposition in human respiratory system: Deposition of 

55 concentrated hygroscopic aerosols. Inhalation toxicology 21(7):619-630.
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This document is a draft for review purposes only and does not constitute Agency policy.

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