Health Effects Support 

Document for 

Perfluorooctanoic Acid 

(PFOA)

United States 

Environmental Protection 

Agency 

Office of Water

Mail Code 4304T

EPA 822-R-16-003

May 2016

Perfluorooctanoic Acid – May 2016 i 
Health Effects Support Document

for 

Perfluorooctanoic Acid (PFOA)

U.S. Environmental Protection Agency 

Office of Water (4304T)

Health and Ecological Criteria Division

Washington, DC 20460 

EPA Document Number: 822-R-16-003

May 2016 

Perfluorooctanoic Acid – May 2016 ii
BACKGROUND

The Safe Drinking Water Act (SDWA), as amended in 1996, requires the Administrator of 

the U.S. Environmental Protection Agency (EPA) to periodically publish a list of unregulated 

chemical contaminants known or anticipated to occur in public water systems and that may

require regulation under SDWA. The SDWA also requires the Agency to make regulatory 

determinations on at least five contaminants on the Contaminant Candidate List (CCL) every 

5 years. For each contaminant on the CCL, before EPA makes a regulatory determination, the 

Agency needs to obtain sufficient data to conduct analyses on the extent to which the 

contaminant occurs and the risk it poses to populations via drinking water. Ultimately, this 

information will assist the Agency in determining the most appropriate course of action in 

relation to the contaminant (e.g., developing a regulation to control it in drinking water, 

developing guidance, or deciding not to regulate it).

The PFOA health assessment was initiated by the Office of Water, Office of Science and 

Technology in 2009. The draft Health Effects Support Document for Perfluoroctanoic Acid 

(PFOA) was completed in 2013 and released for public comment in February 2014. An external 

peer-review panel meeting was held on August 21 and 22, 2014. The final document reflects 

input from the panel as well as public comments received on the draft document. Both the peerreviewed draft and this document include only the sections of a health effects support document 

(HESD) that cover the toxicokinetics and health effects of PFOA. If a decision is made to 

regulate the contaminant, this document will be expanded. 

One of the challenges inherent in conducting this assessment was the wealth of experimental 

data published before and during its development. This section provides a synopsis of the 

approach used in identifying and selecting the publications reflected in the final assessment. 

Data were identified through the following:

• Monthly/bimonthly literature searches conducted by EPA library staff (2009–2015) and 

New Jersey Department of Environmental Protection library staff (2012–2015). 

• Papers identified by EPA internal and external peer reviewers. 

• Papers identified through public comments on the draft assessments. 

• Papers submitted to EPA by the public. 

In mid-2013, the EPA library searches were expanded to cover all members of the 

perfluoroalkane carboxylate family (C4 through C12). Appendix A describes the literature search 

strategy used by the libraries. Through the literature search, documents were identified for 

retrieval, review, and inclusion in the HESD using the following criteria: 

• The study examines a toxicity endpoint or population not examined by studies already 

included in the draft document. 

• Aspects of the study design such as the size of the population exposed or quantification 

approach make it superior to key studies already included in the draft document. 

• The data contribute substantially to the weight of evidence for any of the toxicity 

endpoints covered by the draft document. 

• Elements of the study design merit its inclusion in the draft document based on its 

contribution to the mode of action (MoA) or the quantification approach. 

Perfluorooctanoic Acid – May 2016 iii
• The study elucidates the MoA for any toxicity endpoint or toxicokinetic property 

associated with PFOA exposure. 

• The effects observed differ from those in other studies with comparable protocols.

In addition to each publication being evaluated against the criteria above, the relevance of the 

study to drinking water exposures and to the U.S. population also were considered.

The studies included in the final draft were determined to provide the most current and 

comprehensive description of the toxicological properties of PFOA and the risk it poses to 

humans exposed to it in their drinking water. Appendix B summarizes the studies evaluated for 

inclusion in the HESD following the August 2014 peer review and identifies those selected for 

inclusion in the final assessment. Appendix B includes epidemiology data that provide a highlevel summary of the outcomes across the studies evaluated.

Development of the hazard identification and dose-response assessment for PFOA has 

followed the general guidelines for risk assessment forth by the National Research Council 

(1983) and EPA’s Framework for Human Health Risk Assessment to Inform Decision Making

(USEPA 2014a). Other EPA guidelines used in the development of this assessment include the 

following: 

• Guidelines for the Health Risk Assessment of Chemical Mixtures (USEPA 1986a) 

• Guidelines for Mutagenicity Risk Assessment (USEPA 1986b) 

• Recommendations for and Documentation of Biological Values for Use in Risk 

Assessment (USEPA 1988) 

• Guidelines for Developmental Toxicity Risk Assessment (USEPA 1991) 

• Interim Policy for Particle Size and Limit Concentration Issues in Inhalation Toxicity 

Studies (USEPA 1994a) 

• Methods for Derivation of Inhalation Reference Concentrations and Application of 

Inhalation Dosimetry (USEPA 1994b) 

• Use of the Benchmark Dose Approach in Health Risk Assessment (USEPA, 1995) 

• Guidelines for Reproductive Toxicity Risk Assessment (USEPA 1996) 

• Guidelines for Neurotoxicity Risk Assessment (USEPA 1998) 

• Science Policy Council Handbook: Peer Review (2nd edition) (USEPA 2000a) 

• Supplemental Guidance for Conducting Health Risk Assessment of Chemical Mixtures

(USEPA 2000b) 

• A Review of the Reference Dose and Reference Concentration Processes (USEPA 2002a)

• Guidelines for Carcinogen Risk Assessment (USEPA 2005a) 

• Supplemental Guidance for Assessing Susceptibility from Early-Life Exposure to 

Carcinogens (USEPA 2005b) 

• Science Policy Council Handbook: Peer Review (3rd edition) (USEPA 2006a) 

• A Framework for Assessing Health Risks of Environmental Exposures to Children

(USEPA 2006b) 

• Exposure Factors Handbook (USEPA 2011) 

• Benchmark Dose Technical Guidance Document (USEPA 2012) 

• Child-Specific Exposure Scenarios Examples (USEPA 2014b) 

Perfluorooctanoic Acid – May 2016 iv
AUTHORS

Joyce Morrissey Donohue, Ph.D. (Chemical Manager)

Office of Water

U.S. Environmental Protection Agency, Washington, DC

Tina Moore Duke, M.S. (previously with Office of Water, U.S. Environmental Protection 

Agency) 

John Wambaugh, Ph.D. 

Office of Research and Development

U.S. Environmental Protection Agency, Research Triangle Park, NC 

The following contractors supported the development of this document:

Jennifer Rayner, Ph.D., D.A.B.T. 

Environmental Sciences Division

Oak Ridge National Laboratory, Oak Ridge, TN 

Carol S. Wood, Ph.D., D.A.B.T. 

Environmental Sciences Division

Oak Ridge National Laboratory, Oak Ridge, TN 

This document was prepared under the U.S. EPA Contract No. DW-8992342701, Work 

Assignment No. 2011-001 with Oak Ridge National Laboratory. The Lead U.S. EPA Scientist is 

Joyce Morrissey Donohue, Ph.D., Health and Ecological Criteria Division, Office of Science and 

Technology, Office of Water.

The Oak Ridge National Laboratory is managed and operated by UT-Battelle, LLC., for the U.S. 

Department of Energy under Contract No. DE-AC05-00OR22725. 

Contributors and Reviewers

Internal Contributors and Reviewers

Office of Water, U.S. Environmental Protection Agency

Elizabeth Doyle, Ph.D. (retired) 

Edward Hackett

Office of Research and Development, U.S. Environmental Protection Agency 

Glinda Cooper, Ph.D. 

Barbara Glenn, Ph.D. 

Erin Hines, Ph.D.

Christopher Lau, Ph.D.

Matthew Lorber, Ph.D. 

Jaqueline Moya 

Linda Phillips, Ph.D. 

Paul White, Ph.D. 

Michael Wright, Sc.D. 

Perfluorooctanoic Acid – May 2016 v 
Office of Chemical Safety and Pollution Prevention, U.S. Environmental Protection Agency 

E. Laurence Libelo

Andrea Pfehales-Hutchens, Ph.D. 

Tracy Williamson

David Lai, Ph.D. (retired)

Jennifer Seed, Ph.D. (retired) 

Office of Childrens Health Protection, U.S. Environmental Protection Agency 

Gregory Miller 

Office of Land and Emergency Management, U.S. Environmental Protection Agency 

External Reviewers

James Bruckner, Ph.D. 

Department of Pharmacology and Toxicology 

University of Georgia, Athens, GA 

Deborah Cory-Slechta, Ph.D.

Department of Environmental Medicine 

University of Rochester Medical Center, Rochester, NY

Jamie DeWitt, Ph.D. 

Pharmacology and Toxicology 

East Carolina University, Greenville, NC

Jeffrey Fisher, Ph.D. 

Biochemical Toxicology, NCTR 

U.S. Food & Drug Administration, Jefferson, AK 

William Hayton, Ph.D. 

College of Pharmacy (Emeritus)

The Ohio State University, Columbus, OH 

Matthew Longnecker, M.D., Sc.D. 

Biomarker-based Epidemiology Group 

National Institute of Environmental Health Sciences, Research Triangle Park, NC

Angela Slitt, Ph.D. 

Biomedical and Pharmaceutical Sciences

University of Rhode Island, Kingston, RI 

Perfluorooctanoic Acid – May 2016 vi
CONTENTS

BACKGROUND ........................................................................................................................... iii

ABBREVIATIONS AND ACRONYMS .................................................................................... xiv

EXECUTIVE SUMMARY .......................................................................................................ES-1

1 IDENTITY: CHEMICAL AND PHYSICAL PROPERTIES .............................................. 1-1

2 TOXICOKINETICS ............................................................................................................. 2-1

2.1 Absorption.................................................................................................................... 2-1

2.1.1 Oral Exposure ........................................................................................................ 2-2

2.1.2 Inhalation Exposure ............................................................................................... 2-2

2.1.3 Dermal Exposure ................................................................................................... 2-3

2.2 Distribution .................................................................................................................. 2-3

2.2.1 Oral Exposure ........................................................................................................ 2-7

2.3 Metabolism ................................................................................................................ 2-23

2.4 Excretion .................................................................................................................... 2-23

2.5 Animal Studies ........................................................................................................... 2-24

2.5.1 Mechanistic Studies of Renal Excretion .............................................................. 2-28

2.6 Toxicokinetic Considerations .................................................................................... 2-34

2.6.1 PK Models ........................................................................................................... 2-34

2.6.2 Half-Life Data ...................................................................................................... 2-46

2.6.3 Volume of Distribution Data ............................................................................... 2-51

2.6.4 Toxicokinetic Summary ....................................................................................... 2-51

3 HAZARD IDENTIFICATION ............................................................................................ 3-1

3.1 Human Studies ............................................................................................................. 3-1

3.1.1 Noncancer .............................................................................................................. 3-3

3.1.1.1 Serum Lipids and Cardiovascular Diseases ..................................................... 3-3

3.1.1.2 Cardiovascular Diseases ................................................................................ 3-13

3.1.1.3 Liver Enzymes and Liver Disease ................................................................. 3-13

3.1.1.4 Biomarkers of Kidney Function and Kidney Disease .................................... 3-18

3.1.1.5 Immunotoxicity .............................................................................................. 3-22

3.1.1.6 Thyroid Effects .............................................................................................. 3-26

3.1.1.7 Diabetes and Related Endpoints .................................................................... 3-35

3.1.1.8 Reproductive and Developmental Endpoints ................................................. 3-37

3.1.1.9 Steroid Hormones .......................................................................................... 3-47

3.1.1.10 Neurodevelopment ......................................................................................... 3-48

3.1.1.11 Postnatal Development .................................................................................. 3-51

3.1.1.12 Summary and Conclusions from the Human Epidemiology Studies ............. 3-52

3.1.2 Cancer .................................................................................................................. 3-55

3.1.2.1 Summary and Conclusions from the Human Cancer Epidemiology Studies .. 3-60

3.2 Animal Studies ........................................................................................................... 3-60

3.2.1 Acute Toxicity ..................................................................................................... 3-61

3.2.2 Short-Term Studies .............................................................................................. 3-62

Perfluorooctanoic Acid – May 2016 vii
3.2.3 Subchronic Studies............................................................................................... 3-74

3.2.4 Neurotoxicity ....................................................................................................... 3-77

3.2.5 Developmental/Reproductive Toxicity ................................................................ 3-79

3.2.6 Prenatal Development .......................................................................................... 3-84

3.2.7 Mammary Gland Development and Other Specialized Developmental Studies ... 3-89

3.2.8 Chronic Toxicity ................................................................................................ 3-102

3.2.9 Carcinogenicity .................................................................................................. 3-106

3.3 Other Key Data ........................................................................................................ 3-110

3.3.1 Mutagenicity and Genotoxicity .......................................................................... 3-110

3.3.2 Immunotoxicity .................................................................................................. 3-112

3.3.3 Hormone Disruption .......................................................................................... 3-124

3.3.4 Physiological or Mechanistic Studies ................................................................ 3-129

3.3.5 Structure-Activity Relationship ......................................................................... 3-139

3.4 Hazard Characterization ........................................................................................... 3-140

3.4.1 Synthesis and Evaluation of Major Noncancer Effects ..................................... 3-140

3.4.2 Synthesis and Evaluation of Carcinogenic Effects ............................................ 3-150

3.4.3 Mode of Action and Implications in Cancer Assessment .................................. 3-152

3.4.4 Weight of Evidence Evaluation for Carcinogenicity ......................................... 3-158

3.4.5 Potentially Sensitive Populations ....................................................................... 3-159

4 DOSE-RESPONSE ASSESSMENT .................................................................................... 4-1

4.1 Dose-Response for Noncancer Effects ........................................................................ 4-1

4.1.1 RfD Determination ................................................................................................. 4-1

4.1.1.1 PK Model approach ......................................................................................... 4-8

4.1.1.2 RfD Quantification ......................................................................................... 4-14

4.1.2 RfD Selection ....................................................................................................... 4-16

4.1.3 RfC Determination ............................................................................................... 4-17

4.2 Dose-Response for Cancer Effects ............................................................................ 4-18

5 REFERENCES ..................................................................................................................... 5-1

Appendix A: Literature Search Strategy Developing the Search ............................................... A-1

Appendix B: Studies Evaluated Since August 2014 ....................................................................B-1

Appendix C: Multistage Model for Leydig Cell Tumors ............................................................C-1

Perfluorooctanoic Acid – May 2016 viii
TABLES

Table 1-1. Chemical and Physical Properties of PFOA ............................................................... 1-3

Table 2-1. Protein Binding in Rat, Human, and Monkey Plasma ................................................ 2-4

Table 2-2. Dissociation Constants (Kd) of Binding Between PFOA and Albumin ..................... 2-4

Table 2-3. Percent (%) Binding of PFOA to Human Plasma Protein Fractions .......................... 2-6

Table 2-4. Tissue Distribution of PFOA in Wistar Rats After 28 Days of Treatment ................ 2-9

Table 2-5. Distribution of PFOA in Male Sprague-Dawley Rats After Oral Exposure 

Dose ........................................................................................................................ 2-10

Table 2-6. Distribution of PFOA in Female Sprague-Dawley Rats after Oral Exposure 

Dose ........................................................................................................................ 2-11

Table 2-7. PFOA Concentrations in Wild-type and PPARα-null Mice (μg/mL) ...................... 2-12

Table 2-8. Plasma PFOA Concentrations (μg/ml) in Postweaning Sprague-Dawley Rats ....... 2-14

Table 2-9. Plasma PFOA Concentrations in Male Rats ............................................................. 2-14

Table 2-10. Plasma PFOA Concentrations in Female Rats ....................................................... 2-15

Table 2-11. Maternal Plasma PFOA Levels (μg/ml) in Rats During Gestation and 

Lactation ................................................................................................................. 2-17

Table 2-12. Placenta, Amniotic Fluid, and Embryo/Fetus PFOA Concentrations in Rats 

(μg/ml) .................................................................................................................... 2-17

Table 2-13. Fetus/Pup PFOA Concentration (μg/ml) in Rats During Gestation and 

Lactation ................................................................................................................. 2-18

Table 2-14. PFOA Levels (μg/ml) in Rats Maternal Milk During Lactation ............................ 2-18

Table 2-15. PFOA Levels (ng/ml) in Mice During Gestation and Lactation in Selected 

Fluids and Tissues .................................................................................................. 2-19

Table 2-16. Female Offspring PFOA Levels (ng/ml) in Mice After GD 1-17 Exposure .......... 2-20

Table 2-17. Female Offspring Serum PFOA Levels (ng/ml) in Mice After GD 10-17 

Exposure ................................................................................................................. 2-21

Table 2-18. Serum PFOA Levels (ng/ml) in Mice Over Three Generations ............................. 2-21

Table 2-19. Urine PFOA Concentrations in Male and Female Rats .......................................... 2-25

Table 2-20. Cumulative Percent 14C-PFOA Excreted in Urine and Feces by Rats ................... 2-26

Table 2-21. Cumulative Percent 14C-PFOA Excreted in Urine and Feces ................................ 2-27

Table 2-22. Kinetic Parameters of Perfluorinated Carboxylate Transport by OAT1, 

OAT3, and OATP1a1 ............................................................................................. 2-33

Table 2-23. Plasma and Urine PFOA Concentration 24-hr After Treatment with 30 mg/kg 

PFOA ...................................................................................................................... 2-34

Table 2-24. Model Parameters for 1 and 10 mg/kg Single Doses of PFOA to CD1 Mice ........ 2-37

Table 2-25. Pharmacokinetic Parameters from Wambaugh et al. (2013) Meta-Analysis of 

Literature Data ........................................................................................................ 2-46

Perfluorooctanoic Acid – May 2016 ix
Table 2-26. PK Parameters in Male Rats Following Administration of PFOA ......................... 2-50

Table 2-27. PK Parameters in Female Rats Following Administration of PFOA ..................... 2-50

Table 3-1. Summary of PFOA Occupational Exposure Studies of PFOA and Serum 

Lipids ........................................................................................................................ 3-7

Table 3-2. Summary of High-Exposure Community Studies of PFOA and Serum Lipids ......... 3-9

Table 3-3. Summary of General Population Epidemiology Studies of PFOA with Serum 

Lipids ...................................................................................................................... 3-12

Table 3-4. Summary of Epidemiology Studies of PFOA and Liver Enzymes .......................... 3-14

Table 3-5. Summary of Epidemiology Studies of PFOA and Measures of Kidney 

Function .................................................................................................................. 3-18

Table 3-6. Summary of Epidemiology Studies of PFOA and Immune Suppression 

(Vaccine Response) ................................................................................................ 3-23

Table 3-7. Summary of Epidemiology Studies of PFOA and Thyroid Effects in Adults ......... 3-27

Table 3-8. Summary of Epidemiology Studies of PFOA and Thyroid Effects in Special 

Populations ............................................................................................................. 3-29

Table 3-9. Summary of Epidemiology Studies of PFOA and Pregnancy-Induced 

Hypertension or Preeclampsia ................................................................................ 3-39

Table 3-10. Summary of Epidemiology Studies of PFOA and Birth Weight ........................... 3-41

Table 3-11. Summary of Epidemiology Studies of PFOA and Pubertal Development ............. 3-45

Table 3-12. Summary of PFOA Epidemiology Studies of Kidney and Testicular Cancer ....... 3-56

Table 3-13. Comparison of PPAR-α Related Effects in Rats for PFOA, DEHP, and 

Fenofibrate after a 3-day Exposure ........................................................................ 3-62

Table 3-14. Hepatic Effects of Rats Exposed to PFOA ............................................................. 3-65

Table 3-15. Hepatic Effects in PFOA-Treated Mice ................................................................. 3-68

Table 3-16. Mouse Hepatocyte Ultrastructure After PFOA or Wythe 14,643 Treatment ......... 3-69

Table 3-17. Relative Response of hPPARα, mPPARα, and PPARα-null Mice to PFOA ......... 3-70

Table 3-18. Liver Effects in Male Rats ...................................................................................... 3-76

Table 3-19. Organ Weight Data from F0 Male Rats ................................................................. 3-80

Table 3-20. Organ Weight Data from F1 Male Rats ................................................................. 3-82

Table 3-21. Studies of Pregnant CD-1 Mice Following Administration of PFOA ................... 3-90

Table 3-22. Mammary Gland Measurements at PND 21 from Female Offspring of Dams 

Treated GD 10–17 .................................................................................................. 3-95

Table 3-23. Dosing Regimens Used in the Multigeneration Study of CD-1 Mice .................... 3-96

Table 3-24. Mammary Gland Scores from Three Generations of CD-1 Female Mice ............. 3-97

Table 3-25. Liver Weight Data in Monkeys Administered PFOA for 6 Months .................... 3-103

Table 3-26. Subcellular Liver Enzyme Activities and Liver PFOA Concentrations ............... 3-104

Perfluorooctanoic Acid – May 2016 x 
Table 3-27. Clinical Chemistry Values from Male Rats Given PFOA for 2 Years ................. 3-105

Table 3-28. Incidence of Nonneoplastic Lesions in Rats Given PFOA for 2 Years ............... 3-106

Table 3-29. Incidence of Ovarian Stromal Hyperplasia and Adenoma in Rats ....................... 3-107

Table 3-30. Mammary Gland Tumor Incidence Comparison .................................................. 3-108

Table 3-31. Liver Tumors in Three Strains of Mice at 18 Months with Exposure to PFOA 

Only during Gestation and Lactation ................................................................... 3-109

Table 3-32. Genotoxicity of PFOA In Vitro ............................................................................ 3-111

Table 3-33. Selected Clinical Chemistry Parameters in Mice Treated with PFOA for 5 

Days ...................................................................................................................... 3-118

Table 3-34. Selected Clinical Chemistry Parameters in Mice Treated with PFOA for 

10 Days ................................................................................................................. 3-119

Table 3-35. Impact of PFOA on Splenic and Thymic Lymphocyte Populations .................... 3-120

Table 3-36. Estimated EC50 Values ......................................................................................... 3-126

Table 3-37. Data Collection for Female Mice Gestationally Exposed to PFOA ..................... 3-127

Table 3-38. mRNA Expression of Hepatic PPARα and Related Genes .................................. 3-132

Table 3-39. Activation of Mouse and Human PPAR by PFOA .............................................. 3-135

Table 4-1. NOAEL/LOAEL Data for Oral Subchronic and Chronic Studies of PFOA .............. 4-3

Table 4-2. Shorter-term and Developmental Oral Exposure Studies .......................................... 4-5

Table 4-3. Predicted Final Serum Concentration and Time-Integrated Serum 

Concentration (AUC) for Studies in Rats ................................................................. 4-9

Table 4-4. Predicted Final Serum Concentration and Time-Integrated Serum 

Concentration (AUC) for Studies in Mice .............................................................. 4-10

Table 4-5. Predicted Final Serum Concentration and Time-Integrated Serum (AUC) in 

Studies of Monkeys ................................................................................................ 4-10

Table 4-6. Average Serum Concentrations Derived from the AUC and the Duration of 

Dosing..................................................................................................................... 4-11

Table 4-7. Comparison of Average Serum Concentration and Steady-State Concentration ..... 4-12

Table 4-8. HEDs Derived from the Modeled Animal Average Serum Values ......................... 4-13

Table 4-9. The Impact of Quantification Approach on the RfD Outcomes for the HEDs 

from the PK Model Average Serum Values ........................................................... 4-15

Table 4-10. Summary of Tumor Data from Animal Studies ..................................................... 4-19

Table 4-11. Multistage Cancer Model Dose Prediction Results for a 4% Increase in LCT 

Incidence ................................................................................................................. 4-21

Table B-1. PFOS Epi Papers—Post Peer Review (Retrieved and Reviewed) ............................B-1

Table B-2. PFOA Post Peer Review Animal Toxicity Studies ....................................................B-3

Table B-3. Toxicokinetics: Post Peer Review .............................................................................B-5

Table B-4. Association between Serum PFOA and Serum Lipids and Uric Acid ......................B-6

Perfluorooctanoic Acid – May 2016 xi
Table B-5. Association of Serum PFOA and Biochemical and Hematological Measures ..........B-8

Table B-6. Association between PFOA level and prevalence of thyroid disease and 

thyroid hormone levels .............................................................................................B-9

Table B-7. Association between Serum PFOA and Markers of Immunotoxicity .....................B-11

Table B-8. Association between Serum PFOA and Reproductive and Developmental 

Outcomes ................................................................................................................B-12

Perfluorooctanoic Acid – May 2016 xii
FIGURES

Figure 1-1. Chemical Structures of PFOA and APFO ................................................................. 1-1

Figure 1-2. PFOA Anti-Conformer.............................................................................................. 1-1

Figure 1-3. PFOA Lowest Energy Conformer ............................................................................. 1-2

Figure 2-1. PFOA Binding Sites on HSA .................................................................................... 2-5

Figure 2-2. Localization of Transport Proteins .......................................................................... 2-29

Figure 2-3. Schematic for a Physiologically Motivated Renal Resorptions PK Model. ........... 2-35

Figure 2-4. Physiologically Motivated Pharmacokinetic Model Schematic for PFOAExposed Rats .......................................................................................................... 2-36

Figure 2-5. Schematic for One-Compartment Model. ............................................................... 2-37

Figure 2-6. PK Model of Gestation and Lactation in Mice ....................................................... 2-39

Figure 2-7. Structure of the PFOA PBPK Model in Monkeys and Humans ............................. 2-40

Figure 2-8. Structure of the PBPK Model for PFOA in the Adult Sprague-Dawley Rat .......... 2-42

Figure 2-9. PBPK Model Structure for Simulating PFOA and PFOS Exposure During 

Pregnancy in Humans (Maternal, Left; Fetal, Right) ............................................. 2-43

Figure 3-1. PPARα Agonist MoA for Liver Tumors ............................................................... 3-154

Figure 4-1. BMD Model Results for LCTs (Butenhoff et al. 2012) .......................................... 4-21

Perfluorooctanoic Acid – May 2016 xiii
ABBREVIATIONS AND ACRONYMS

8-OH-dG 8-hydroxydeoxyguanosine 

Acot1 acyl-CoA thioesterase (human)

Acox acyl-CoA oxidase 

ADHD attention deficit hyperactivity disorder

ADX adrenalectomized

AIC Akaike’s Information Criterion

ALP alkaline phosphatase

ALT alanine aminotransferase

ANOVA analysis of variance

APFO ammonium perfluorooctanoate

Areg amphiregulin 

AST aspartate aminotransferase

ATP adenosine triphosphate

AUC area under the plasma concentration time curve

AUCINF area under the plasma concentration time curve, extrapolated to infinity

AUCINF/D area under the plasma concentration time curve, extrapolated to infinity, 

normalized to dose 

BAX BCL2-associated X protein

BMD benchmark dose 

BMDL lower 95th percentile confidence bound on benchmark dose 

BMDS Benchmark Dose Software

BMI body mass index 

BrdU Bromodeoxyuridine (5-bromo-2-deoxyuridine) 

BSA bovine serum albumin 

BSEP bile salt export pump

BSP sulfobromophthalein 

BUN blood urea nitrogen 

bw body weight 

C Celsius

Cmax peak plasma concentration at the first intestinal absorption loci

CaMKII calcium/calmodulin-dependent protein kinase II 

CAR constitutive androstane receptor

CAT carnitine acyltransferase

CCK cholecystokinin 

CCL Contaminant Candidate List

CCL 3 Contaminant Candidate List 3 

CFSE 6-carboxyfluorescein succinimidyl ester 

ChAT choline acetyltransferase

CHO Chinese hamster ovary

CI confidence interval 

CL clearance

Clp plasma clearance

CLR renal clearance

CoA coenzyme A

ConA concanavalinA 

COPD chronic obstructive pulmonary disease 

CORT corticosterone 

Perfluorooctanoic Acid – May 2016 xiv
Cox II cytochrome c oxidase subunit II 

Cox IV cytochrome c oxidase subunit IV 

CPT carnitine palmitoyltransferase

Crl Charles River Laboratory

CSF cancer slope factor

Cte acyl-CoA thioesterase (rat)

CYP4A10 cytochrome P450 4a10 

d day 

DCDQ Developmental Coordination Disorder Questionnaire

Ddit3 DNA damage inducible transcript

DEHP di(2-ethylhexyl) phthalate 

DHT 5α-dihydroxy-testosterone

dL deciliter

DMSO dimethyl sulfoxide 

DNA deoxyribonucleic acid

DR dose rate

DTH delayed-type hypersensitivity 

DWI drinking water intake 

E2 17-β estradiol

E3S estrone-3-sulfate

EC50 half maximal effective concentration

ECF electrochemical fluorination

eGFR estimated glomerular filtration rate

EGFR epidermal growth factor receptor

EPA U.S. Environmental Protection Agency 

ER endoplasmic reticulum

ERα estrogen receptor α

Errα estrogen-related receptor α

FID flame ionization detector

FSH follicle-stimulating hormone 

FT4 free thyroxine 

FXR farnesoid receptor 

g gram 

GAP-43 growth-associated protein-43 

GD gestation day 

GEE generalized estimating equation

GFR glomerular filtration rate

GGT gamma-glutamyl transpeptidase 

GJIC gap junction intercellular communication 

GlyT glycogen trophoblast cell

GnRH gonadotropin releasing hormone 

GSD geometric standard deviation

GST glutathione-S-transferase

hCG human chorionic gonadotropin 

HDL high-density lipoprotein

HED human equivalent dose 

HEK human embryonic kidney 

HESD health effects support document 

HET heterozygous 

Perfluorooctanoic Acid – May 2016 xv
HFD high-fat diet

HGFα hepatocyte growth factor 

HL-60 human promyelocytic leukemia cell line 

HMG-CoA 3-hydroxy-3-methylglutaryl coenzyme A 

HPLC/MS High-performance liquid chromatography mass spectrometry

HPLC/MS/MS High-performance liquid chromatography tandem mass spectrometry

HR hazard ratio

HRBC horse red blood cells 

HSA human serum albumin 

HSD17β1 hydroxysteroid 17β dehydrogenase 1

HSD3β1 hydroxysteroid 3β dehydrogenase 1

IC50 half-maximal inhibiting concentration 

ICR imprinting control region

IDL intermediate density lipoprotein

IgE immunoglobulin E 

IGF-I insulin like growth factor I

IgM immunoglobulin M 

IHD ischemic heart disease

IL-6 interleukin 6

INUENDO Biopersistent Organochlorines in Diet and Human Fertility study 

IQR interquartile range

IRR incidence rate ratio

IU international unit 

IV intravenous 

Ka adsorption rate constant 

Kd dissociation constant

Ke elimination rate constant

kg kilogram 

Km substrate concentration at which the initial reaction rate is half maximal

Koc organic carbon water partitioning coefficient 

Kt affinity constant 

L liter

L-FABP liver fatty acid binding protein

LC50 lethal concentration for 50% of animals 

LCT Leydig cell tumor 

LD lactation day

LD50 lethal dose for 50% of animals 

LDH lactic dehydrogenase

LDL low-density lipoprotein

LH luteinizing hormone 

LHWA Little Hocking Water Association

LLOQ lower limit of quantification

LOAEL lowest observed adverse effect level

LOD limit of detection

LOQ limit of quantitation 

LPS lipopolysaccharide 

m meter

MCAD medium chain acyl-CoA dehydrogenase 

MDA malondialdehyde 

Perfluorooctanoic Acid – May 2016 xvi
Mdr2 multidrug resistance protein 2

μg microgram

mg milligram

min minute 

mL milliliter

μm micrometer

μmol micromole

MMAD mass median aerodynamic diameter

MOA mechanism of action

MoA mode of action 

mol mole

mPL mouse placental lactogen

mPLP mouse prolactin-like protein

mRNA messenger ribonucleic acid

MRP multidrug resistance-associated protein

MTT 3-(4,5-dimethylthiazol-2-yl)-2,5-diphenyltetrazolium bromide

Nd2 NADH dehydrogenase 2 

Ndufs8 NADH dehydrogenase iron-sulfur protein 8 

ng nanogram 

NHANES National Health and Nutrition Examination Survey 

NIEHS National Institute for Environmental Health Sciences

NJDEP New Jersey Department of Environmental Protection 

NK natural killer

NM not monitored 

nmol nanomolar 

NMR nuclear magnetic resonance

NMRI Naval Medical Research Institute

NOAEL no observed adverse effect level

Nrf1 nuclear respiratory factor 1

Nrf2 nuclear respiratory factor 2

NTCP sodium-taurocholate contransporting polypeptide

OAT organic anion transporter

OATP organic anion transporting polypeptide 

OR odds ratio 

OVA ovalbumin 

OVX ovariectomized

OW Office of Water

P progesterone 

PACT pancreatic acinar cell tumor

PAH polycyclic aromatic hydrocarbon 

PB phenobarbital 

PBMC peripheral blood mononuclear cells 

PBPK physiologically based pharmacokinetic 

PCNA proliferating cell nuclear antigen

PenH enhanced pause airway respiration

PFAA perfluoroalkyl acid

PFAS perfluoroalkyl substance 

PFC plaque-forming cell 

PFDA perfluorodecanoic acid 

Perfluorooctanoic Acid – May 2016 xvii
PFHxA perfluorohexanoic acid 

PFHxS perfluorohexanesulfonic acid 

PFNA perfluorononanoic acid 

PFOA perfluorooctanoic acid 

PFOS perfluorooctane sulfonate 

Pgc-1α peroxisome proliferator-activated receptor gamma coactivator 1α

PH peroxisomal bifunctional protein 

PHA phytohemagglutinin 

PK pharmacokinetic 

pKa acid dissociation constant

PND postnatal day 

POD point of departure 

PPAR peroxisome proliferator-activated receptor

ppb parts per billion

ppm parts per million

PT peroxisomal thiolase 

PWG pathology working group 

PXR pregnane X receptor

Q flow in and out of tissues 

Qfilc median fraction of blood flow to the filtrate

RfC reference concentration

RfD reference dose

RFD regular fat diet

ROS reactive oxygen species

RR relative risk

RSA rodent serum albumin 

RSC relative source contribution

RT-PCR reverse transcription polymerase chain reaction

RXRα retinoid X receptor alpha

SD standard deviation

SDH sorbitol dehydrogenase 

SDQ Strengths and Difficulties Questionnaire

SDWA Safe Drinking Water Act

SHBG sex hormone-binding globulin 

SIAR SIDS Initial Assessment Report

SIR standardized incidence ratio

SMR standardized mortality ratio

SOD superoxide dismutase

SPI Society of the Plastics Industry

SRBC sheep red blood cells 

S-TGC sinusoidal trophoblast giant cells 

T3 triiodothyronine 

T4 thyroxine 

T1/2 elimination half-time

Tm transporter maximum

Tmax time of maximum plasma concentration

TC total cholesterol

TCPOBOP 1,4-bis[2-(3,5-dichloropyridyloxy)] benzene 

Tfam transcription factor A

Perfluorooctanoic Acid – May 2016 xviii
TG triglyceride

TH tyrosine hydroxylase 

TNFα tumor necrosis factor α

TPO thyroid peroxidise 

TRR total reactive residues

TSH thyroid stimulating hormone 

TTP time to pregnancy 

TTR thyroid hormone transport protein, transthyretin 

UA uric acid

UCMR Unregulated Contaminant Monitoring Rule 

UCMR 1 Unregulated Contaminant Monitoring Rule 1 

UCMR 2 Unregulated Contaminant Monitoring Rule 2 

UF uncertainty factor

URAT urate transporter

USGS U.S. Geological Survey

Vd volume of distribution 

Vmax maximum initial rate of an enzyme catalyzed reaction

VLCAD very long chain acyl-CoA dehydrogenase 

VLDL very low-density lipoprotein 

VOC volatile organic compound 

WHO World Health Organization 

WRF Water Research Foundation

Perfluorooctanoic Acid – May 2016 xix
EXECUTIVE SUMMARY

Perfluorooctanoic acid (PFOA) is a synthetic, fully fluorinated, organic acid used in a variety 

of consumer products and in the production of fluoropolymers and generated as a degradation 

product of other perfluorinated compounds. Because of strong carbon-fluorine bonds, PFOA is 

stable to metabolic and environmental degradation. PFOA is one of a large group of 

perfluoroalkyl substances (PFASs) that are used to make products more resistant to stains, 

grease, and water. These compounds have been widely found in consumer and industrial 

products as well as in food items. Major U.S. manufacturers voluntarily agreed to phase out 

production of PFOA by the end of 2015. Exposure to PFOA in the United States remains 

possible due to its legacy uses, existing and legacy uses on imported goods, degradation of 

precursors, and extremely high persistence in the environment and the human body. 

Extensive data on humans and animals indicate ready absorption of PFOA and distribution of 

the chemical throughout the body by noncovalent binding to plasma proteins. Studies of 

postmortem human tissues identify its presence in liver, lung, kidney, and bone. PFOA is not 

readily eliminated from the human body as evidenced by the half-life of 2.3 years among 

members of the general population. In contrast, half-life values for the monkey, rat, and mouse 

are 20.8 days, 11.5 days, and 15.6 days, respectively. 

Human epidemiology data report associations between PFOA exposure and high cholesterol, 

increased liver enzymes, decreased vaccination response, thyroid disorders, pregnancy-induced 

hypertension and preeclampsia, and cancer (testicular and kidney). Epidemiology studies 

examined workers at PFOA production plants, a high-exposure community population near a 

production plant in the United States (i.e., the C8 cohort), and members of the general population 

in the United States, Europe, and Asia. These studies examined the relationship between serum 

PFOA concentration (or other measures of PFOA exposure) and various health outcomes. 

Exposures in the highly exposed C8 community are based on the concentrations in contaminated 

drinking water and serum measures. Exposures among the general population typically included 

multiple PFASs as indicated by serum measurements. The correlation among eight carbon 

PFASs is often moderately strong (e.g., Spearman r > 0.6 for PFOA and perfluorooctane 

sulfonate (PFOS) in the general population). Mean serum levels among the occupational cohorts 

ranged from approximately 1 to 4 micrograms per milliliter (μg/mL) and in the C8 cohort ranged 

from 0.01 to 0.10 μg/mL. Geometric mean serum values for the National Health and Nutrition 

Examination Survey (NHANES) general population (≥ age 12; 2003–2008) were 0.0045 μg/mL 

for males and 0.0036 μg/mL for females.

These epidemiology studies have generally found positive associations between serum PFOA 

concentration and total cholesterol (TC) in the PFOA-exposed workers and the high-exposure 

community (i.e., increasing lipid level with increasing PFOA); similar patterns are seen with 

low-density lipoproteins (LDLs) but not with high-density lipoproteins (HDLs). These 

associations were seen in most of the general population studies, but similar results also were 

seen with PFOS, and the studies did not always adjust for these correlations. Associations 

between serum PFOA concentrations and elevations in serum levels of alanine aminotransferase 

(ALT) and gamma-glutamyl transpeptidase (GGT) were consistently observed in occupational 

cohorts, the high-exposure community, and the U.S. general population. The associations are not 

large in magnitude, but indicate the potential for PFOA to affect liver function.

Perfluorooctanoic acid (PFOA) – May 2016 ES-1 
Diagnosed thyroid disease in females and female children was increased both in the highexposure C8 study population and in females with background exposure; thyroid hormones are 

not consistently associated with PFOA concentration. Associations between PFOA exposure 

and risk of infectious diseases (as a marker of immune suppression) were not identified, but a 

decreased response to vaccines in relation to PFOA exposure was reported in studies in adults 

in the high-exposure community population and in studies of children in the general 

population; in the latter studies, it is difficult to distinguish associations with PFOA from those 

of other correlated PFASs. Studies in the high-exposure community reported an association 

between serum PFOA and risk of pregnancy-related hypertension or preeclampsia, conditions 

related to renal function during pregnancy; this outcome has not been examined in other 

populations. An inverse association between maternal PFOA (measured during the second or 

third trimester) or cord blood PFOA concentrations and birth weight was seen in several 

studies. It has been suggested that low glomerular filtration rate (GFR) can impact fetal birth 

weight (Morken et al. 2014). Pharmacokinetic (PK) analyses have shown, however, that in 

individuals with low GFR, there are increased levels of serum PFOA and lower birth weights.

Thus, the impact on body weight is likely due to a combination of the low GFR and the serum 

PFOA.

The epidemiology studies did not find associations between PFOA and diabetes, 

neurodevelopmental effects, or preterm birth and other complications of pregnancy. 

Developmental outcomes including delayed puberty onset in girls also have been reported; 

however, in the two studies examining PFOA exposure in relation to menarche, conflicting 

results were observed: either no association or a possible indication of an earlier menarche seen 

with higher maternal PFOA levels in one study and a later menarche seen with higher maternal 

PFOA levels in the other study. Increased risk of ulcerative colitis was reported in the highexposure community study as well as in a study limited to workers in that population. 

For PFOA, oral animal studies of short-term subchronic and chronic duration are available in 

multiple species including monkeys, rats, and mice. These studies report developmental effects, 

liver and kidney toxicity, immune effects, and cancer (liver, testicular, and pancreatic). 

Developmental effects observed in animals include decreased survival, delayed eye opening and 

reduced ossification, skeletal defects, altered puberty (delayed vaginal opening in females and 

accelerated puberty in males), and altered mammary gland development. 

In most animal studies, changes in relative and/or absolute liver weight appear to be the most 

common effect observed with or without other hepatic indicators of adversity, identifying 

increased liver weight as a common indicator of PFOA exposure. The liver also contains the 

highest levels of PFOA when analyzed after test animal sacrifice. The increases in liver weight 

and hypertrophy, however, also can be associated with activation of cellular peroxisome 

proliferator-activated receptor α (PPARα) receptors, making it difficult to determine if this 

change is a reflection of PPARα activation or an indication of PFOA toxicity. The PPARα 

response is greater in rodents than it is in humans. The U.S. Environmental Protection Agency 

(EPA) evaluated liver disease and liver function resulting from PFOA exposure in studies where 

liver weight changes and other indicators of adversity such as necrosis, inflammation, fibrosis, 

and/or steatosis (fat accumulation in the liver) or increases in liver or serum enzymes indicative 

of liver damage were observed. 

Perfluorooctanoic acid (PFOA) – May 2016 ES-2 
In repeat PFOA dosing studies, rats given 0.64 milligrams per kilogram per day (mg/kg/day) 

for 13 weeks and monkeys given 3 mg/kg/day for 26 weeks had increased liver weight 

accompanied by hepatocellular hypertrophy. As part of a two-generation study, male rats had 

increased liver and kidney weights as well as decreased body weight at 1 mg/kg/day. In shorter 

term studies, slightly higher or lower doses to rats resulted in increased liver weight, liver 

necrosis, and developmental delays. In mice, developmental toxicity and increased spleen weight 

was observed at a dose of 1 mg/kg/day accompanied by increased liver weight. Other doses of 

similar magnitudes in mice were associated with developmental delays and liver necrosis. 

Slightly higher doses resulted in decreased immunoglobulin levels. As supported by the 

epidemiology data, suppression of the immune system in response to PFOA exposure is an area 

of concern for humans as well as animals.

PFOA is known to activate PPAR pathways by increasing transcription of mitochondrial and 

peroxisomal lipid metabolism, sterol, and bile acid biosynthesis and retinol metabolism genes. 

Based on PFOA-induced transcriptional activation of many other genes in PPARα-null mice, 

however, other receptors such as the constitutive androstane receptor (CAR), farnesoid receptor 

(FXR), and pregnane X receptor (PXR) could be involved in PFOA-induced toxicity. 

EPA used a peer-reviewed PK model to calculate the average serum concentrations 

associated with candidate no observed adverse effect levels (NOAELs) and lowest observed 

adverse effect levels (LOAELs) from six studies for multiple effects to calculate corresponding 

human equivalent doses (HEDs) for the derivation of candidate reference doses (RfDs). Overall, 

the toxicity studies available for PFOA demonstrate that the developing fetus is particularly 

sensitive to PFOA-induced toxicity. In addition to the critical developmental effects described 

above, other adverse effects include decreased survival, delays in eye opening and ossification, 

skeletal defects, delayed vaginal opening in females, and altered mammary gland development. 

The EPA Office of Water (OW) selected an RfD of 0.00002 mg/kg/day based on effects 

observed in a developmental toxicity study in mice for PFOA (Lau et al. 2006). The RfD is 

based on reduced ossification and accelerated puberty (in males). The total uncertainty factor 

(UF) applied to the HED LOAEL from Lau et al. (2006) is 300 and includes a UF of 10 for 

intrahuman variability, a UF of 3 to account for toxicodynamic differences between animals and 

humans, and a UF of 10 to account for use of a LOAEL as the point of departure (POD). 

Decreased pup body weights also were observed in studies conducted in mice receiving 

external doses within the same order of magnitude (1, 3, and 5 mg/kg/day, respectively) as those 

chosen for the RfD. These studies, however, lacked serum levels and were not amenable to 

modeling. Overall, the developmental and reproductive toxicity studies available for PFOA 

demonstrate that the developing fetus is particularly sensitive to PFOA-induced toxicity. The 

selected RfD is supported by the other candidate RfDs (also 0.00002 mg/kg/day) based on 

effects on the immune system in a 15-day short-term study by DeWitt et al. (2008) and on the 

kidneys of F0 and F1 males in a two-generation study of developmental and reproductive 

toxicity.

Under EPA’s Guidelines for Carcinogen Risk Assessment (USEPA 2005a), there is 

“suggestive evidence of carcinogenic potential” for PFOA. Epidemiology studies demonstrate an 

association of serum PFOA with kidney and testicular tumors among highly exposed members of 

the general population. Two chronic bioassays of PFOA support a positive finding for its ability 

to be tumorigenic in one or more organs of rats, including the liver, testes, and pancreas. EPA 

estimated a cancer slope factor (CSF) of 0.07 (mg/kg/day)-1 based on testicular tumors. As a 

Perfluorooctanoic acid (PFOA) – May 2016 ES-3 
comparative analysis, the concentration of PFOA in drinking water that would have a one-in-amillion increased cancer risk was calculated using the oral slope factor for testicular tumors, 

assuming a default adult body weight of 80 kg and a default drinking water intake (DWI) rate of 

2.5 liter per day (L/day) (USEPA 2011). This concentration is lower than the concentration for 

cancer (also derived with adult exposure values), indicating that a guideline derived from the 

developmental endpoint will be protective for the cancer endpoint.

Perfluorooctanoic acid (PFOA) – May 2016 ES-4 
1 IDENTITY: CHEMICAL AND PHYSICAL PROPERTIES

Perfluorooctanoic acid (PFOA) is a completely fluorinated organic synthetic acid used to 

produce fluoropolymers. It is manufactured by the Simons electrochemical fluorination (ECF) 

process or by telomerization. In the ECF process, the carbon-hydrogen bonds on molecules of 

organic feedstock are replaced with carbon-fluorine bonds when an electric current is passed

through a solution of hydrogen fluoride and the organic feedstock. In the telomerization process, 

fluorine-bearing chemicals and tetrafluoroethylene react to produce fluorinated intermediates 

that are converted into PFOA (HSDB 2006). The telomerization process produces linear chains 

(Beesoon et al. 2011). Ammonium perfluorooctanoate (APFO) is the salt of PFOA and is a 

processing aid in the manufacture of certain fluoropolymers, especially as an emulsifier in 

aqueous solution during the emulsion polymerization of tetrafluoroethylene (see Figure 1-1). 

APFO is not consumed during the polymerization process (SPI 2005). Some sources of PFOA in 

the atmosphere result from the atmospheric degradation or transformation or surface deposition 

of precursors, including related fluorinated chemicals (e.g., fluorotelomer alcohols, olefins, and 

perfluoroalkyl sulfonamide substances) (Wallington et al. 2006). 

PFOA APFO

Source: SIAR 2006

Figure 1-1. Chemical Structures of PFOA and APFO

Although PFOA is not a polar molecule, each of the carbon fluoride bonds is a dipole as a 

result of the electronegativity difference between fluoride (4.1) and carbon (2.5), placing a partial 

negative charge on each of the covalently bound fluorines and a partial positive charge on each 

of the fluorinated carbons. Charge repulsion of the partially negative fluorines and steric factors 

favor a PFOA conformation in which carbons 2 through 7 adopt an anti arrangement of 

substituents resulting in a fairly linear molecular shape as the lowest energy conformer (see 

Figure 1-2 and Figure 1-3). 

Figure 1-2. PFOA Anti-Conformer 

Perfluorooctanoic acid (PFOA) – May 2016 1-1 
Figure 1-3. PFOA Lowest Energy Conformer

The favored PFOA conformer is very similar to the preferred conformation of the eightcarbon fatty acid, octanoic acid (also known as caprylic acid) except for the sphere of partial 

negative charge on the fluorines of the exterior surface. The ionized carboxylate grouping and 

the fluorine’s partial negative charges favor electrostatic interactions between PFOA and 

positively charged surfaces on proteins and other macromolecules. 

The ECF process produces branched chain isomers, about 80% linear and 20% branched 

(Loveless et al. 2008). The samples studied by Loveless et al. (2008) had the following mole 

(mol) percents of branched chain isomers: 12.6% internal monomethyl (nonalpha), 9% isopropyl, 

0.2%, tert-butyl, 0.1 gem-dimethyl, and 0.1 alpha monomethyl. A study by Yoshikane et al. 

(2010) reported finding perfluoro-6-methylheptanoic acid (the isopropyl isomer) using mass 

spectroscopy analysis of environmental fluorosurfactants in Japan. Branched chain samples 

evaluated by Beesoon and Martin (2015) had a 7 carbon linear chain with methyl groups on 

carbons 3, 4, 5, or 6, designated as 3m, 4m, 5m, or 6m (iso), respectively. The composition of a 

PFOA product is important because the toxicokinetic and physiological properties of the linear 

and branched chain isomers are different. The nomenclature for the branched chain isomers 

varied between authors and indicates that differences exist in the composition of the commercial 

products that were evaluated. 

The physical and chemical properties and other reference information for PFOA and its salt 

APFO are provided in Table 1-1. These properties help to define the behavior of PFOA in living 

systems and the environment. PFOA and its salt are highly stable compounds. They are solids at 

room temperature with low vapor pressures. The melting point for PFOA is identified as 

54.3 degrees Celsius, and vapor pressures increase at temperatures near the melting point. 

PFOA is moderately soluble in water and APFO is even more soluble. Both compounds are 

considered insoluble in nonpolar solvents, which results in their being described as olephobic. 

Water solubility is increased by the presence of other ions and is an important factor governing 

solubility in body fluids. As the concentration of PFOA in aqueous solution increases, it forms 

colloidal micelles with the carboxyl functional groups on the exterior and the fluorocarbon chain 

on the interior. The critical micelle concentration has been identified as 3.6–3.7 g/L. Once the 

critical concentration has been reached, micelles will form and the PFOA molecules will 

colloidally distribute in the aqueous environment. At levels below the critical micelle 

concentration, the individual molecules are individually distributed in the solvent. 

The acid dissociation constant (pKa) for PFOA has been reported as 2.8. As a result, it will 

be present in most biological fluids (gastric secretions excluded) primarily as the 

perfluorooctanoate anion. This is an important feature in governing absorption and membrane 

transport.

Perfluorooctanoic acid (PFOA) – May 2016 1-2 
Table 1-1. Chemical and Physical Properties of PFOA

Property Perfluorooctanoic Acid Source

Chemical Abstracts 

Service Registry No. 

(CASRN) a

335-67-1

CA Index Name 2,2,3,3,4,4,5,5,6,6,7,7,8,8,8-

pentadecafluorooctanoic acid

Synonyms PFOA; Pentadecafluoro-1-octanoic acid; 

Pentadecafluoro-n-octanoic acid; Octanoic 

acid, pentadecafluoro-; Perfluorocaprylic 

acid; Pentadecafluorooctanoic acid; 

Perfluoroheptanecarboxylic acid 

Chemical Formula C8HF15O2

Molecular Weight (g/mol) 414.09 (HSDB 2012); (Lide 2007); (SRC 2016)

Color/Physical State White powder (ammonia salt) (HSDB 2012); (Lewis 2004) 

Boiling Point 192.4 oC; Stable when bound (HSDB 2012); (Lide 2007); (SRC 2016)

Melting Point 54.3 oC (HSDB 2012); (Lide 2007); (SRC 2016)

Vapor Pressure 0.525 mm Hg at 25 oC (measured)

0.962 mm Hg at 59.25 °C (measured)

(Hekster et al. 2003); (HSDB 2012); 

(SRC 2016)

(ATSDR 2015); (Kaiser et al. 2005)

Henry’s Law constant Not measureable (ATSDR 2015)

pKa 2.80 (SRC 2016)

Koc 2.06 (Higgins and Luthy 2006)

Kow Not measurable (ATSDR 2015); (EFSA 2008)

Solubility in water 9.50 x 103 mg/L at 25 oC (estimated) (ATSDR 2015); (Hekster et al. 2003); 

(HSDB 2012); (Kauck and Diesslin 

1951); (SRC 2016)

Half-life in water (25oC) Stable (UNEP 2015)

Half-life in air Stable when bound (UNEP 2015)

Note: aThis CASRN is for linear PFOA, but the toxicity studies are based on a mixture of linear and branched and the RfD 

applies to both.

Perfluorooctanoic acid (PFOA) – May 2016 1-3 
2 TOXICOKINETICS

PFOA is stable to metabolic and environmental degradation because of strong carbonfluorine bonds. It also is resistant to metabolic biotransformation. Thus, the toxicity of the parent 

compound is the concern. Because of its impact on cellular receptors and proteins, it possesses

the ability to impact the biotransformation of dietary constituents, intermediate metabolites, and 

other xenobiotic chemicals by altering enzyme activities and transport kinetics. PFOA is known 

to activate PPAR pathways by increasing transcription of mitochondrial and peroxisomal lipid 

metabolism, sterol, and bile acid biosynthesis and retinol metabolism genes. Based on 

transcriptional activation of many genes in PPARα-null mice, however, the effects of PFOA 

involve far more than activation of PPAR and consequent peroxisome proliferation. The data 

indicate that it also can activate the CAR, FXR, and PXR and metabolic activities linked to these 

nuclear receptors. 

PFOA is not readily eliminated from humans and other primates. Toxicokinetic profiles and 

the underlying mechanism for half-life differences are not completely understood, although 

many of the differences appear to be related to elimination kinetics and factors that control 

membrane transport. Thus far, three transport families appear to play a role in PFOA absorption, 

distribution, and excretion: organic anion transporters (OATs), organic anion transporting 

polypeptides (OATPs), and multidrug resistance-associated proteins (MRPs) (Klaassen and 

Aleksunes 2010; Launay-Vacher et al. 2006). The transporters are critical for gastrointestinal 

absorption, uptake by the tissues, and excretion via bile and the kidney. These transport systems 

are located at the membrane surfaces of the intestines, liver, lungs, heart, blood brain barrier, 

blood placental barrier, blood testes barrier, and mammary glands where they function to protect 

the organs, tissues, and fetus from foreign compounds (Ito and Alcorn 2003; Klaassen and 

Aleksunes 2010, Zaïr et al. 2008). 

There are differences in transporters across species, genders, and individuals. For example, 

more PFOA-specific information is available about the OAT and OATP families than about the 

MRPs. These limitations have hindered the development of PK models for use in predicting 

effects in humans based on the data from animal studies. Abbreviations for the various 

transporters are not totally standardized, and there are inconsistencies across individual 

publications. The current convention for distinguishing between the transporters in humans and 

those in animals is to use uppercase letters for humans and lowercase letters for animals. In this 

document, uppercase letters are used uniformly, thus, the abbreviations indicate the transporter 

family and not the species studied. 

2.1 Absorption

Absorption data are available for oral, inhalation, and dermal exposure in laboratory animals, 

and extensive data are available from humans demonstrating the presence of PFOA in serum. 

These data demonstrate absorption by one or more routes but do not quantify the amounts 

absorbed relative to dose. 

The absorption process requires transport across the interface of the gut, lung, and skin with 

the external environment. Since PFOA is moderately soluble in aqueous solutions and 

oleophobic (i.e., minimally soluble in body lipids), movement across the apical and basal 

membrane surfaces of the lung, gastrointestinal tract, and skin involves transporters or 

mechanisms other than simple diffusion across the lipid bilayer. As discussed above, there are 

Perfluorooctanoic acid (PFOA) – May 2016 2-1 
data that identify involvement of OATs, OATPs, and MRPs in enterocytes in uptake of PFOA 

(Klaassen and Aleksunes 2010; Zaïr et al. 2008). OAT2, OAT3, OATP2b1, and MRP2 are 

located in the apical membrane of the microvilli, and MRP1, 3, and 4 are located along the 

basolateral membrane. Together they function in the uptake of organic anions from 

gastrointestinal contents and transport of those anions into the portal blood supply (Zaïr et al. 

2008). Few studies have been conducted of the intestinal transporters for PFOA in humans or 

laboratory animals. Most of the research has focused on the kidney and has been carried out 

using cultured carrier cells transfected with the transporter proteins. 

2.1.1 Oral Exposure 

Based on animal data, PFOA is well absorbed following oral exposure. Gibson and Johnson 

(1979) administered a single dose of 14C-PFOA averaging 11.4 mg/kg by gavage to groups of 

three male 10-week-old CD rats. Twenty-four hours after administration, at least 93% of total 

carbon-14 was absorbed. In another study, Cui et al. (2010) exposed male Sprague-Dawley rats 

(10 per group) to PFOA (96% active ingredient) at 0, 5, and 20 mg/kg/day once daily by gavage 

for 28 days. The percent of the dose absorbed was 92.8% and 92.3% for the low and high dose, 

respectively, under the assumption that fecal excretion over the first 24 hours after dosing was 

estimated to be unabsorbed material and did not include biliary loss. 

The data from studies of adverse effects on monkeys, rats, and mice receiving PFOA in 

capsules, food, or drinking water demonstrate gastrointestinal absorption. In cynomolgus 

monkeys, steady-state serum PFOA levels were reached within 4–6 weeks after dosing with 

capsules containing 3, 10, and 20 mg/kg PFOA for 6 months (Butenhoff et al. 2004b). Urine 

steady-state levels were reached after 4 weeks. Serum PFOA concentration in male rats fed diets 

containing 0.06, 0.64, 1.94, and 6.5 mg PFOA/kg for 90 days was 7.1, 41, 70, and 138 μg/mL, 

respectively (Perkins et al. 2004). Peak blood levels of PFOA were attained 1–2 hours following 

a 25-mg/kg dose to male and female rats (Kennedy et al. 2004). Blood levels of PFOA over time 

were similar in female rats given a single dose of 25 mg PFOA/kg to a female rat given 10 daily 

doses of 25 mg PFOA/kg (Kennedy et al. 2004). Plasma PFOA concentrations in male SpragueDawley rats fed a diet containing 300 parts per million (ppm) PFOA for 1, 7, and 28 days were 

259, 234, and 252 μg/mL, respectively (Elcombe et al. 2010). 

In rats, a marked gender difference in serum and tissue levels exists following PFOA 

administration. Males consistently have much higher levels than females with the difference 

maintained and becoming more pronounced over time. Female rats show much greater urinary 

excretion of PFOA than do male rats with serum half-life values in hours for females compared 

with days for males. These differences account for variability in postexposure serum 

concentrations between males and females.

2.1.2 Inhalation Exposure

Hinderliter (2003) measured the serum concentrations of PFOA following single and 

repeated inhalation exposures in Sprague-Dawley rats. For the single-exposure study, male and 

female rats (3/gender/group) were exposed to a single nose-only exposure of an aerosol of 0, 1, 

10, and 25 mg/m3 PFOA. Preliminary range-finding studies demonstrated that aerosol particle 

sizes were 1.8–2.0 μm mass median aerodynamic diameter (MMAD) with geometric standard 

deviations (GSDs) ranging from 1.9 to 2.1 μm. Blood samples were collected before exposure; at 

0.5, 1, 3, and 6 hours during exposure; and at 1, 3, 6, 12, 18, and 24 hours after exposure. Plasma 

Perfluorooctanoic acid (PFOA) – May 2016 2-2 
was analyzed by liquid chromatography-mass spectrometry (LC-MS). PFOA plasma 

concentrations increased proportional to aerosol exposure concentrations. 

The male plasma Cmax values were approximately 2–3 times higher than the female Cmax. The 

female Cmax occurred approximately 1 hour after the exposure period with plasma concentrations 

then declining. In males, Cmax was observed immediately after the exposure period ended and 

persisted for up to 6 hours. The data are illustrative of absorption of PFOA via inhalation and are 

consistent with the gender differences in rate of excretion. 

2.1.3 Dermal Exposure

There is evidence that PFOA is absorbed following dermal exposure. Kennedy (1985) treated 

rabbits and rats dermally with a total of 10 applications of PFOA at doses of 0, 20, 200, and 

2,000 mg/kg. Treatment resulted in elevated blood organofluorine levels that increased in a doserelated manner. Organofluorine was measured because, at the time of the study, reliable 

analytical techniques for measuring serum or plasma PFOA were still under development. 

O’Malley and Ebbens (1981) treated groups of two male and two female New Zealand White 

rabbits dermally with doses of 100, 1,000, and 2,000 mg/kg PFOA for 14 days. Mortality among 

the exposed animals demonstrated dermal uptake. All of the animals died at the highest dose, 

three of four died in the mid-dose group, and none in the low-dose group. Although these data 

demonstrate dermal absorption, they do not provide quantitative dose-response data for effects 

other than mortality. 

The results of in vitro percutaneous absorption studies of PFOA through rat and human skin 

have been reported (Fasano et al. 2005). The permeability coefficient for PFOA was calculated 

to be 3.25 ± 1.51 x 10-5 centimeters per hour (cm/h) and 9.49 ± 2.86 x 10-7 cm/h in rat and 

human skin, respectively. 

2.2 Distribution

Distribution of absorbed material requires vascular transport from the portal of entry to 

receiving tissues. It has been suggested that PFOA circulates in the body by noncovalently 

binding to plasma proteins. Several studies have investigated the binding of PFOA to plasma 

proteins in rats, humans, or monkeys to gain an understanding of its absorption, distribution, and 

elimination as well as information on species and gender differences.

Protein Binding. Protein binding in plasma from cynomolgus monkeys, rats, and humans was 

tested with PFOA via in vitro methods (Kerstner-Wood et al. 2003). The results are summarized 

in Table 2-1. Rat, human, and monkey plasma proteins were able to bind 97–100% of the PFOA 

added at concentrations ranging from 1 to 500 ppm. Human serum albumin (HSA) carried the 

largest portion of the PFOA among the protein components of human plasma. Serum albumin is 

a common carrier of hydrophobic materials in the blood, including short- and medium-chain 

fatty acids, thyroxine (T4), heme, inorganic ions, and some pharmaceuticals (Fasano et al. 2005). 

Approximately 60% of the serum protein in humans and rats is albumin (Harkness and Wagner 

1983; Saladin 2004). At 68%, the percentage bound to albumin in mice is slightly higher than in 

humans and rats (Harkness and Wagner 1983). 

Perfluorooctanoic acid (PFOA) – May 2016 2-3 
Table 2-1. Protein Binding in Rat, Human, and Monkey Plasma

PFOA Concentration

(ppm) Rat (%) Monkey (%) Human (%)

1 ~100 ~100 ~100

10 99.5 99.8 99.9

100 98.6 99.8 99.9

250 97.6 99.8 99.6

500 97.3 99.5 99.4

Source: Kerstner-Wood et al. 2003

Note: % binding values reported as “~100” reflect a nonquantifiable amount of test article in the plasma water below the 

quantifiable limit <6.25 ng/mL. 

Han et al. (2003) investigated the binding of PFOA to rat and human plasma proteins in vitro. 

The authors concluded that there was no correlation between the PFOA persistence and binding 

of the PFOA to rat serum. The primary PFOA binding protein in plasma was serum albumin. 

However, the method used (ligand blotting) would not theoretically allow the identification of 

low-abundance proteins with high affinity for PFOA. Further investigation of purified rodent and 

HSA binding using labeled 19F nuclear magnetic resonance (NMR) allowed the calculation of 

disassociation constants for PFOA binding to rodent and HSA. No significant difference in 

binding to the serum albumin of rat versus human was detected (Table 2-2).

Male and female rats treated in vivo showed no gender difference in the binding of PFOA to 

serum proteins though the persistence of PFOA in vivo is much greater in male than female rats.

Table 2-2. Dissociation Constants (Kd) of Binding Between PFOA and Albumin

Parameter Method RSA HSA

Kd (mM) NMRa 0.29 ± 0.10c

Kd (mM) micro-SECb 0.36 ± 0.08c 0.38 ± 0.04

Number of Binding Sites micro-SECb 7.8 ± 1.5 7.2 ± 1.3

Source: Han et al. 2003

Notes:

RSA = rodent serum albumin; HSA= human serum albumin

a = Average of the two Kd values (0.31 ± 0.15 and 0.27 ± 0.05 mM) obtained by NMR.

b = Values were obtained from three independent experiments and their SDs are shown.

c = On the basis of the result of unpaired t-test at 95% confidence interval, the difference of Kd values determined by NMR and 

micro-SEC is statistically insignificant.

Wu et al. (2009) examined the interaction of PFOA and HSA. The authors tested their 

hypothesis that PFOA, after absorption, was transported bound to albumin by dialyzing PFOA 

solutions in the presence and absence of HSA. In the absence of HSA, 98% of the dissolved 

PFOA crossed the dialysis membrane into the dialysate within 4 hours. In the presence of HSA, 

the amount of PFOA found in the dialysate after 4 hours decreased in direct proportion to the 

albumin concentration, demonstrating binding to the protein. No albumin was identified in the 

dialysate. 

Using the dialysis data and thermodynamic considerations, the authors concluded that 

albumin could bind up to 12 PFOA molecules on its surface via chemical monolayer absorption 

with a 13th molecule bound noncovalently in the more hydrophobic interior of the protein. The 

surface nature of the binding could well indicate potential binding to other serum proteins as 

well. Circular dichroism measurements of the albumin/PFOA complex suggested a 

conformational change in the protein as a result of the PFOA binding. The beta-pleated sheet 

Perfluorooctanoic acid (PFOA) – May 2016 2-4 
content of the albumin decreased, and the alpha-helix content increased by 15%. These 

conformational changes could interfere with the functional properties of serum albumin or other 

serum proteins impacted by surface monolayers of PFOA. For example, albumin’s ability to 

transport its natural ligands could be decreased by the presence of PFOA on the protein surface. 

The interaction of albumin with target cellular receptors also could be altered. 

MacManus-Spencer et al. (2010) used a variety of approaches to quantify the binding of 

PFOA to serum albumin (e.g., surface tension measurements, 19FNMR spectroscopy, and 

fluorescence spectroscopy). When taken together, the results from these analyses suggest the 

presence of primary and secondary binding sites on albumin. The PFOA-albumin association 

constants for the primary site (K1

a

) was about 1.5 ± 0.2 x 105/mol bovine serum albumin (BSA) 

while the association constant for the secondary site (K2

a

) was 0.8 ± 0.1 x 102/mol BSA at a 

concentration of 1μmol. When the concentration of BSA increased to 10 μmol, the binding per 

mol of BSA decreased K1

a

 = 0.33 ± 0.004 and K2

a

 = 0.53 ± 0.1. Qin et al. (2010) also used 

fluorescence spectroscopy quenching analysis to study PFOA binding to BSA and concluded that 

van der Waals forces and hydrogen bonds were the dominant intermolecular binding forces. 

The results of the fluorescence spectroscopy suggested a conformational change in BSA 

following binding of PFOA that moved a tryptophan residue (#214) from a slightly polar region 

of the protein to a less polar region. The shift in a tryptophan position is consistent with the 

observations of Wu et al. (2009) and Qin et al. (2010), who reported that BSA underwent a 

conformational change following the binding of PFOA. The authors considered the results from 

the fluorescence spectroscopy to be relevant to the potential physiological impact of PFOA at 

levels found in the environment. Because serum albumin is a carrier for a variety of endogenous 

and exogenous substrates, a change in conformation can alter the bonding constants between 

albumin and other serum constituents. 

A modeling study by Salvalaglio et al. (2010) was conducted to determine the binding sites 

of PFOA on HSA and classify them by their interaction energy using molecular modeling; this 

study builds on the binding studies of Wu et al. (2009) and MacManus-Spencer et al. (2010). It 

was estimated that the maximum number of PFOA binding sites on HSA was nine. The site 

locations were common to the natural binding sites for fatty acids, T4, Warfarin, indol, and 

benzodiazepine (see Figure 2-1) (Salvalaglio et al. 2010). The binding site closest to tryptophan 

residue #214 had the highest binding affinity (-8.0 kilocalorie/mol).

Figure 2-1. PFOA Binding Sites on HSA

Perfluorooctanoic acid (PFOA) – May 2016 2-5 
Weiss et al. (2009) screened several perfluorinated compounds (n = 30), differing by carbon 

chain length C4–18, fluorination degree, and functional groups for potential binding to the serum 

thyroid hormone transport protein, transthyretin (TTR), using a radioligand-binding assay. The 

natural ligand of TTR is T4. PFOA was one of the chemicals evaluated. Human TTR was 

incubated overnight with 125I-labeled T4, unlabeled T4, and 10-10,000 nanomolar (nmol) PFOA 

as a competitor for the T4 binding sites. The unlabeled T4 was used as a reference compound. 

The levels of T4 in the assay were close to the lower range for total T4 measured in healthy 

adults. The authors concluded that binding affinity for TTR was highest for the fully fluorinated 

compounds tested and those having at least an eight carbon length chain, characteristics that 

apply to PFOA. PFOA demonstrated a high binding affinity for TTR with 949 nmol, causing a 

50% inhibition of T4 binding to the TTR. 

Beesoon and Martin (2015) examined differences in the binding of the linear and branched 

chain isomers to serum albumin and human serum proteins. The linear PFOA molecule was 

found to bind more strongly to calf serum albumin than the branched chain isomers. When 

arranged in order of increasing binding, the order was 4m < 3m <5m < 6m (iso) <linear. In the 

isomer-specific binding to spiked total human serum protein, the linear molecule clearly had the 

strongest binding potential with about 7–10% free. The relationship for the other isomers was 5m 

> 6m >4m> 3m (15–30% free). Binding was estimated based on the concentrations in the 

ultrafiltrate after spiking with 5–60 mg/L of technical PFOA. The human serum was diluted 

tenfold before spiking. 

When incubated with separate human-derived plasma protein fractions (Kerstner-Wood et al. 

2003), PFOA was highly bound (99.7%) to albumin and showed some affinity for LDLs, 

formerly beta-lipoproteins (9.6%) with limited binding to alpha-globulins (11.0%) and gammaglobulins (3.0%). Low levels of binding to alpha-2-macroglobulin and transferrin were measured 

when the protein concentrations were approximately 10% of physiological concentration (see 

Table 2-3).

Table 2-3. Percent (%) Binding of PFOA to Human Plasma Protein Fractions

Fraction ~10% Physiological Conc. 100% Physiological Conc.

Albumin 96.4 99.7

Gamma-globulin 3.5 3.0

Alpha-globulin 28.5 11.0

Fibrinogen 5.4 <0.1

Alpha-2-macroglobulin 7.9 <0.1

Transferrin 1.0 2.1

LDLs 19.6 39.6

Source: Kerstner-Wood et al. 2003 

It also is possible that PFOA will display nonspecific binding to proteins within the cellular 

matrix as well as in the serum but little work has been done to investigate that probability. 

Luebker et al. (2002) conducted in vitro studies of the ability of a variety of perfluorinated 

chemicals to displace a fluorescent substrate (11-(5-dimethylamino-napthalenesulphonyl)-

undecanoic acid) from liver fatty acid binding protein (L-FABP). L-FABP is an intracellular 

lipid carrier protein that reversibly binds long-chain fatty acids, phospholipids, and an assortment 

of peroxisome proliferators (Erol et al. 2003). It constitutes 2–5% of the cytosolic protein in 

hepatocytes. Luebker et al. (2002) reported that PFOA (IC50>10μmol) exhibited some binding to 

Perfluorooctanoic acid (PFOA) – May 2016 2-6 
L-FABP, but the binding potential was only about 50% of that for PFOS (IC50 = 4.9 μmol) and 

far less than that of oleic acid (IC50 = 0.01 μmol).

L. Zhang et al. (2013) cloned the human L-FABP gene and used it to produce purified 

protein for evaluation of the binding of PFOA and other PFASs. Nitrobenzoxadiazole-labeled 

lauric acid was the fluorescent substrate used in the displacement assays. IC50 values and 

dissociation constants were generated for the PFASs studied. Oleic and palmitic acids served as 

the normal substrates for L-FABP binding. The nitrobenzoxadiazole-labeled lauric acids 

indicated that there were two distinct binding sites for fatty acids in human FABP with the 

primary site having a twentyfold higher affinity than the secondary site. The IC50 value for 

PFOA was 9.0 ± 0.7 μmol, suggesting that it has a lower binding affinity than PFOS (IC50=3.3 ± 

0.1 μmol). A similar approach was used to compare perfluorohexanoic acid (PFHxA, 

perfluorohexanesulfonic acid (PFHxS), PFOA, and perfluorononanoic acid (PFNA). The affinity 

of PFNA for human L-FABP was found to be greater than that for PFOA. The affinities of 

PFHxA and PFHxS for the protein were much lower. Both PFOA and PFNA bound to the carrier 

protein in a 1:1 ratio and the interaction was mediated by electrostatic interactions and hydrogen 

binding of the PFAS with the fatty acid binding site.

2.2.1 Oral Exposure 

Tissue Distribution

Human. No clinical studies are available that examined tissue distribution in humans following

administration of a controlled dose of PFOA. However, samples collected in biomonitoring and 

epidemiology studies provide data showing distribution of PFOA. Olsen et al. (2001a) analyzed 

human sera and postmortem liver samples and found that more than 90% of the liver samples 

(n = 30) were < limit of quantitation (LOQ). Serum levels ranged from <LOQ–7.0 nanograms 

per milliliter (ng/mL). PFOA concentrations above the LOQ were detected in 5/6 postmortem 

liver samples from males in Catalonia, Spain. In females, only 1/6 liver samples was above LOQ 

(Kärrman et al. 2009). Pirali et al. (2009) measured intrathyroidal PFOA levels (0.4–6.0 ng/g) in 

thyroid surgical patients and found no correlation between serum and thyroid PFOA 

concentration. PFOA has been detected in breast milk samples (Tao et al. 2008; Völkel et al. 

2008) and cord blood samples (Apelberg et al. 2007; Monroy et al. 2008) at concentrations 

above the LOQ. These studies indicate that PFOA is distributed within the body and that 

maternal transfer to offspring can occur.

Pérez et al. (2013) collected tissue samples from 20 adult subjects (aged 28–83) who had 

been living in Catalonia, Spain, for 10 years and died of a variety of causes. Autopsies and tissue 

collection (liver, kidney, brain, lung, and bone) were carried out in the first 24 hours after death. 

The tissues were analyzed for 21 perfluorinated compounds. PFOA was present in 45% of the 

samples but could be quantified in only 20% (median 1.9 ng/g). PFOA accumulated primarily in 

the bone (60.2 ng/g), lung (29.2 ng/g), liver (13.6 ng/g), and kidney (2.0 ng/g), with levels below 

detection in brain based on the mean wet weight tissue concentrations. Detection levels varied 

with the tissue evaluated.

Animal. Studies of tissue distributions are available for several species including monkeys, rats, 

and mice. The data are categorized by species in the sections that follow. The distribution data 

derived from studies during pregnancy and lactation follow the data on nonpregnant animals. 

Perfluorooctanoic acid (PFOA) – May 2016 2-7 
Monkey. Butenhoff et al. (2002, 2004b) studied the fate of PFOA in cynomolgus monkeys in a 

6-month oral exposure study. Groups of four to six male monkeys each were administered PFOA 

daily via oral capsule at dose rates (DRs) of 0, 3, 10, or 20 mg/kg. The highest dose was initially 

30 mg/kg, but due to its toxicity, it was suspended after 12 days. Dosing was resumed on test day 

22 using the 20 mg/kg/day dose for the remainder of the 6-month period, resulting in a 

normalized dose of 20 mg/kg/day for the study. Serum, urine, and fecal samples were collected 

at 2-week intervals and were analyzed for PFOA concentrations. Liver samples were collected at 

time of sacrifice.

Serum concentrations reached steady-state levels within 4–6 weeks in all dose groups. 

Steady-state concentrations of PFOA in serum were 77 ± 39, 86 ± 33, and 158 ± 100 μg/ml after 

6 weeks and 81 ± 40, 99 ± 50, and 156 ± 103 μg/ml after 6 months for the 3-, 10-, and 20-mg/kg 

dose groups, respectively (Butenhoff et al. 2002, 2004b). The mean serum concentration of 

PFOA in control monkeys was 0.134–0.203 μg/ml. Urine PFOA concentrations reached steady

state after 4 weeks and were 53 ± 25, 166 ± 83, and 181 ± 100 μg/ml in the 3, 10, and 20-mg/kg 

dose groups, respectively, for the duration of the study. Liver PFOA concentrations at terminal 

sacrifice in the 3-mg/kg and 10-mg/kg dose groups were similar and ranged from 6.29 to 

21.9μg/g. Liver PFOA concentrations in two monkeys exposed to 20 mg/kg were 16.0 and 

83.3 μg/g. Liver PFOA concentrations in two monkeys dosed with 10 mg/kg/day at the end of a 

13-week recovery period were 0.08 and 0.15 μg/g (Butenhoff et al. 2004b). 

Rat. Ylinen et al. (1990) administered newly weaned Wistar rats (18/gender/group) doses of 3, 

10, and 30 mg/kg/day PFOA by gavage for 28 days. At necropsy, serum was collected as well as 

brain, liver, kidney, lung, spleen, ovary, testis, and adipose tissue. The concentration of PFOA in 

the serum and tissues was determined with capillary gas chromatography equipped with a flame 

ionization detector (FID). A mass spectrometer was used in the selected ion monitoring mode 

when the PFOA concentration was below the LOQ of the FID (1 μg/ml).

The concentration of PFOA in the serum and tissues following 28 days of administration is 

presented in Table 2-4. PFOA was not detected in the adipose tissue. The concentrations of 

PFOA in the serum and tissues were much higher in males than in females. In the males, the 

levels of PFOA in the serum were generally lower in the 30 mg/kg/day dose group than in the 

10 mg/kg/day dose group due to increased urinary elimination in the 30 mg/kg/day group. The 

tissue levels were similar for the 10 and 30 mg/kg/day doses. In females, there was a dose-related 

increase in tissue levels while the serum levels were comparable for the 10 and 30 mg/kg/day 

dose groups. Among solid tissues, the liver had the highest tissue concentration followed by the 

testis, spleen, lung, kidney, and brain, respectively. In females, the concentration in the kidneys 

exceeded that in the liver for the 10 and 30 mg/kg/day doses but not at the lowest dose. Ovary 

and spleen tissue had similar concentrations followed by lower levels in the lung and brain. 

Perfluorooctanoic acid (PFOA) – May 2016 2-8 
Table 2-4. Tissue Distribution of PFOA in Wistar Rats After 28 Days of Treatment

Tissue

Dose (Malesa)

mg/kg/day

Dose (Femalesa)

mg/kg/day

3 10 30 3 10 30

Serum 48.6 ± 10.3 87.27 ± 20.09 51.65 ± 1.47 2.4b 12.47 ± 4.07 13.92 ± 6.06

Liver 39.9 ± 7.25 51.71 ± 11.18 49.77 ± 10.76 1.81 ± 0.49 3.45 ± 1.36 6.64 ± 2.64

Kidney 1.55 ± 0.71 40.56 ± 14.94 39.81 ± 17.67 0.06 ± 0.02 7.36 ± 3.19 12.54 ± 8.24

Spleen 4.75 ± 1.66 7.59 ± 3.5 4.1 ± 1.57 0.15 ± 0.04 0.38 ± 0.17 1.59 ± 0.49

Lung 2.95 ± 0.54 22.58 ± 4.59 23.71 ± 5.42 0.24b 0.22 ± 0.15 0.75 ± 0.26

Brain 0.398 ± 0.144 1.464 ± 0.211 0.71 ± 0.32 < LOQc 0.029 ± 0.019 0.044 ± 0.018

Ovary < LOQ 0.41 ± 0.27 1.16 ± 0.58

Testis 6.24 ± 2.04 9.35 ± 4.02 7.22 ± 3.17

Source: Ylinen et al. 1990

Notes:

a n = 6, mean ± SD, μg/ml tissue. b n = 3, no SD. c

 Below LOQ of 1μg/mL

Kemper (2003) examined the distribution of PFOA in tissues of male and female SpragueDawley rats following administration by gavage. Rats were administered 1, 5, and 25 mg/kg 14CPFOA by oral gavage. Tissue concentrations were determined at the time of maximum plasma 

concentration (Tmax) and at the time that plasma concentration had fallen to one half the 

maximum (Tmax/2). Values for Tmax and Tmax/2 for male and female rats were determined from PK 

experiments. In those experiments, plasma was collected over the course of several days and 

PFOA concentration was analyzed. Noncompartmental PK models were applied to identify Tmax

and elimination half-time (T1/2) from the data. The Tmax/2 was calculated as the time (hr) for the 

maximum plasma concentration plus the elimination half-time (hr) (Tmax + T1/2). In some cases, 

elimination could occur in a rapid phase followed by a slower elimination phase. For cases in 

which biphasic elimination was evident, the rapid phase T1/2 was used for calculation of Tmax/2. 

Tissues from male rats were collected at 10.5 hours (Tmax) and 171 hours (Tmax/2) after 

dosing. Tissues from female rats were collected at 1.25 hours (Tmax) and 4 hours (Tmax/2) after 

dosing. The results are summarized in Table 2-5 for males and Table 2-6 for females. Liver, 

blood, skin, muscle, bone, G.I. tract, and fat were the primary tissues for distribution of 

14C-PFOA. In males, the fraction of the dose found in the liver increased between Tmax and 

Tmax/2, but remained constant or decreased in other tissues. In females, the fraction of the dose 

present in all tissues remained constant or decreased between Tmax and Tmax/2. Liver-to-blood 

concentration ratios for 14C at Tmax in males were greater than 1 and increased between Tmax and 

Tmax/2. In females distribution levels in blood were between 1 and 2 at all dose levels and 

remained relatively constant between Tmax and Tmax/2. In males, the blood distributions levels 

were tenfold or higher than kidney levels at Tmax and declined slightly at Tmax/2. 

Examination of the residuals from the administered PFOA in the male tissues at Tmax/2 

(171 hours) indicate that 40–60% of the dosed PFOA retained was present in the liver, blood, 

skin, and muscle tissues in decreasing amounts (Table 2-5). In males, about 1% of the label was 

present in the gastrointestinal tissues and contents at Tmax/2, while the value for females was about 

10%. However, the samples were collected at 1.25 and 4 hours in females and 10.5 and 171 

hours in males, providing more time for absorption in the males. 

Perfluorooctanoic acid (PFOA) – May 2016 2-9 
Table 2-5. Distribution of PFOA in Male Sprague-Dawley Rats After Oral Exposure Dose

Tissue

1 mg/kg 5 mg/kg 25 mg/kg

% at Tmax % at Tmax/2 % at Tmax % at Tmax/2 % at Tmax % at Tmax/2

Prostate 0.083 ± 0.039 0.030 ± 0.002 0.071 ± 0.045 0.057 ± 0.020 0.067 ± 0.018 0.028 ± 0.012

Skina 14.77 ± 2.135 6.061 ± 0.274 15.565 ± 0.899 7.233 ± 0.430 13.836 ± 0.969 5.419 ± 0.237

Blooda 22.148 ± 0.692 8.232 ± 1.218 24.919 ± 1.942 11.140 ± 0.624 22.905 ± 1.177 7.904 ± 1.032

Brain 0.071 ± 0.018 0.022 ± 0.002 0.051 ± 0.021 0.023 ± 0.008 0.063 ± 0.007 0.019 ± 0.002

Fata 2.281 ± 0.467 0.593 ± 0.136 2.815 ± 0.225 0.916 ± 0.205 2.153 ± 0.430 0.628 ± 0.110

Heart 0.451 ± 0.119 0.195 ± 0.024 0.443 ± 0.037 0.252 ± 0.030 0.461 ± 0.053 0.164 ± 0.032

Lungs 0.74 ± 0.147 0.341 ± 0.043 0.593 ± 0.376 0.344 ± 0.194 0.863 ± 0.103 0.303 ± 0.057

Spleen 0.086 ± 0.011 0.045 ± 0.006 0.096 ± 0.017 0.060 ± 0.007 0.106 ± 0.015 0.042 ± 0.005

Liver 21.708 ± 5.627 32.627 ± 3.601 18.750 ± 2.434 25.231 ± 1.289 17.528 ± 0.900 20.145 ± 3.098

Kidney 1.949 ± 0.402 1.14 ± 0.215 2.170 ± 0.354 1.212 ± 0.115 2.293 ± 0.286 1.003 ± 0.122

G.I. tract 2.930 ± 0.929 0.980 ± 0.300 2.508 ± 0.713 1.052 ± 0.202 2.784 ± 0.608 0.808 ± 0.189

G.I. 

contents

2.083 ± 0.625 0.239 ± 0.025 2.632 ± 0.934 0.270 ± 0.028 4.186 ± 1.349 0.210 ± 0.084

Thyroid 0.008 ± 0.005 0.004 ± 0.003 0.011 ± 0.006 0.004 ± 0.002 0.009 ± 0.002 0.005 ± 0.001

Thymus 0.085 ± 0.008 0.051 ± 0.018 0.085 ± 0.012 0.053 ± 0.003 0.120 ± 0.025 0.045 ± 0.010

Testes 0.755 ± 0.079 0.356 ± 0.037 0.693 ± 0.180 0.372 ± 0.062 0.623 ± 0.098 0.224 ± 0.031

Adrenals 0.019 ± 0.004 0.010 ± 0.001 0.022 ± 0.004 0.009 ± 0.001 0.026 ± 0.004 0.009 ± 0.003

Musclea 12.025 ± 0.648 4.984 ± 0.745 13.565 ± 0.576 6.429 ± 0.648 12.855 ± 0.841 4.253 ± 0.358

Bonea 3.273 ± 0.538 1.120 ± 0.094 3.047 ± 0.544 1.375 ± 0.169 3.062 ± 0.438 0.906 ± 0.100

Totalb 85.465 ± 6.426 57.026 ± 3.379 88.033 ± 1.420 56.031 ± 1.025 83.937 ± 3.680 42.112 ± 4.740

Source: Kemper 2003

Notes: Percent of dose recovered at Tmax and Tmax/2 in tissues. a Percent recovery scaled to whole animal assuming the following: skin=19%, whole blood=7.4%, fat=7%, muscle=40.4%, 

bone=7.3% of body weight.

b Totals are calculated from individual animal data.

In the female tissues at Tmax/2 (4 hours), approximately 30% of the dosed PFOA retained was 

present in the liver, blood, kidney, muscle, and skin tissues in decreasing amounts (Table 2-6). 

About 14% of the administered dose remained in the gastrointestinal tissues and contents. Based 

on the timing of the measurements and the results, females appear to absorb and excrete PFOA 

more rapidly than males.

Lau et al. (2006) studied PFOA’s toxicokinetic properties in rats as part of a larger study. The 

authors gavage-dosed adult male and female Sprague-Dawley rats (n = 8) with 10 mg/kg for 20 

days and sacrificed them 24 hours after the last treatment. After 20 days of treatment, male rats 

had serum PFOA levels of 111 μg/mL compared to 0.69 μg/mL in female rats. 

Martin et al. (2007) administered 20 mg PFOA/kg to adult male Sprague-Dawley rats (n = 4 

or 5) for 1, 3, and 5 days by oral gavage and determined the liver and serum levels of PFOA. 

Blood was collected via cardiac puncture and PFOA concentration was determined by highperformance liquid chromatography-electrospray tandem mass spectrometry (LC-MS/MS). The 

mean liver PFOA concentration was 92 ± 6, 250 ± 32, and 243 ± 23 μg/g after 1, 3, and 5 daily 

doses of 20 mg PFOA/kg/day, respectively. The mean serum concentration was 245 ± 41 μg/mL 

after 3 daily doses of 20 mg PFOA/kg/day. Serum PFOA concentration was not determined after 

1 day and 5 days of dosing due to sample availability. 

Perfluorooctanoic acid (PFOA) – May 2016 2-10
Table 2-6. Distribution of PFOA in Female Sprague-Dawley Rats after

Oral Exposure Dose

Tissue

1 mg/kg 5 mg/kg 25 mg/kg

% at Tmax % at Tmax/2 % at Tmax % at Tmax/2 % at Tmax % at Tmax/2

Skina 0.434 ± 0.162 0.403 ± 0.096 0.624 ± 0.142 0.307 ± 0.121 0.380 ± 0.166 0.415 ± 0.175

Blooda 5.740 ± 1.507 4.438 ± 1.625 8.089 ± 2.080 5.411 ± 1.466 7.158 ± 2.232 6.407 ± 1.406

Brain 0.037 ± 0.009 0.047 ± 0.008 0.066 ± 0.019 0.045 ± 0.010 0.058 ± 0.008 0.058 ± 0.018

Fata 0.134 ± 0.032 0.164 ± 0.079 0.220 ± 0.111 0.110 ± 0.069 0.147 ± 0.053 0.148 ± 0.065

Heart 0.198 ± 0.079 0.253 ± 0.055 0.388 ± 0.057 0.236 ± 0.051 0.317 ± 0.035 0.287 ± 0.069

Lungs 0.454 ± 0.148 0.546 ± 0.082 0.827 ± 0.102 0.570 ± 0.179 0.678 ± 0.067 0.775 ± 0.204

Spleen 0.063 ± 0.027 0.058 ± 0.006 0.101 ± 0.021 0.060 ± 0.012 0.091 ± 0.007 0.070 ± 0.002

Liver 7.060 ± 1.266 6.817 ± 1.537 11.190 ± 2.192 7.176 ± 0.982 10.538 ± 1.723 9.080 ± 0.895

Kidney 3.288 ± 0.948 2.769 ± 0.784 4.293 ± 0.771 2.685 ± 0.736 5.867 ± 0.946 4.749 ± 0.393

G.I. tract 10.699 ± 9.066 8.462 ± 6.519 7.142 ± 2.594 8.255 ± 8.967 6.923 ± 1.846 3.547 ± 1.306

G.I. 

contents

21.956 ± 13.48 3.891 ± 2.395 2.896 ± 2.305 5.601 ± 6.165 2.491 ± 1.548 1.121 ± 1.010

Thyroid 0.010 ± 0.003 0.016 ± 0.021 0.008 ± 0.002 0.006 ± 0.002 0.009 ± 0.003 0.007 ± 0.002

Thymus 0.052 ± 0.017 0.058 ± 0.024 0.105 ± 0.030 0.068 ± 0.021 0.091 ± 0.032 0.077 ± 0.020

Ovaries 0.047 ± 0.019 0.048 ± 0.006 0.071 ± 0.012 0.041 ± 0.012 0.071 ± 0.012 0.070 ± 0.012

Adrenals 0.014 ± 0.005 0.018 ± 0.004 0.026 ± 0.005 0.015 ± 0.004 0.031 ± 0.005 0.021 ± 0.001

Musclea 0.170 ± 0.051 0.258 ± 0.089 0.325 ± 0.010 0.229 ± 0.031 0.441 ± 0.116 0.304 ± 0.099

Uterus 0.243 ± 0.091 0.374 ± 0.247 0.354 ± 0.046 0.247 ± 0.068 0.358 ± 0.124 0.365 ± 0.029

Bonea 0.101 ± 0.017 0.153 ± 0.052 0.174 ± 0.057 0.142 ± 0.078 0.157 ± 0.072 0.181 ± 0.090

Totalb 50.698 ± 16.48 28.772 ± 10.98 36.897 ± 3.187 31.201 ± 12.63 35.803 ± 2.554 27.680 ± 2.569

Source: Kemper 2003

Notes: Percent of dose recovered at Tmax and Tmax/2 in tissues. a Percent recovery scaled to whole animal assuming the following: skin=19%, whole blood=7.4%, fat=7%, muscle=40.4%, 

bone=7.3% of body weight.

b Totals are calculated from individual animal data.

Mouse. Lau et al. (2006) gavage-dosed adult male and female CD-1 mice (5–7/group) with 

20 mg/kg for 7 and 17 days. The animals were sacrificed 24 hours after the last treatment. After 

7 days of treatment, male mice had serum PFOA levels of 181 μg/mL and females had levels of 

178 μg/mL. After 17 days of treatment, male mice had serum PFOA levels of 199 μg/mL and 

females had levels of 171 μg/mL. These data suggest that the gender difference observed by Lau 

et al (2006) in rats was not seen in the mice under the conditions of this study. 

As part of a physiologically based pharmacokinetic (PBPK) modeling exercise, Lou et al. 

(2009) administered single doses of 1 and 10 mg/kg to groups of three male and three female 

CD-1 mice. The mice were sacrificed for analysis of plasma, liver, and kidney tissues after 4, 8, 

and 12 hours and at 1, 3, 6, 9, 13, 20, 27, 34, and 48 days after dosing. This study was repeated 

for a second analysis that extended the sacrifice times to 55, 62, 70, and 80 days. 

Measures of PFOA in serum were presented graphically and indicate that the order of 

magnitude difference between the doses led to a comparable order of magnitude difference in

serum concentrations for both males and females across the 80-day observation period. [The 

study procedures indicated that serum was collected and analyzed, but the graphic presentation 

described the values as plasma values. Contact with one of the authors confirmed that the values 

should have been listed as serum rather than plasma.] The peak serum concentrations were 

Perfluorooctanoic acid (PFOA) – May 2016 2-11
10 and 100 mg/L for the 1 and 10 mg/kg/day doses, respectively. Declines in serum 

concentrations for females over time were roughly parallel reaching concentrations of about 

2 mg/L and <0.2 mg/L for the high and low doses, respectively, at the end of 80 days. Peak 

serum concentrations were slightly lower in the males (~8 and 80 mg/L) than in the females, and 

final serum concentrations were higher in the males (~0.5 and 8 mg/L). Liver and kidney 

concentrations also were higher in males than in females for each of the two doses. These data 

suggest a longer half-life in males than in females. 

Lou et al. (2009) also collected serum data for up to 28 days after administration of a 

60-mg/kg dose to groups of three female mice. Based on the graphic presentation of the data, the 

60-mg/kg dose was cleared from the serum much more rapidly than the 1- and 10-mg/kg doses. 

For example, a serum concentration of about 0.4 mg/L was reached in about 28 days for the 

60-mg/kg dose, 61 days for the 10-mg/kg dose, and 70 days for the 1-mg/kg dose (values 

estimated from Figure 3 in Lou et al. [2009]). No measurements were made for liver or kidney in

the high-dose animals.

In the final experimental portion of the study, Lou et al. (2009) exposed groups of five 

female CD-1 mice to 20 mg/kg/day for 17 days. Serum samples were collected 24 hours after the 

final dose and analyzed for PFOA. The mean serum concentration was 130 ± 23 mg/L, which is 

comparable to that of 171 μg/mL reported by Lau et al. (2006). 

Minata et al. (2010) orally administered 0, 12.5, 25, and 50 micromole per kilogram 

(μmol/kg) PFOA (~0, 5.4, 10.8, and 21.6 mg/kg PFOA) to groups of male wild-type 

129S4/SvlmJ mice (n = 39) and PPARα- null 129S4/SvJae-Pparαtm1Gonz/J mice (n = 40) for 

4 weeks. Blood, liver, and bile were collected for determination of PFOA concentration at the 

end of 4 weeks, as shown in Table 2-7. The PFOA concentration in whole blood and the liver 

were similar between wild-type and PPARα-null mice at the same dose level and appeared to 

increase in proportion to dose. In bile, PFOA concentration in wild-type mice increased by a 

factor of 13.8 from 12.5 to 25 μmol/kg and by a factor of 2.8 from 25 to 50 μmol/kg. In the bile 

of PPARα-null mice, PFOA concentration increased by a factor of 3.2 from 12.5 to 25 μmol/kg 

and by a factor of 19.5 from 25 to 50 μmol/kg. The data suggested saturation of PFOA transport 

from the liver to bile ducts in wild-type mice, but not PPARα-null mice. This may indicate that 

PPARα plays a role in the clearance of PFOA. 

Table 2-7. PFOA Concentrations in Wild-type and PPARα-null Mice (μg/mL)

Dose 

μmol/kg

Whole Blood Bile Liver

Wild-type PPARα-null Wild-type PPARα-null Wild-type PPARα-null

0 ND ND ND ND ND ND

12.5 20.6 ± 2.4 19.3 ± 2.2 56.8 ± 26.9 19.6 ± 2.2 181.2 ± 6.3 172.3 ±8.9

25 46.9 ± 3.2 36.4 ± 2.7* 784 ± 137.6 62.9 ± 16.7** 198.8 ± 15.4 218.3 ± 14.5

50 64.2 ± 6.5 71.2 ± 8.0 2174 ± 322.4 383 ± 109.9** 211.6 ± 13.3 239.7 ± 25.0

Source: Minata et al. 2010

Notes: Mean ± SD; ND= not detected (< 0.001 μg/mL); *p < 0.05; ** p < 0.01 

Tissue Transporters. As identified earlier, protein transporters from a number of families play a 

role in the tissue uptake of orally ingested PFOA. The transporters are located at the interface 

between the serum and the liver, kidneys, lungs, heart, brain, testes, ovaries, placenta, and uterus 

(Klaassen and Aleksunes 2010). The liver is an important uptake site for PFOA. OATPs and 

MRPs, at least one OAT, and the sodium-taurocholate cotransporting polypeptide (NTCP), a 

Perfluorooctanoic acid (PFOA) – May 2016 2-12
hepatic bile uptake transporter, have been identified at the interface of the liver with the portal 

blood and/or the canalicular membranes within the liver (Kim 2003; Kusuhara and Sugiyama 

2009; Zaïr et al. 2008). 

The impact of PFOA on several membrane transporter systems linked to biliary transport was 

studied by Maher et al. (2008) as part of a more detailed study of perfluorodecanoic acid 

(PFDA). A dose of 80 mg/kg by intraperitoneal (i.p.) injection (propylene glycol: water vehicle) 

was found to significantly increase (p<0.05) the expression of MRP3 and MRP4 in the livers of 

C57BL/6 mice 2 days after treatment as reflected in quantification of their deoxyribonucleic acid 

(DNA) transcripts. MRP3 and MRP4 are believed to protect the liver from accumulation of bile 

acids, bilirubin, and potentially toxic exogenous substances by promoting their excretion in bile. 

There were significant increases in serum bilirubin and bile acids after PFDA exposure, signifying 

increased export. Conversely, there were significant decreases (p<0.05) in the protein levels for 

OATP1a1, OATP1a4, and OATP1b2 as determined by Western Blot analysis and messenger 

ribonucleic acid (mRNA) measurements following exposure to 40 mg PFOA/kg (Cheng and 

Klaassen 2008). There was no significant impact on NTCP protein or the serum levels of bile 

acids. The OATPs are transporters responsible for the uptake of bile acids and other hydrophobic 

substances such as steroid conjugates, ecosinoids, and thyroid hormones into the liver.

These studies, all by the same laboratory, were carried out at high, single-dose exposures, 

which limit their value in extrapolating to low- and repeat-dose scenarios. The results suggest a 

decrease in the uptake of favored substrates into the liver and an increase in removal of favored 

substrates from the liver via bile. Upregulation of MRP3 and MRP4, coupled with decreased 

OATp levels, could be beneficial due to increased biliary excretion of bile acids, bilirubin, and 

conjugated metabolites of toxic chemicals, including PFOA. Based on the results with the more 

extensive evaluation of PFDA including mouse strains null for several receptors (PPARα, CAR, 

PXR, and FXR), the authors concluded that the changes in receptor proteins were primarily 

linked to activation of PPARα.

Impact of Developmental Age. Han (2003) administered groups of 4–8-week-old SpragueDawley rats (10 per gender per age) a single dose of 10 mg/kg/day PFOA by oral gavage. Blood 

samples were collected 24 hours after dosing and the plasma concentration of PFOA was 

measured by high-performance liquid chromatography mass spectrometry (HPLC/MS). In the 

4-week-old rats, the concentration of plasma PFOA was approximately 2.7 times higher in males 

than in the females (Table 2-8). In the 5- and 6-week-old female rats, the plasma PFOA 

concentrations were about twofold lower than in the 4-week-old rats. However, in the 5-weekold males, the concentration of plasma PFOA was about fivefold higher than in the 4-week-old 

group, suggesting a developmental change in excretion rate. Plasma concentrations did not differ 

appreciably among 5-, 6-, 7-, and 8-week-old rats within each gender but did differ between 

genders. In fact, PFOA plasma concentrations were 35–65-fold higher in males than in females 

at every age except at 4 weeks. Thus, it appears that maturation of the transport features

responsible for the gender difference in elimination occurs between the ages of 4 and 5 weeks in 

the rat.

Perfluorooctanoic acid (PFOA) – May 2016 2-13
Table 2-8. Plasma PFOA Concentrations (μg/ml) in Postweaning Sprague-Dawley Rats

Age (weeks) Males Females

4 7.32 ± 1.01a 2.68 ± 0.64

5 39.24 ± 3.89 1.13 ± 0.46

6 43.19 ± 3.79 1.18 ± 0.52

7 37.12 ± 4.07 0.57 ± 0.29

8 38.55 ± 5.44 0.81 ± 0.27

Source: Han 2003

Notes: 

a Mean ± SD; samples from 10 animals/gender/group 

Hinderliter (2004) and Hinderliter et al. (2006a) continued the investigation of the 

relationship between age and plasma PFOA in male and female Sprague-Dawley rats. Immature 

rats at 3, 4, and 5 weeks of age were administered PFOA via oral gavage at a single dose of 10 or 

30 mg/kg. Rats were not fasted prior to dosing. Two hours after dosing, five rats per gender per 

age group and dose group were sacrificed and blood samples were collected. The remaining five 

rats per gender per age and dose group were placed in metabolism cages for 24-hour urine 

collection. These rats were sacrificed at 24 hours and blood samples were collected.

In the male rats, plasma PFOA concentrations for either the 10- or 30-mg/kg dosage groups 

did not differ significantly by sample time (at 2 and 24 hours) or by animal age (3, 4, and 

5 weeks), except at 2 hours for the 5-week-old group (p<0.01), which showed the lowest PFOA 

level (Table 2-9). PFOA plasma concentrations following a 30-mg/kg dose were 2–3 times 

higher than those following a 10-mg/kg dose. These data do not demonstrate a difference 

between the 5-week-old rats and the younger 3- and 4-week-old groups at 24 hours after dosing, 

and thus do not support the observations from the Han study (2003). 

Table 2-9. Plasma PFOA Concentrations in Male Rats

Age

(weeks)

Dose 

(mg/kg)

Plasma PFOA (μg/ml)

2 Hours Post-Dose 24 Hours Post-Dose

Mean SD Mean SD

3 10 41.87 4.01 34.22 7.89

4 10 39.92 4.45 42.94 5.33

5 10 26.32* 6.89 40.60 3.69

3 30 120.65 12.78 74.16 18.23

4 30 117.40 18.10 100.81 13.18

5 30 65.66* 15.53 113.86 23.36

Source: Hinderliter 2004

Note: *Statistically significantly different by sample time and animal age (p<0.01).

In the female rats, plasma PFOA concentrations were significantly lower in the 5-week-old 

group than in the 3- or 4-week-old groups at the 24-hour time period for both doses and for the 

30-mg/kg dose group at 2 hours (Table 2-10). Plasma PFOA concentrations following a 

30-mg/kg dose were approximately one and one half to four times higher than those observed 

following a 10-mg/kg dose. 

At 24 hours post-dose, plasma PFOA levels in the female rats were significantly lower than the 

plasma PFOA levels in male rats, especially at 5 weeks of age. The data for the 5-week-old female 

rats compared to the 3- and 4-week-old groups at 24 hours are consistent with the Han (2003) data 

Perfluorooctanoic acid (PFOA) – May 2016 2-14
in that they demonstrate a decline in plasma levels compared to their earlier measurements. Thus, 

the developmental change is one that appears to be unique to the female rat.

Table 2-10. Plasma PFOA Concentrations in Female Rats

Age

(weeks)

Dose 

(mg/kg)

Plasma PFOA (μg/ml)

2 Hours Post-Dose 24 Hours Post-Dose

Mean SD Mean SD

3 10 37.87 5.77 13.55 b 3.83

4 10 29.88 12.15 18.98 b 7.01

5 10 33.23 7.41 1.36 a, b 0.87

3 30 84.86 10.51 51.43 b 13.61

4 30 80.67 14.10 28.01 b 9.90

5 30 56.90 a 29.66 3.42 a, b 1.95

Source: Hinderliter 2004

Notes: 

a Statistically significantly different from the 3- and 4-week values (p < 0.01). b Statistically significantly different from 2-hour values (p < 0.01). 

The data demonstrate that both dose and gender influence plasma levels. Post-dosing 

clearance (CL) is slow for both doses at 2 and 24 hours in males and females at postnatal weeks 

3 and 4. At 5 weeks, however, the plasma levels after 24 hours are greater than those at 2 hours 

in males. In females, for the high dose at 2 hours, plasma levels are similar to those in males, 

while at 24 hours they are only 3% of the value for males. This suggests that uptake from the 

intestines is similar while the rate of excretion at 5 weeks and beyond is considerably greater for 

female rats than males. They are comparable for postnatal weeks 3 and 4. 

In a supplemental study to determine the effect of fasting (Hinderliter [2004] and Hinderliter 

et al. [2006a]), 4-week-old rats, 4 rats per gender, were administered 10 mg/kg PFOA via oral 

gavage. Animals (two per gender) were fasted overnight for 12 hours before dosing with PFOA. 

All the rats were sacrificed at 24 hours post dosing and blood was collected for analysis of PFOA 

in plasma. Plasma PFOA concentrations in male rats were 64.95 and 30.00 μg/ml for the fasted 

and nonfasted animals, respectively. Plasma PFOA concentrations in the female rats were 68.16 

and 26.54 μg/ml for the fasted and nonfasted animals, respectively. Given the consistency in the 

4-week-old rat plasma PFOA concentrations, the authors concluded that age-dependent changes 

in female PFOA elimination are observable between 3 and 5 weeks of age. PFOA uptake was 

greater in the fasted animals than the fed animals, suggesting competition for uptake in the 

presence of food components that share common transporters and/or decreased contact of PFOA 

with the intestinal epithelium in the presence of dietary materials. 

Distribution during Pregnancy and Lactation 

Humans. T. Zhang et al. (2013) recruited pregnant females for a study to examine the 

distribution of PFOA between maternal blood, cord blood, the placenta, and amniotic fluid. 

Thirty-two females from Tianjin, China, volunteered to take part in the study. Samples were 

collected at time of delivery. Maternal ages ranged from 21 to 39 years, gestation periods ranged 

from 35 to 37 weeks. It was the first child for 26 of the females and a second child for 6. The 

study yielded 31 maternal whole blood samples, 30 cord blood samples, 29 amniotic fluid 

samples, and 29 placentas. The maternal blood contained variable levels of 10 PFASs, eight 

acids, and two sulfonates. The mean maternal blood concentration was highest for PFOS 

Perfluorooctanoic acid (PFOA) – May 2016 2-15
(14.6 ng/mL) followed by PFOA (3.35 ng/mL). In both cases, the mean was greater than the 

median, indicating a distribution skewed toward the higher concentrations. 

PFOA was found in all fluids/tissues sampled. PFOA was transferred to the amniotic fluid to 

a greater extent than PFOS, based on their relative proportions in the maternal blood and cord 

blood. Compared to the mean PFOA blood levels in the pregnant females, the mean levels in the 

cord blood, placenta, and amniotic fluid were 47%, 59%, and 1.3%, respectively, of those in the 

mother’s blood. The correlation coefficients between the maternal PFOA blood levels and 

placenta, cord blood, and amniotic fluid levels were good (0.7–0.9) and the relationships 

statistically significant (p<0.001).

Rat. An oral two-generation reproductive toxicity study of PFOA in rats was conducted 

(Butenhoff et al. 2004a). Five groups of rats (30 gender/group) were administered PFOA by 

gavage at doses of 0, 1, 3, 10, and 30 mg/kg/day. At scheduled sacrifice, after completion of the 

cohabitation period in F0 male rats and on lactation day (LD) 22 in F0 female rats, blood 

samples (3/gender/group-control; 10/gender/group-treated) were collected from animals dosed 

with 0, 10, and 30 mg/kg for analysis of PFOA. Serum analysis for the F0 generation males in 

the control, 10-, and 30-mg/kg/day groups sampled at the end of cohabitation showed that PFOA 

was present in all samples tested, including controls. Control males had an average concentration 

of 0.0344 ± 0.0148 μg/ml PFOA. Levels of PFOA were similar in the two male dose groups; 

treated males had 51.1 ± 9.30 and 45.3 ± 12.6 μg/ml, respectively, for the 10- and 30-mg/kg/day 

dose groups. In the F0 female controls, serum PFOA was below LOQ (0.00528 μg/ml). Levels of 

PFOA found in female sera increased between the two dose groups; treated females had an 

average concentration of 0.37 ± 0.0805 and 1.02 ± 0.425 μg/ml, respectively, for the 10- and 

30-mg/kg/day dose groups. 

PFOA levels during gestation and lactation were studied by Hinderliter et al. (2005) and 

Mylchreest (2003). Groups of 20 pregnant Sprague-Dawley rats were dosed with 0, 3, 10, and 

30 mg/kg/day of PFOA during days 4–10, 4–15, and 4–21 of gestation, or from gestation day 

(GD) 4 to LD 21. Maternal blood samples were collected at 2 hours ± 30 minutes (mins) postdose on a daily basis. Clinical observations and body weights were recorded daily. Five animals 

per dose group were sacrificed at specific time periods to harvest the conceptus and/or placenta 

and amniotic fluid. On GD 10, only embryos were recovered, and on GDs 15 and 21, the 

placentas, amniotic fluid, and embryos/fetuses were collected.

The remaining five rats per group were allowed to deliver their pups. On LDs 0, 3, 7, 14, and 

21, the pups were counted, weighed (genders separate), and examined for abnormal appearance 

and behavior. Randomly selected pups were sacrificed and blood samples were collected. On 

LDs 3, 7, 14, and 21, the dams were anesthetized and milk and blood samples were collected; 

dams were removed from their litters 1–2 hours prior to collection.

Plasma, milk, amniotic fluid extract, and tissue homogenate (placenta, embryo, and fetus) 

supernatants were analyzed for PFOA concentrations by HPLC/MS. Maternal PFOA plasma 

levels during gestation and lactation are presented in Table 2-11. Maternal plasma levels at 

2 hours post-dosing (approximately the time of peak blood levels following a gavage dose) were 

fairly similar during the course of the study with a mean level of 11.2, 26.8, and 66.6 μg/ml in 

the 3-, 10-, and 30-mg/kg/day groups, respectively; PFOA levels in the control group were below 

the LOQ (0.05 μg/ml). 

Perfluorooctanoic acid (PFOA) – May 2016 2-16
Table 2-11. Maternal Plasma PFOA Levels (μg/ml) in Rats During Gestation and Lactation

Exposure Period Sample Time

Dose

3 mg/kg/day 10 mg/kg/day 30 mg/kg/day

GD 4 - GD 10 GD 10 plasma 8.53 ± 1.06 23.32 ± 2.15 70.49 ± 8.94

GD 4 - GD 15 GD 15 plasma 15.92 ± 12.96 29.40 ± 14.19 79.55 ± 3.11

GD 4 - GD 21 GD 21 plasma 14.04 ± 2.27 34.20 ± 6.68 76.36 ± 14.76

GD 4 - LD 3 LD 3 plasma 11.01 ± 2.11 22.47 ± 2.74 54.39 ± 17.86

GD 4 - LD 7 LD 7 plasma 10.09 ± 2.90 25.83 ± 2.07 66.91 ± 11.82

GD 4 - LD 14 LD 14 plasma 9.69 ± 0.92 23.79 ± 2,81 54.65 ± 11.63

GD 4 - LD 21 LD 21 plasma 9.04 ± 1.01 28.84 ± 5.15 64.13 ± 1.45

NA Average plasma 11.19 ± 2.76 26.84 ± 4.21 66.64 ± 9.80

Source: Hinderliter et al. 2005; Mylchreest 2003

Notes: Mean ± SD; samples were from five dams/group/time point and were collected 2 hours post-dosing. 

PFOA levels in the placenta, amniotic fluid, and embryo/fetus are presented in Table 2-12. 

The levels of PFOA in the placenta on GD 21 were approximately twice the levels observed on 

GD 15, and the levels of PFOA in the amniotic fluid were approximately four times higher on 

GD 21 than on GD 15. The concentration of PFOA in the embryo/fetus was highest in the GD 10 

embryo and lowest in the GD 15 embryo; PFOA levels in the GD 21 fetus were intermediate.

Table 2-12. Placenta, Amniotic Fluid, and Embryo/Fetus PFOA Concentrations

in Rats (μg/ml)

Exposure Period Tissue

Dose

3 mg/kg/day 10 mg/kg/day 30 mg/kg/day

GD 4–GD 10 GD 10—embryo 1.40 ± 0.30 3.33 ± 0.81 12.49 ± 3.50

GD 4–GD 15 GD 15—placenta

—amniotic fluid

—embryo

2.22 ± 1.79

0.60 ± 0.69

0.24 ± 0.19

5.10 ± 1.70

0.70 ± 0.15

0.53 ± 0.18

13.22 ± 1.03

1.70 ± 0.91

1.24 ± 0.22

GD 4–GD 21 GD 21—placenta

—amniotic fluid

—fetus

3.55 ± 0.57

1.50 ± 0.32

1.27 ± 0.26

9.37 ± 1.76

3.76 ± 0.81

2.61 ± 0.37

24.37 ± 4.13

8.13 ± 0.86

8.77 ± 2.36

Source: Hinderliter et al. 2005; Mylchreest 2003

Note: Mean ± SD; samples were pooled by litter and were collected 2 hours post-dosing. 

The concentrations of PFOA in the plasma of the GD 21 fetus were approximately half the 

levels observed in the maternal plasma (Table 2-11). The values were about twice as high in the 

dams as in the pups with mean values of 14.04, 34.20, and 76.36 μg/ml, respectively, in the 3-, 

10-, and 30-mg/kg/day groups for the dams and 5.88, 14.48, and 33.11 μg/ml, respectively, for 

the pups. Pup plasma levels decreased between birth and LD 7 (Table 2-13) and were, thereafter, 

similar to the levels observed in the milk (Table 2-14). The pups were not separated by gender. 

The concentration of PFOA in the milk also was fairly similar throughout lactation and was 

approximately one-tenth of the PFOA levels in the maternal plasma (see Table 2-11); the mean 

values for maternal milk were 1.1, 2.8, and 6.2 μg/ml in the 3-, 10-, and 30-mg/kg/day groups, 

respectively (Table 2-14). 

Perfluorooctanoic acid (PFOA) – May 2016 2-17
Table 2-13. Fetus/Pup PFOA Concentration (μg/ml) in Rats During

Gestation and Lactation

Exposure Period Tissue

Dose

3 mg/kg/day 10 mg/kg/day 30 mg/kg/day

GD 4–GD 21 GD21––fetal plasma 5.88 ± 0.69 14.48 ± 1.51 33.11 ± 4.64

GD 4–LD 3 LD 3––pup plasma 2.89 ± 0.70 5.94 ± 1.44 11.96 ± 1.66

GD 4–LD 7 LD 7––pup plasma 0.65 ± 0.20 2.77 ± 0.58 4.92 ± 1.28

GD 4–LD 14 LD 14––pup plasma 0.77 ± 0.10 2.22 ± 0.38 4.91 ± 1.12

GD 4–LD 21 LD 21––pup plasma 1.28 ± 0.72 3.25 ± 0.52 7.36 ± 2.17

Source: Hinderliter et al. 2005; Mylchreest 2003

Note: Mean ± SD; samples were pooled by litter and were collected 2 hours post-dosing. 

Table 2-14. PFOA Levels (μg/ml) in Rats Maternal Milk During Lactation

Exposure Period Sample Time

Dose

3 mg/kg/day 10 mg/kg/day 30 mg/kg/day

GD 4–LD 3 LD 3–milk 1.07 ± 0.26 2.03 ± 0.33 4.97 ± 1.20

GD 4–LD 7 LD 7–milk 0.94 ± 0.22 2.74 ±0.91 5.76 ±1.26

GD 4–LD 14 LD 14–milk 1.15 ± 0.06 3.45 ± 1.18 6.45 ± 1.38

GD 4–LD 21 LD 21–milk 1.13 ± 0.08 3.07 ± 0.51 7.48 ± 1.63

NA Average milk 1.07 ± 0.09 2.82 ± 0.60 6.16 ± 1.06

Source: Hinderliter et al. 2005; Mylchreest 2003

Notes: Mean ± SD; samples were from 5 dams/group/time point and were collected 2 hours post-dosing. 

Mouse. Fenton et al. (2009) orally dosed pregnant CD-1 mice (n = 25/group) with 0, 0.1, 1, and 

5 mg PFOA/kg on GD 17. On GD 18, five dams/group were sacrificed and trunk blood, urine, 

amniotic fluid, and the fourth and fifth mammary glands were collected. One fetus/dam was 

euthanized and retained for whole-pup analysis. The remaining dams were allowed to litter. 

Biological samples as described above excluding amniotic fluid also were collected on postnatal 

days (PNDs) 1, 4, 8, and 18. As before, at each time-point, a single pup was euthanized and 

retained for whole-pup analysis. Blood from the remaining pups was collected and pooled. Milk 

was collected from dams on PNDs 2, 8, 11, and 18 following a 2-hour separation of the pups 

from the dam. 

The concentration of PFOA in dam serum was approximately twice that detected in amniotic 

fluid (Table 2-15). Compared to the amniotic fluid, the concentration of PFOA in the fetuses was 

increased by 2.3-, 3.1-, and 2.7-fold at 0.1, 1, and 5 mg/kg, respectively. The highest 

concentration of PFOA was detected in the serum of nursing dams. In the dams, the 

concentration of PFOA in the serum exhibited a U-shaped response curve; the lowest serum 

concentration was observed at the time of peak lactation. Dam mammary tissue and milk PFOA 

concentrations showed a U-shaped response that mirrored that found in the dam’s serum. The 

concentration of PFOA in pup’s serum was significantly higher than PFOA concentration in 

dam’s serum and appeared to decrease as the time for weaning approached. When pup PFOA 

concentration was calculated with consideration for pup body weight gain, PFOA body burden 

increased through the peak of lactation and began to decrease by PND 18, showing an inverse 

U-shaped response curve.

Perfluorooctanoic acid (PFOA) – May 2016 2-18
Table 2-15. PFOA Levels (ng/ml) in Mice During Gestation and Lactation in Selected 

Fluids and Tissues

Tissue Day

Dose

0.1 mg/kg 1 mg/kg 5 mg/kg

Dam Serum GD 18 143 ± 19 1697 ± 203 7897± 663

PND 1 217.5 ± 35 1957.0 ± 84 9845.6 ± 1478

PND 4 110.0 ± 12 1269.4 ± 235 6776.6 ± 561

PND 8 46.7 ± 21 360.8 ± 98 1961.8 ± 414

PND 18 123.3 ± 41 1035.2 ± 305 5156.5 ± 1201

Amniotic Fluid GD 18 99.0 ± 28 865.3 ± 191 3203.8 ± 492

Dam Urine GD 18 21.9 ± 8.6 104.9 ± 69.7 666.7 ± 169

PND 1 7.7 ± 1.7 116.8 ± 64 492.3 ± 119

PND 4 8.4 ± 6.4 53.5 ± 15 401.5 ± 117

PND 8 0.8 ± 0.22 11.6 ± 6.2 40.1 ± 17

PND 18 1.8 ± 1.1 18.7 ± 8.6 91.7 ± 49

Mammary Gland GD 18 18.9 ± 1.9 307.2 ± 30.4 1429± 186

PND 1 27.4 ± 6.8 343.8 ± 53 1933.5 ± 194

PND 4 9.6± 8.4 239.2 ± 53 1461.8 ± 267

PND 8 2.4 ± 3.8 71.7 ± 22 411.8 ± 78

PND 18 17.1 ± 10 239.9 ± 76 1372.8 ± 240

Milk PND 2 32.5 ± 12 716.7 ± 145 1236.6 ± 1370

PND 8 11.6 ± 8.1 77.4 ± 19 245.1 ± 26

PND 11 5.4 ± 1.0 42.3 ± 9.1 282.5 ± 162

PND 18 43.5 ± 19 251.8 ± 147 909.8 ± 308

Whole Pup GD 18 136.3 ± 15 1665.8 ± 213 6256.5 ± 751

PND 1 150.9 ± 21 1606.9 ± 288 7134.5 ± 1097

PND 4 91.8 ± 8.9 1183.2 ±187 5071.4 ± 267

PND 8 60.9 ± 16 729.0 ± 92 3118.5 ± 424

PND 18 17.5 ± 11 251.9 ± 112 1391.5 ± 118

Pup Serum PND 1 324.7 ± 36 3926.8 ± 480 16,286.4 ± 1372

PND 4 267.6 ± 47 3020.8 ± 223 11,925.2 ± 1077

PND 8 260.2 ± 56 2548.2 ± 245 9215.8 ± 594

PND 18 111.8 ± 30 1124.8 ± 236 5894.3 ± 743

Source: Fenton et al. 2009

Pregnant C57BL/6/Bkl mice were fed diets containing 0.3 mg PFOA/kg/day from GD 1 

through the end of pregnancy. At birth, the PFOA concentrations in the offspring were 

0.7 ± 0.1 μg/g in the brain and 16.3 ± 4.1 μg/g in the liver (Onishchenko et al. 2011). 

Perfluorooctanoic acid (PFOA) – May 2016 2-19
Macon et al. (2011) gavage-dosed CD-1 mice with 0, 0.3, 1.0, or 3.0 mg PFOA/kg from GD 

1 to GD 17 or with 0, 0.01, 0.1, and 1.0 mg PFOA/kg from GD 10 to GD 17. In the full gestation 

experiment (GD 1–17), offspring were sacrificed on PNDs 7, 14, 21, 28, 42, 63, and 84, and in 

the half gestation experiment (GDs 10–17), female offspring were sacrificed on PNDs 1, 4, 7, 14, 

and 21. Serum, liver, and brain from the offspring were analyzed for PFOA by HPLC/MS/MS.

At the lowest dose, PFOA concentration in the serum peaked at or before PND 7, but the two 

higher doses peaked around PND 14 (Table 2-16). Calculated blood burdens which take into 

account the increasing blood volumes and body weights for females showed an inverted 

U-shaped curve peaking at PND 14 for all doses. In the liver, PFOA concentration decreased 

over time with the highest concentration observed at PND 7. Lower concentrations of PFOA 

were detected in the brain of the offspring on PND 7 and 14. 

Table 2-16. Female Offspring PFOA Levels (ng/ml) in Mice After GD 1-17 Exposure 

Tissue Day

Dose

0.3 mg/kg 1.0 mg/kg 3.0 mg/kg

Serum PND 7 4980 ± 218 11026 ± 915 20700 ± 3900

PND 14 4535 ± 920 16950 ± 3606 26525 ± 2446

PND 21 1194 ± 394 377 ± 607 8343 ± 1078

PND 28 630 ± 162 1247 ± 208 4883 ± 1378

PND 42 377± 81 663 ± 185 2058 ± 348

PND 63 55 ± 17 176 ± 85 ‒

PND 84 16 ± 5 71 ± 8 125

Liver PND 7 2078 ± 90 8134 ± 740 16700 ± 749

PND 14 972 ± 124 4152 ± 483 10290 ± 1028

PND 21 1188 ± 182 1939 ± 637 2339 ± 1241

PND 28 678 ± 130 2007 ± 560 7124 ± 1081

PND 42 342 ± 87 617 ± 145 1145 ± 274

PND 63 118 ± 22 320 ± 113 417 ± 160

PND 84 43 ± 12 55± 12 235 ± 79

Brain PND 7 150 ± 26 479 ± 41 1594 ± 162

PND 14 65 ± 12 241 ± 20 650 ± 44

PND 21 <LOQ 31 ± 5 133 ± 23

PND 28 <LOQ <LOQ 62 ± 93

PND 42 <LOQ <LOQ <LOQ

PND 63 <LOQ <LOQ <LOQ

PND 84 <LOQ <LOQ <LOQ

Source: Macon et al. 2011

Notes: 

‒ = not measured

LOQ: serum full gestation = 10-20 ng/g; liver = 35 ng/g; brain = 35 ng/g; late gestation serum = 5 ng/mL

After an exposure to low doses of PFOA from GD 10 to GD 17, serum PFOA concentration 

in the female offspring declined from PND 1 through the end of the experiment (Table 2-17). 

Calculated blood burden showed a gradual increase from PND 1 to PND 14, followed by a 

decline through PND 21. 

Perfluorooctanoic acid (PFOA) – May 2016 2-20
Table 2-17. Female Offspring Serum PFOA Levels (ng/ml) in Mice

After GD 10-17 Exposure 

Tissue Day

Dose

0.01 mg/kg 0.1 mg/kg 1.0 mg/kg

Serum PND 1 284.5 ± 21.0 2303.5 ± 114.4 16305.5 ± 873.5

PND 4 184.1 ± 12.1 ‒ ‒

PND 7 150.7 ± 20.9 1277.8 ± 122.6 11880.3 ± 1447.6

PND 14 80.2 ± 13.9 645.4 ± 114.2 6083.7 ± 662.6

PND 21 16.5 ± 2.1 131.7 ± 24.5 2025.1 ± 281.9

Blood Burden 

(calculated)

PND 1 15.2 ± 1.7 114.3 ± 5.4 926.0 ± 47.6

PND 4 20.6 ± 0.1 ‒ ‒

PND 7 27.3 ± 3.8 221.7 ± 24.9 1965.9 ± 256.7

PND 14 27.0 ± 4.6 218.5 ± 39.8 2033.6 ± 293.5

PND 21 7.9 ± 1.0 66.4 ± 12.8 984.7 ± 142.8

Source: Macon et al. 2011

Note: ‒ = not measured, blood burden determined by (body weight x (58.5/1000) x serum x 0.55)

White et al. (2011) measured serum PFOA concentrations in three generation of CD-1 mice

(Table 2-18). Pregnant mice (F0, n = 10–12 dams/group) were gavage-dosed with 0, 1, and 5 mg 

PFOA/kg from GD 1–17. A separate group of pregnant mice (n = 7–10 dams/group) were 

gavage-dosed with either 0 or 1 mg PFOA/kg from GD 1–17 and received drinking water 

containing 5 parts per billion (ppb) PFOA beginning on GD 7 and continuing until the end of the 

study for their offspring, except during breeding and early gestation, to simulate a chronic lowdose exposure. An increase in serum PFOA concentration was observed in the control + 5 ppb 

PFOA groups in the F1 and F2 generations and in the 1-mg/kg + 5-ppb PFOA group of the F2 

generation. A decrease was observed for the remaining groups. 

Table 2-18. Serum PFOA Levels (ng/ml) in Mice Over Three Generations

Generation/

Day

Dose

0 mg/kg + 5 ppb 1 mg/kg 1 mg/kg + 5 ppb 5 mg/kg

Dams at weaning F0/ PND 22 74.8 ± 11.3 6658.0 ± 650.5 4772.0 ± 282.4 26980.0 ± 1288.2

F1/~PND 91 86.9 ± 14.5 9.3 ± 2.6 173.3 ± 36.4 18.7 ± 5.2

Offspring F1/PND 22 21.3 ± 2.1 2443.8 ± 256.4 2743.8 ± 129.7 10045 ± 1125.6

F1/PND 42 48.9 ± 4.7 609.5 ± 72.2 558.0 ± 55.8 1581.0 ± 245.1

F1/PND 63 66.2 ± 4.1 210.7 ± 21.9 187.0 ± 24.1 760.3 ± 188.3

F2/PND 22 26.6 ± 2.4 4.6 ± 1.2 28.5 ± 3.7 7.8 ± 1.9

F2/PND 42 57.4 ± 2.9 0.4 ± 0.0 72.8 ± 5.8 0.4 ± 0.0

F2/PND 63 68.5 ± 9.4 1.1 ± 0.5 69.2 ± 4.3 1.2 ± 0.5

Source: White et al. 2011

Subcellular Distribution. Han et al. (2005) examined the subcellular distribution of PFOA in 

the liver and kidney of male and female rats. Male and female Sprague-Dawley Crl:CD (SD)IGS 

BR rats were gavage-dosed with 25 mg/kg [14C] PFOA and sacrificed 2 hours after dosing. 

Blood was collected and the liver and kidneys were removed. Five subcellular fractions (nuclei 

Perfluorooctanoic acid (PFOA) – May 2016 2-21
and cell debris, lysosome and mitochondria, microsome, light microsome and ribosome, and 

membrane-free cytosol) were obtained by differential centrifugation. The radioactivity per gram

(g) of each fraction and the total radioactivity were measured. 

In the male liver, the highest proportion of total reactive residues (TRR) of PFOA was 

located in the nuclei and cell debris (40%). The TRR for the other subcellular fractions were as 

follows: membrane-free cytosol 26 percent% TRR, lysosome and mitochondria ~14% TRR, and 

microsome ~16% TRR. The level of PFOA in the light microsome and ribosome was ~1% TRR. 

In the female liver, the highest proportion of PFOA was found in the membrane-free cytosol, 

48% TRR. The TRR were nuclei and cell debris ~31% TRR, lysosome and mitochondria 

~12% TRR, and microsome ~8% TRR. As observed in the males, the level of PFOA in the light 

microsome and ribosome was ~1% TRR (Han et al. 2005). 

In the male kidney, the level of PFOA was 79% TRR in the membrane-free cytosol, 

15% TRR in the nuclei and cell debris, and 4% TRR in the lysosome and 

mitochondria/microsome/ light microsome and ribosome (combined). In the female kidney, the 

level of PFOA was 71% TRR in the cytosol, 21% TRR in the nuclei and cell debris, and 

8% TRR in the lysosome and mitochondria/ microsome/light microsome and ribosome 

(combined). Further examination showed that in both genders, 98% of PFOA in the plasma was 

protein bound. The protein-bound fraction of PFOA in the liver cytosol was 56% TRR. In the 

kidney, the protein-bound fraction of PFOA in males was 42% TRR and 17% TRR in females 

(Han et al. 2005). 

Based on the results, the authors concluded that subcellular distribution of PFOA in the rat 

liver was gender-dependent because the proportion of PFOA in the liver cytosol of female rats 

was almost twice that of the male rats. They hypothesized that the female might have a greater 

amount than the male of an unknown liver cytosolic binding protein with an affinity for 

perfluorinated acids. They also hypothesized that the unknown protein or protein complex might 

normally aid in transport of fatty acids from the liver. In the kidney, the subcellular distribution

did not show the gender difference seen with the liver; however, the protein-bound fraction for 

the males (42%) was about twice that for the females (17%) (Han et al. 2005). 

Inhalation Exposure 

In a repeated exposure study, Hinderliter (2003) and Hinderliter et al. (2006b) exposed 6–8-

week-old male and female rats (5 per gender per group) to 0-, 1-, 10-, and 25-mg/m3

 aerosol 

concentrations of PFOA for 6 hours/day, 5 days/week for 3 weeks. Blood was collected 

immediately before and after the daily exposure period 3 days/week. The aerosols had MMADs 

of 1.3–1.9 μm with GSDs of 1.5–2.1. PFOA plasma concentrations were proportional to the 

inhalation exposure concentrations, and repeated exposures produced little plasma carryover in 

females, but significant day-to-day carryover in males. Male rats reached steady-state plasma 

levels of 8, 21, and 36 μg/ml for the 1-, 10-, and 25-mg/m3

 groups, respectively, by 3 weeks. In 

females, the post-exposure plasma levels were 1, 2, and 4 μg/ml for the 1-, 10-, and 25-mg/m3

groups, respectively. When measured immediately before the next daily exposure, plasma levels 

had returned to baseline in females, demonstrating CL within 24 hours of each daily dose. 

Dermal Exposure

No data were identified on tissue distribution following dermal exposures. 

Perfluorooctanoic acid (PFOA) – May 2016 2-22
2.3 Metabolism

Several studies have examined metabolism of PFOA. However, no studies show clear 

evidence of metabolism. Ophaug and Singer (1980) found no change in fluoride ion level in the 

serum or urine following oral administration of PFOA to female Holtzman rats. Ylinen et al. 

(1989) found no evidence of phase II metabolism of PFOA following a single intraperitoneal 

PFOA dose (50 mg/kg) in male and female Wistar rats. The free anionic and possible conjugated 

forms of PFOA in the urine were separated using BondElut tubes. The tubes contain NH2, which 

is a weaker anion exchange sorbent and a good choice for retaining strong anions. The samples 

were aspirated through the tube, washed with water, and eluted with sodium 

bicarbonate/carbonate-buffer. The aspirate and eluate from the separation method were analyzed 

by gas chromatography. PFOA was not detected in the aspirate, but was retained with the 

cationic amino phase found in the eluate. This also occurred in control blanks spiked with PFOA. 

The authors concluded that because the PFOA anion was completely bound to the weak cationic 

amino phase in both the spiked controls and urine samples, PFOA in urine is not altered by phase 

II metabolism (Ylinen et al. 1989). 

2.4 Excretion

Excretion data are available for oral exposure in humans and laboratory animals. Several 

studies have investigated the elimination of PFOA in humans, cynomolgus monkeys, and rats. In 

human females, elimination pathways include pregnancy (cord blood) and lactation (breast milk) 

(Apelberg et al. 2007; Tao et al. 2008; Thomsen et al. 2010; Völkel et al. 2008; von Ehrenstein et 

al. 2009). 

Elimination half-lives differ among species. There are also significant gender differences in 

humans and some laboratory animal species. Information from humans does not, at this time, 

provide sufficient data to determine the magnitude of interindividual and gender differences in 

excretory half-lives. The transporters appear to play an important role in renal excretion of 

PFOA and possibly its biliary elimination as well.

Humans. The urinary excretion of PFOA in humans is impacted by the isomeric composition of 

the mixture present in blood and the gender/age of the individuals. The half-lives of the 

branched-chain PFOA isomers are shorter than those for the linear molecule, an indication that 

renal resorption is less likely with the branched chains. 

Y. Zhang et al. (2013) determined half-lives for PFOA isomers based on paired serum 

samples and early morning urine samples collected from healthy volunteers in two large Chinese 

cities. Half-lives were determined using a one compartment model and an assumption of first 

order CL. The Vd applied in the analysis as determined by Thompson et al. (2010) was 

170 mL/kg. CL was estimated from the concentration in urine normalized for creatinine and 

assuming excretion of 1.2 and 1.4 L/day of urine and 0.9 and 1.1 mg creatinine/day for males 

and females, respectively. The mean half-life for the sum of all PFOA isomers in younger 

females (n = 12) was 2.1 years (range 0.19–5.2 years) while that for all males and older females 

(n = 31) was 2.6 (range 0.0059–14 years); the medians were 1.8 and 1.7 years, respectively. The 

mean values for the four branched-chain isomers of PFOA were lower than the value for the 

linear chain, suggesting that resorption transporters might favor uptake of the linear chain over 

the branched-chain isomers. Older females and males have longer half-lives than young females, 

suggesting the importance of monthly menstruation as a pathway for excretion (Y. Zhang et al. 

2013). 

Perfluorooctanoic acid (PFOA) – May 2016 2-23
T. Zhang et al. (2014) derived estimates for PFOA’s urinary excretion rate using paired urine 

and blood samples from 54 adults (29 males and 25 females) in the general population and 27 

pregnant females in Tainjin, China. The age range for the general population was 22–62 years 

and for the pregnant females was 21–39 years. Urinary excretion was calculated based on the 

concentration in the urine times volume of urine wherein a urinary volume of 1,200 mL/day was 

applied to all females and 1,600 mL/day for all males. Urine samples were first-draw morning 

samples. Total daily intakes for PFOA were calculated from the concentration in blood using 

first order assumptions, a half-life of 2.3 years (Bartell et al. 2010) and a Vd of 170 mL/kg 

(Lorber and Egeghy 2011; Thompson et al. 2010). PFOA was detected in the blood samples for 

all participants but for only 76% of the urine samples from the general population and 30% for 

the pregnant females. There was a direct correlation between the PFOA concentrations in blood 

and creatinine adjusted urine (r = 0.348 p = 0.013) for the general population but not for the 

pregnant females. When limited to the eight females who had detectable levels in both blood and 

urine, there was a significant correlation (r= 0.724, p = 0.042). 

Among the general population, the daily urinary excretion rate accounted for 25% of the 

estimated intake with the excretion higher in males (31%) than in females (19%). The urine: 

blood ratio was lower for pregnant females than for nonpregnant females (0.0011 versus 0.0029), 

suggesting other removal pathways such as placenta and cord blood. There was little difference 

between the younger menstruating females (21–50 years versus 51–61 years), but there is no 

indication that data were collected from the participants relative to menstruation status on the day 

of blood and urine collection. 

Wong et al. (2014) looked at the role of menstrual blood as an excretory pathway to explain 

the shorter half-life of PFOS in females than in males. They fit a population-based PK model to 

six cross-sectional NHANES data sets (1999–2012) for males and females. They concluded that 

menstruation could account for about 30% of the PFOS elimination half-life difference between 

females and males. Although Wong et al. (2014) studied PFOS and not PFOA, their findings are 

relevant to both chemicals.

Elimination of PFOA by way of the gastrointestinal tract was reported in a case history of a 

single human male with high serum levels of perfluorinated chemicals that appeared to originate 

from household dust following the installation of new carpeting (Genuis et al. 2010). Treatment 

with cholestyramine, a bile acid sequestrant for 20 weeks (4g/day, three times a day), lowered his 

serum PFOA concentration from 5.9 ng/g serum to 4.1 ng/g serum. More dramatic decreases 

were observed with serum PFOS (23–14.4 ng/g serum) and PFHxS (58–46.8 ng/g serum), which 

were present at higher levels in the serum. This observation suggests that excretion with bile and 

possible enterohepatic resorption via intestinal transporters limits the loss of absorbed PFOA via 

feces in the absence of a binding agent such as cholestyramine.

2.5 Animal Studies

Oral Exposure 

Monkey. Butenhoff et al. (2004b) studied the fate of PFOA in cynomolgus monkeys in a 

6-month oral exposure study. Groups of four to six male monkeys each were administered PFOA 

daily via oral capsule at DRs of 0, 3, 10, and 30/20 mg/kg for 6 months. Two monkeys exposed 

to 10 mg/kg and three monkeys exposed to 20 mg/kg were monitored for 21 weeks (recovery 

period) following dosing. Urine and fecal samples were collected at 2-week intervals and were 

analyzed for PFOA concentrations. 

Perfluorooctanoic acid (PFOA) – May 2016 2-24
Urine PFOA concentrations over the duration of the study were 53 ± 25, 166 ± 83, and 

181 ± 100 μg/ml in the 3-, 10-, and 30-/20-mg/kg dose groups, respectively, and reached steadystate after 4 weeks. Within two weeks of recovery, urine PFOA concentrations were <1% of the 

value measured during treatment and decreased slowly thereafter. Fecal PFOA concentrations 

were 6.8 ± 5.3, 28 ± 20, and 50 ± 33 μg/g in the 3-, 10-, and 20-mg/kg dose groups, respectively. 

Within two weeks of recovery, fecal PFOA concentrations dropped to less than 10% of the last 

value during treatment, and then declined slowly. These results are consistent with both renal and 

biliary excretion in male monkeys. 

Rat. There have been a number of studies of excretion in rats because of the gender differences 

noted in serum levels. Hinderliter (2004) and Hinderliter et al. (2006a) investigated the 

relationship between age and urine PFOA concentrations in male and female Sprague-Dawley

rats. Immature rats 3, 4, or 5 weeks of age were administered PFOA via oral gavage as a single 

dose of 10 or 30 mg/kg. Two hours after dosing, five rats per gender per age group and dose 

group were sacrificed and blood samples were collected (see section 2.2.1). The remaining five 

rats per gender per age and dose group were placed in metabolism cages for 24-hour urine 

collection. Urinary output (volume) was not quantified or standardized for creatinine levels. 

Urine PFOA concentrations differed significantly with age, dose, and gender (p<0.01, Table 

2-19). Urinary excretion of PFOA was substantially higher in females than in males, and the 

female urine PFOA concentrations increased with age. In male rats, 24-hour urine PFOA 

concentrations decreased with age up to five weeks. In both genders, urine PFOA was higher 

(2.5 to 6.5 times) at the 30-mg/kg dose as compared to the 10-mg/kg dose. 

There was a difference in urinary excretion between the 3-week-old and 4/5-week-old male 

rats, with the older rats excreting ~50% less PFOA in the urine than the younger rats at 10 mg/kg 

and 30 mg/kg. If the data from urine are integrated with the plasma data in the same study 

(Table 2-9), the male plasma levels increased from the 3-week value and were relatively stable 

for weeks 4 and 5. In the females, urine excretion increased gradually with age (Table 2-19) and 

plasma concentrations decreased (Table 2-10).

Table 2-19. Urine PFOA Concentrations in Male and Female Rats

Age

(weeks)

Dose 

(mg/kg)

Urine PFOA (μg/ml at 24 hours post-dose)

Male Female

Mean SD Mean SD

3 10 9.57 4.86 21.17 8.95

4 10 4.53 2.45 23.26 15.27

5 10 4.03 2.36 49.77 24.64

3 30 51.76 28.86 94.89 26.26

4 30 28.70 18.84 104.12 28.97

5 30 15.65 6.24 123.16 51.56

Source: Hinderliter 2004

Hundley et al. (2006) examined excretion of PFOA in one male and one female CD rat 

(sexually mature). Each was given a single dose of 10 mg/kg 14C-PFOA and housed in a 

metabolism cage. Urine and feces were collected at 12, 24, 48, 72, 96, and 120 hours post-dose. 

The female rat excreted more PFOA over the 120-hour collection period than the male rat. In the 

male rat, 25.6% and 9.2% 14C-PFOA were excreted in the urine and feces, respectively. In the 

female rat, 73.9% and 27.8% 14C-PFOA were excreted in the urine and feces, respectively. The 

Perfluorooctanoic acid (PFOA) – May 2016 2-25
female rat excreted almost all of the PFOA by 48 hours compared with only 19% of the dose 

excreted by the male rat over the same amount of time. The cumulative percent of the dose 

excreted is shown in Table 2-20. 

Table 2-20. Cumulative Percent 14C-PFOA Excreted in Urine and Feces by Rats

Rat

Hours After Dosing

12 24 48 72 96 120

Male 0.6 8.7 19.2 23.4 30.2 34.3

Female 52.5 96.4 99.8 100.0 100.0 100.0

Source: Hundley et al. 2006

Adult male Sprague-Dawley rats (n = 7) were given a single gavage dose of 0.5 mg PFOA/kg 

and monitored for 38 days (Benskin et al 2009). Over the course of the study, the rats were held 

in metabolic cages and urine and feces were collected. The mean blood PFOA concentration was 

1.1 μg/mL 24 hours post-dose. During the first 24 hours post-dose, 65% of PFOA was excreted 

in the urine; most of the PFOA that was not absorbed was excreted in the feces. After that time 

period, 91–95% of the daily excreted PFOA was eliminated in the urine. On day 3, the mean 

PFOA concentration in urine and feces were 265 ng/g and 28 ng/g. The half-life for elimination 

from plasma in male rats was 13.4 days.

Cui et al. (2010) exposed 2-month-old male Sprague-Dawley rats (10 per group) to PFOA 

(96% active ingredient) at 0, 5, and 20 mg/kg/day once daily by gavage for 28 days. Urine and 

fecal samples were collected through use of metabolism cages at 24-hour intervals immediately 

following dosing on days 1, 2, 5, 7, 10, 14, 18, 21, 24, and 28 of the study. Daily urine volume 

and fecal weight were comparable across all groups throughout the study. As measured by 

excretion 24-hours after the first dose, 17.9% of the applied dose was excreted in the urine of the 

low-dose group and 22% for the high-dose group. The percent of the absorbed dose was 92.8% 

and 92.3% for the low and high doses, respectively, when the fecal excretion over the 24 hours 

following dosing was estimated to be unabsorbed material. During week 1, a sharp increase in 

urinary and fecal excretion expressed as percent of dose/day was observed in rats of both groups. 

The excretion rate leveled off at about 50% for the low-dose animals for the remainder of the 

28 days. In the case of the high-dose animals, the urinary excretion remained level at about 80% 

for the second and third weeks and then increased sharply to about 140% at 28 days. The fecal 

excretion rates were 7.2% and 7.7% for rats in the 5- and 20-mg/kg groups, respectively, during 

the first 24 hours post-dosing and continued an upward trend throughout the 28 days with the 

terminal percent/day about 25% for the low-dose group and 40% for the high-dose group. 

Dose is an important variable that impacts excretion. Rigden et al. (2015) exposed groups of 

five male Sprague-Dawley rats to doses of 0, 10, 33, and 100 mg/kg/day for 3 days and 

maintained them for 3 additional days; overnight urine was collected and body weight was 

measured daily. Of greatest interest relative to the limitations on renal resorption, is the doserelated increase in urine PFOA concentration and urine PFOA concentration per mg creatinine 

for the 33- and 100-mg/kg/day groups compared to the 10-mg/kg/day group. The peak in PFOA 

excretion normalized to creatinine occurred on day 3 after the cessation of dosing. The 

concentration at 33 mg/kg/day was 500 times greater than that at 10 mg/kg/day. At the 

100-mg/kg/day dose, the peak concentration was about 3,200 times greater than for the low dose. 

The low-dose excretion was only slightly greater than the controls. The urine results support the 

renal resorption hypothesis concept and suggest that there is a threshold limit on resorption that, 

Perfluorooctanoic acid (PFOA) – May 2016 2-26
once exceeded, dramatically increases PFOA loss in urine. As a consequence, half-life for 

continuous low-dose exposures will be longer than for single or short-term high-dose exposures. 

Other Species. Hundley et al. (2006) examined excretion of PFOA in CD mice, BIO-15.16 

hamsters, and New Zealand White rabbits. One male and one female of each species was given a 

single dose of 10-mg/kg 14C-PFOA and housed in metabolism cages. Urine and feces were 

collected at 12, 24, 48, 72, 96, and 120 hours post-dose. Additional samples were collected from 

rabbits at 144 and 168 hours post-dose. 

Over 120 hours, the male mouse excreted 3.4% 14C-PFOA in urine and 8.3% 14C-PFOA in 

feces, and the female mouse excreted 6.7% 14C-PFOA in urine and 5.7% 14C-PFOA in feces. The 

mice were similar in the amounts excreted. The male hamster excreted 90.3% and 8.2% 14CPFOA in urine and feces, respectively, and the female hamster excreted 45.3% and 9.3%14CPFOA. The male hamster excreted a greater amount of 14C-PFOA than the female hamster. Over 

84% of 14C-PFOA was excreted 24 hours after dosing by the male hamster compared to less than 

25% of 14C-PFOA excreted by the female hamster at 24 hours after dosing. Over 168 hours, the 

male rabbit excreted 76.8% and 4.2% 14C-PFOA in urine and feces, respectively, and the female 

rabbit excreted 87.9% and 4.6% 14C-PFOA. Both rabbits excreted most of the dose by 24 hours. 

The cumulative percentage of 14C-PFOA excreted is shown in Table 2-21. 

Table 2-21. Cumulative Percent 14C-PFOA Excreted in Urine and Feces

Species Gender

Hours After Dosing

12 24 48 72 96 120 168

Mouse Male 0.4 4.1 6.7 8.6 9.1 10.8 -

Female 0.2 4.1 6.5 8.4 9.0 11.0 -

Hamster Male 67.3 84.5 96.1 97.4 98.2 98.4 -

Female 11.3 24.6 36.4 43.9 50.1 54.0 -

Rabbit Male 77.8 80.2 80.4 80.4 80.4 80.4 80.4

Female 86.7 90.5 92.0 92.2 92.7 92.9 93.0

Source: Hundley et al. 2006

When the data in Table 2-21 are integrated with the data from rats, the gender differences in 

PFOA excretion rate appear to be species-specific. Female rats, male hamsters, and both genders 

of rabbits appear to be good excreters based on their response to a radiolabeled dose of 

10 mg/kg. Most of the dosed material is excreted within 24 hours after dosing. Female hamsters 

apparently are moderate excreters. Males and female mice excreted only about 10% of the dose 

over the 120 hours (5 days) after dosing. Mice do not show a gender difference but retain more 

of the dose than do hamsters, rabbits, and female rats. The long half-lives in humans suggest that 

their excretion rates are more like mice or male rats.

Inhalation Exposure 

Although no data were identified on urine or fecal excretion of PFOA following inhalation 

exposures, the Hinderliter study (2003) provides evidence of CL following single and repeated 

inhalation exposures in Sprague-Dawley rats. Plasma PFOA concentrations following a single 

exposure to 1, 10, and 25 mg/m3

 PFOA declined 1 hour after exposure in females and 6 hours 

after exposure in males. In females, the elimination of PFOA was rapid at all exposure levels 

and, by 12 hours after exposure, their plasma levels had dropped below the analytical LOQ

(0.1 μg/ml). In males, the plasma elimination was much slower and, at 24 hours after exposure, 

the plasma concentrations were approximately 90% of the peak concentrations at all exposure 

Perfluorooctanoic acid (PFOA) – May 2016 2-27
levels. In the repeated exposure study, male and female rats were exposed to the same 

concentrations for 6 hours/day, 5 days/week for 3 weeks. Steady-state plasma levels were 

reached in males by 3 weeks, but plasma PFOA levels in females returned to baseline with 

24 hours of each dose. The data are illustrative of distinct toxicokinetic differences between male 

and female rats in their response to PFOA exposure (Hinderliter 2003). 

Dermal Exposure

No data were identified on excretion following dermal exposures. Minimal fecal excretion is 

anticipated for the dermal route of exposure although the biliary pathway can be a route for 

excretion of material absorbed through the skin, distributed to the liver, and discharged to the 

gastrointestinal tract. 

2.5.1 Mechanistic Studies of Renal Excretion

Several studies have been conducted to elucidate the cause of the gender difference in the 

elimination of PFOA by rats. Many of the studies have focused on the role of transporters in the 

kidney tubules. Most studies have examined the OATs located in the proximal portion of the 

descending tubule. OATs are found in other tissues as well and were discussed earlier for their 

role in absorption and distribution. In the kidney, they are responsible for delivery of organic 

anions, including a large number of medications from the serum into the kidney tubule for 

excretion as well as reabsorption of anions from the glomerular filtrate. The transporters are 

particularly important in excretion of PFOA because it binds to surfaces of serum proteins 

(particularly albumin), which makes much of it unavailable for removal during glomerular 

filtration. Other transporter families believed to be involved in renal excretion are the OATPs 

and the MRPs. However, they have not been evaluated as extensively as the OATs for their role 

in renal excretion. 

OATs are located on both the basolateral (serum interface) and apical surfaces of the brush 

boarder of the proximal tubule inner surface. At the basolateral surface, the OATs transport the 

perfluorooctanoate anion from the serum to the tubular cells (Anzai et al. 2006; Cheng and 

Klaassen 2008; Klaassen and Aleksunes 2010; Klaassen and Lu 2008; Nakagawa et al. 2007, 

2009). OAT1, 2, and 3 are located on the basolateral membrane surface. OAT4 and OAT5 are 

located on the apical surface of the tubular cells, where they reabsorb the PFOA anions from the 

glomerular filtrate. Figure 2.2 diagrams the flow of organic anions such as the PFOA anion from 

serum to the glomerular filtrate for excretion and resorption of organic acids from the glomerular 

filtrate with transport back to serum. OATs can function for uptake into the cell across both the 

basolateral and apical surfaces. 

Several MRP transporters also appear to function in the kidney and move organic anions in 

and out of cells at both the basolateral surface (e.g., MRP2/4) and the apical surface (e.g., MRP1) 

as well as one or more OATPs on each surface (Cheng and Klaassen 2009; Klaassen and 

Aleksunes 2010; Klaassen and Lu 2008; Kusuhara and Sugiyama 2009; Launay-Vacher et al. 

2006; Yang et al. 2009). Bidirectional movement of PFOA across both the basolateral and apical 

surfaces is driven by concentration gradients and/or active transport. Far more data exist on 

PFOA and OATs in the kidneys than on OATPs and MRPs. Abbreviations for individual 

transporters on the basolateral and apical surfaces differ across publications. The accepted 

convention is to use uppercase letters to refer to human transporters and lowercase letters to refer 

to animal transporters. For this report, the data are not reported by species but by transporter 

family and the uppercase letters are used. 

Perfluorooctanoic acid (PFOA) – May 2016 2-28
Source: Klaassen and Aleksunes 2010

Figure 2-2. Localization of Transport Proteins 

Knowledge about specific OAT, OATP, and MRP transporters in the kidneys is rapidly 

evolving. A low membrane density or blockage of basolateral OATs will decrease PFOA 

excretion while low membrane densities or blockage of apical OATs will increase excretion 

because they decrease resorption of anions from the glomerular filtrate.

The earliest studies of the impact of gender on urinary excretion were conducted by 

Hanhijarvi et al. (1982) using probenecid, an inhibitor of renal excretion of organic acids on 

PFOA excretion in male and female Holtzman rats. The female rats that had not received the 

probenecid excreted 76% of the administered dose of PFOA over a 7-hour period, while males 

excreted only 7.8% of the administered dose over the same period of time. Probenecid 

administration modified the cumulative excretion curve for males only slightly. In females, 

however, probenecid markedly reduced PFOA elimination to 11.8%. The authors concluded that 

the female rat possesses an active secretory mechanism that rapidly eliminates PFOA from the 

body that male rats do not possess. 

Kudo et al. (2002) examined the role of sex hormones and OATs on the renal clearance 

(CLR) of PFOA. Renal mRNA levels of specific OATs in castrated male and ovariectomized 

(OVX) female Wistar rats also were determined. Castration of male rats caused a 14-fold 

increase in CLR of PFOA. The elevated PFOA CLR in castrated males was reduced by treating 

them with testosterone. Treatment of male rats with estradiol increased the CLR of PFOA. In 

female rats, ovariectomy caused a significant increase in CLR of PFOA (a twofold increase), but 

the administration of estradiol to OVX female rats returned CLR of PFOA to normal values. 

Treatments of female rats with testosterone reduced the CLR of PFOA.

Treatment with probenecid, a known inhibitor of OAT1–6 and OAT8, markedly reduced the 

CLR of PFOA in male rats, castrated male rats, and female rats (Kudo et al. 2002). Accordingly, 

the male sex hormones appear to decrease the presence of OATs in the renal basolateral 

membranes while the female sex hormones appear to increase the transporters. 

Perfluorooctanoic acid (PFOA) – May 2016 2-29
To identify the transporter molecules responsible for PFOA transport in the rat kidney, renal 

mRNA levels of specific OATs were determined in male and female rats under various hormonal 

states and compared with the CLR of PFOA. The level of OAT2 mRNA in male rats was only 

13% of the level in female rats. Castration or estradiol treatment increased the level of OAT2 

mRNA whereas treatment of castrated male rats with testosterone reduced it. Ovariectomy of 

female rats significantly increased the level of OAT3 mRNA. Multiple regression analysis of the 

data suggested that OAT2 and OAT3 are responsible for urinary elimination of PFOA in the rat; 

however, the possibility of a resorption process mediated by OATP1 was mentioned as a 

possible factor in male rat retention of PFOA. OAT2 and OAT3 are located on the basolateral 

cell surface. OATP1 is located on the apical surface of the renal tubule cells (Kudo et al. 2002). 

Cheng et al. (2006) examined whether sex hormones influenced gender-specific OATP

expression in the kidneys of adult male and female C57BL/6 mice. Gonadectomized mice were 

used for the studies in conjunction with hormone replacement measures (5α-dihydroxytestosterone [DHT] or 17-β estradiol [E2]). OATP1a1 and OATP3a1 were evaluated. Treatment 

with DHT resulted in significant increase in both OATPs in the kidneys of male and female 

gonadectomized mice. In both cases, the change in males was greater than the change in females. 

Treatment with E2 almost abolished the expression of OATP1a1 in the kidneys but caused no 

significant change in OATP3a1. In the intact control animals, almost no expression of OATP1a1 

occurred in the kidneys of females and a significantly lower expression of OATP3a1 (p<0.05) 

occurred. In the gonadectomized control animals, little or no expression of OATP1a1 occurred in 

either gender, and expression of OATP3a1 was equivalent in both genders.

Nakagawa et al. (2007) investigated the role of OATs in the renal excretion of PFOA using 

in vitro methods. HEK293-transformed cells, derived from human embryonic kidney (HEK), 

were transfected with human or rat OAT1, OAT2, or OAT3 constructs. Cells from the S2 

segment of the proximal tubule were transfected with human or rat OAT2 constructs. HEK293 

and S2 cells transfected with the vector served only as control cells. The transfected HEK293 

cells were incubated for 1 min with or without 0, 10, and 100 μmol [14C]PFOA and/or varying 

concentrations of favored OAT substrates to determine inhibitory effects of PFOA as follows: 

5 μmol [14C]para-aminohippuric acid (OAT1), 20 nmol [14C]estrone sulfate (OAT3), and 

10 nmol [14C]prostaglandin F2α (OAT2). 

PFOA significantly inhibited para-aminohippuric acid and estrone sulfate uptake mediated by 

OAT1 and OAT3, respectively. At 10 μmol PFOA, uptake of 5 μmol [14C] para-aminohippuric 

acid was 75–85% of the control level and, at 100 μmol PFOA, uptake was reduced to 35–45% of 

control. Estrone sulfate uptake by human OAT3 was 65% of the control level at 10 μmol PFOA 

and 40% of control at 100 μmol PFOA. Estrone sulfate uptake by rat OAT3 was 15% of the 

control level in the presence of 10 μmol PFOA and was almost completely inhibited at 100 μmol

PFOA. Prostaglandin F2α uptake by OAT2 was inhibited moderately by PFOA, 75–85% of 

control at 10 μmol PFOA, and 65% of control at 100 μmol PFOA.

In the second part of their study, Nakagawa et al. (2007) incubated HEK293 and S2 

transfected cells with 10 μmol [14C]PFOA for 1 min to determine uptake. Time-dependent 

uptake of 5 μmol [14C]PFOA from 0 to 30 mins was conducted in the HEK293 cells transfected 

with human or rat OAT1, OAT2, or OAT3. Experiments were conducted in triplicate. Uptake of 

PFOA was stimulated (p<0.001) in cells transfected with human or rat OAT1 or OAT3, while no 

uptake was stimulated in cells transfected with OAT2 in either cell line. In the time-dependent 

experiments, uptake by human or rat OAT1 or OAT3 increased linearly up to 2 mins and reached 

a plateau in about 15 mins. Kinetic evaluations resulted in substrate concentration at which the 

Perfluorooctanoic acid (PFOA) – May 2016 2-30
initial reaction rate is half maximal (Km) values of 48.0, 51.0, 49.1, and 80.2 μmol for human 

OAT1, rat OAT1, human OAT3, and rat OAT3, respectively. The authors showed that both 

human and rat OAT1 and OAT3 transport PFOA in the kidney while human and rat OAT2 do 

not (Nakagawa et al. 2007). 

Yang et al. (2009) investigated the role of OAT polypeptide 1a1 (OATP1a1) in the renal 

elimination of PFOA. The polypeptide is located on the apical side of proximal tubule cells and 

could be the mechanism for renal reabsorption of PFOA in rats. The level of mRNA of 

OATP1a1 in male rat kidney is 5–20-fold higher than in female rat kidney, OATP1a1 protein 

expression is higher in male rat kidneys, and it is regulated by sex hormones. One of its known 

substrates is estrone-3-sulfate (E3S). A substantial presence of OATP1a1 in male rats would 

favor resorption of PFOA in the glomerular filtrate and reduce excretion. 

Chinese hamster ovary (CHO) cells were transfected with rat OATP1a1 complementary 

DNA. The transfected CHO cells were incubated with 4 μmol [14C]PFOA for up to 10 mins or 

with 0–1,000 μmol [14C]PFOA for 2 mins to determine uptake. The difference between the 

uptake velocities of CHO OATP1a1-transfected cells and CHO vector-transfected cells was 

defined as active PFOA uptake by the tubular epithelium. The transfected CHO cells were 

incubated with 5 μmol [14C]PFOA for 2 mins in the absence or presence of inhibitors (e.g., BSP, 

taurocholate, probenecid, p-aminohippurate, and naringin [a flavonoid found in grapefruit]) for 

inhibition studies. The transfected CHO cells were incubated with 2 μmol E3S and 0, 0.1, or 

1 mM perfluorocarboxylates with carbon chain lengths ranging from 4 to 12, including PFOA 

(C8) for 30 seconds for E3S inhibition studies. 

In time-dependent uptake experiments, uptake of PFOA by OATP1a1-transfected cells 

increased proportionally to time during the first 2 mins of incubation. Vector-transfected cells 

had a significant level of uptake of PFOA attributed to nonspecific passive diffusion. In the 

concentration-dependent uptake experiments, uptake velocity of PFOA in OATP1a1-transfected 

cells increased with increasing concentration and saturation levels were not reached. In vectortransfected cells, uptake velocities increased linearly with increasing concentration of PFOA, 

demonstrating a passive diffusion mechanism. Active PFOA uptake—the difference between the 

uptake of the OATP1a1 cells and the vector-transfected cells—could be described by the 

Michaelis-Menton equation and exhibited saturable kinetics. 

Inhibition experiments with substrates of OATs and OATPs showed that BSP, taurocholate, 

and naringin inhibited PFOA uptake to 10–30% of control and p-aminohippurate inhibited PFOA 

uptake to 62% of control. Probenecid, an OAT inhibitor, did not inhibit PFOA uptake at all. In 

OATP1a1-transfected cells, uptake of E3S was inhibited to less than 10% of control uptake 

following incubation with 1 mM [14C]PFOA. Inhibition of E3S was less than 50% of control 

uptake after incubation with 0.1 mM [14C]PFOA. Based on the results of the uptake and 

inhibition experiments, the authors suggested that passive diffusion could be an important route 

of PFOA distribution and that renal reabsorption in the male rat could be mediated by OATP1a1. 

Nakagawa et al. (2009) investigated the role that the human organic acid transporter (OAT4) 

plays in transporting PFOA. Human OAT4 is located on the apical side of proximal tubule cells 

and mediates reabsorption of organic anions. Transformed cells derived from HEK cells, 

HEK293, were transfected with human OAT1, OAT3, or OAT4 constructs. HEK293 cells 

transfected with only the vector served as control cells. The transfected HEK293 cells were 

incubated with 10 μmol [14C]PFOA for 15 mins to determine uptake. Transfected cells also were 

incubated with 10 μmol [14C]PFOA for 15 mins and then washed with incubation medium 

Perfluorooctanoic acid (PFOA) – May 2016 2-31
containing 1%, 3%, and 5% BSA to investigate the contribution of nonspecific binding of PFOA 

on the cell membrane. Experiments were conducted in triplicate.

Uptake of PFOA was significantly stimulated (p<0.01) in cells transfected with human 

OAT1, OAT3, and OAT4. Uptake of PFOA in human OAT1 transfected cells was 1.6-fold 

higher than in control cells. In human OAT3 transfected cells, PFOA uptake was ~2.4-fold 

higher than in control cells. In human OAT4 transfected cells, PFOA uptake was 2.7-fold higher 

than in control cells. Accumulation of PFOA in transfected human OAT4 cells also was 

significantly greater than in human OAT1 cells (p<0.01). Washing the cells with BSA reduced 

PFOA uptake by 30% at most, suggesting mediation by the transporters into the transfected cells. 

The experiments showed that human OAT4 transports PFOA and that human OAT4 activity 

might play a role in reabsorption of PFOA from the tubule, resulting in poor urinary excretion. 

Yang et al. (2010) examined cellular uptake of PFOA by OATP1A2, OAT4, and urate 

transporter 1 (URAT1) to determine their roles in mediating human renal reabsorption. CHO and 

HEK293 cells were transfected with OATP1A2, OAT4, and URAT1 plasmid DNA or vector 

DNA (control). In uptake studies, PFOA incubation times were 10 seconds (OAT4) and 

30 seconds (URAT1). Cells transfected with OAT4 were incubated with 5 μmol PFOA for up to 

1 min in time-dependent uptake experiments. In inhibition studies, cells transfected with OAT4 

were incubated with 5 μmol [14C]PFOA for 10 seconds in the presence and absence of 100 μmol

sulfobromophthalein (BSP), probenecid, glutarate, or polycyclic aromatic hydrocarbon (PAH). 

Perfluorinated carboxylates with differing chain lengths (C4–C12) were used in chain lengthdependent inhibition experiments. Incubations with 3

H-E3S (OAT4 and OATP1A2) or 6 μmol

C14-uric acid (URAT1) in the presence and absence of 100 μmol perfluorinated carboxylate 

lasted 10 seconds (OAT4), 30 seconds (OATP1A2), and 1 min (URAT1). 

PFOA uptake in OATP1A2-transfected HEK293 cells was no different than uptake in control 

cells. At 100 μmol, E3S uptake was inhibited ~30% by PFOA (C8), ~62% by C9, ~70% by C10, 

~42% by C11, and ~18% by C12. E3S uptake was not inhibited by C4–C7. In CHO cells 

transfected with OAT4, time-dependent uptake experiments showed a saturation phase after an 

incubation time of approximately 10 seconds. A pH-dependent increase in PFOA uptake was 

observed with approximately 90% uptake at pH 8 and 250% at pH 5.5. 

In concentration-dependent uptake experiments, uptake increased with increasing PFOA 

concentration (0–1000 μmol) in OAT4-transfected CHO cells at pH 7.4 and 6. PFOA uptake was 

cis-inhibited by BSP and probenecid and trans-stimulated by PAH and glutarate at pH 7.4. A 

chain length-dependent effect was observed in E3S inhibition on OAT4-expressing cells in the 

presence of C7 (30%) through C10 (~80%). Inhibition in the presence of C11 and C12 were 

~52% and ~30%, respectively. Inhibition of E3S in the presence of C4, C5, and C6 was less than 

20% for each. 

PFOA uptake in HEK293 cells transfected with URAT1 was not statistically different from 

control cells in the presence and absence of Cl-

. Under both conditions, PFOA intake was 

enhanced especially in the absence of Cl- in which PFOA uptake was greater than fourfold 

compared to uptake in control cells. Time-dependent PFOA (5 μmol) uptake by URAT1 

increased with time during the 5-min incubation period, and a concentration-dependent increase 

in PFOA uptake was observed (0–700 μmol). Urate uptake was inhibited in a chain lengthdependent manner. Inhibition in the presence of C7–C10 was ~70% each, ~60% in the presence 

of C6 and C11, ~50% in the presence of C5, ~30% in the presence of C12, and ~25% in the 

presence of C4. Based on the results, Yang et al. (2010) concluded that PFOA was not a 

Perfluorooctanoic acid (PFOA) – May 2016 2-32
substrate for OATP1A2, but that OAT4 and URAT1 were probably involved in the renal 

reabsorption of PFOA. 

Weaver et al. (2010) published in vitro studies on the transport activities of the rat renal 

transporters OAT1, OAT2, OAT3, OATP1a1, and URAT1. The transporters were transfected 

into one of several cell lines and exposed to a series of perfluorinated carboxylates having chain 

lengths ranging from 2 to 18 carbons (C). The activity of the perfluorinated carboxylate on the 

transporters was quantified on the basis of its ability to inhibit the transport of a favored 

radiolabeled substrate. The PFAS inhibition of the individual transporters varied with chain 

length. The perfluorinated carboxylate with 6, 7, and 8 carbon chains caused a significant 

decrease in OAT1 transport of tritiated p-aminohippurate, with the C7 acid having the strongest 

effect. The perfluorinated carboxylates with 5 through 10 carbon chains caused a significant 

decrease in transport of tritiated E3S by OAT3, with C8 and C9 acids having the strongest effect. 

The transport of tritiated estadiol-17β-glucuronide by OATP1a1 was significantly inhibited by 

perfluorinated carboxylates with 6 through 11 carbon chains, with C10 acid having the strongest 

effect. The perfluorinated carboxylate did not inhibit OAT2 or URAT1 transport of favored 

substrates.

The kinetic response of the OAT1, OAT3, and OATP1a1 transporters to increasing 

concentrations of selected perfluorinated carboxylates also was evaluated by Weaver et al. 

(2010). The change in transport velocity (ng/mg protein/min) with increasing concentrations of 

the perfluorinated carboxylate exhibited a Michaelis-Menton-type response. The kinetic data 

were analyzed to determine the Km and Vmax, and the results are summarized in Table 2-22 

below.

Table 2-22. Kinetic Parameters of Perfluorinated Carboxylate Transport by OAT1, OAT3, 

and OATP1a1 

Transporter PFAS Km (μmol) Vmax (nmol/mg protein/min)

OAT1 C7 50.5 ± 13.9 2.2 ± 0.2

C8 43.2 ± 15.5 2.6 ± 0.3

OAT3 C8 65.7 ± 12.1 3.8 ± 0.5

C9 174.5 ± 32.4 8.7 ± 0.7

OATP1a1 C8 126.4 ± 23.9 9.3 ± 1.4

C9 20.5 ± 6.8 3.6 ± 0.5

C10 28.5 ± 5.6 3.8 ± 0.3

Source: Weaver et al. 2010

The Michaelis-Menton kinetic data (Km and Vmax [maximum initial rate of an enzyme 

catalyzed reaction]) indicate that there are substantial differences in the affinity of the 

perfluorinated carboxylate with 8 and 9 carbon chains for OAT3, with the C8 acid favored over 

the C9 acid. OAT3 is an export transporter located on the basolateral side of the tubular cells; 

thus, when present in a mixture consisting of comparable concentrations of both, renal tubular 

excretion of the C8 acid would tend to decrease excretion of the C9 acid. For OATP1a1, a 

resorption transporter located on the apical side of the renal tubular cells, the C9 and C10 acid 

have a greater affinity for the transport protein than the C8 acid. The kinetic data suggest that the 

net impact of these relationships would be to favor excretion of the C8 acid over the C9 acid and 

possibly the C10 acid when all three fluorocarbons are present in the exposure matrix at 

approximately equal concentrations. There were minimal kinetic differences between transport 

Perfluorooctanoic acid (PFOA) – May 2016 2-33
of the C7 and C8 acids by OAT1, an export transporter on the basolateral surface of the renal 

tubular cells.

Based on the Hinderliter study (2004), a developmental change in renal transport occurs in 

female rats between 3 and 5 weeks of age that allows for expedited excretion of PFOA. When 

the transporters become active, there is a decrease in plasma PFOA levels and an increase in 

urinary excretion (Table 2-23). The developmental change in male rats appears to have the 

opposite effect. Sexual maturity appears to influence these events because castrated male rats 

become more like females and OVX females become more like males in their PFOA excretion 

capabilities. The change in female rats seems to involve the OATs (Kudo et al. 2002) while the 

change in males seems to involve the OATPs (Cheng et al. 2006). 

Table 2-23. Plasma and Urine PFOA Concentration 24-hr After Treatment

with 30 mg/kg PFOA

Age

(weeks)

Female Male

Plasma (μg/ml) Urine (μg/ml) Plasma (μg/ml) Urine (μg/ml)

3 51.43 ± 13.61 94.89 ± 26.26 74.16 ± 18.23 51.76 ± 28.86

4 28.01 ± 9.90 104.12 ± 28.97 100.81 ± 13.18 28.70 ± 18.84

5 3.42 ± 1.95 123.16 ± 51.56 113.86 ± 23.36 15.65 ± 6.24

Source: Hinderliter 2004

When considered together, the studies of the transporters suggest that female rats are efficient 

in transporting PFOA across the basolateral and apical membranes of the proximal kidney 

tubules into the glomerular filtrate, but male rats are not. Males, on the other hand, have a higher 

rate of resorption than females for the smaller amount they can transport into the glomerular 

filtrate via OATP1a1 in the apical membrane. This scenario might explain the inverse 

relationship between the levels of PFOA in female urine and plasma and the plateau of plasma 

PFOA in male rats compared to their losses via urine.

Unfortunately, much work remains to be done to explain the gender differences between 

male and female rats and to determine whether it is relevant to humans. Similarities are possible 

because the long half-life in humans suggests that they might be more like the male rat than the 

female rat. There is a broad range of half-lives in human epidemiology studies suggesting a 

variability in the unbound fraction of PFOA in serum or in human transport capabilities resulting 

from genetic variations in structures and consequently in function. Genetic variations in human 

OATs and OATPs are described in a review by Zaïr et al. (2008). 

2.6 Toxicokinetic Considerations 

2.6.1 PK Models

One of the earliest PK models was done using the post-dosing plasma data from the 

Butenhoff et al. study (2004b) in cynomolgus monkeys (Andersen et al. 2006). In this study, 

groups of six monkeys (three per gender per group) were dosed for 26 weeks with 0, 3, 10, and 

20 mg/kg PFOA (high-dose =30 mg/kg PFOA for the first 12 days), followed for >160 days after 

dosing. Metabolism cages were used for overnight urine collection. Since urine specimens could 

account for only overnight PFOA excretion, the total volume and total PFOA were extrapolated 

to 24-hour values based on the excretion rate (volume/hour) for the volume collected and the 

hours of collection. 

Perfluorooctanoic acid (PFOA) – May 2016 2-34
The Andersen et al. model (2006) was based on the hypothesis that saturable resorption 

capacity in the kidney would possibly account for the unique half-life properties of PFOA across 

species and genders. The model structure, shown in Figure 2-3, was derived from a published 

model for glucose resorption from the glomerular filtrate via transporters on the apical surface of 

renal tubule epithelial cells (Andersen et al. 2006). 

Figure 2-3. Schematic for a Physiologically Motivated Renal Resorptions PK Model.

The renal-resorption model includes a central compartment that receives the chemical from 

the oral dose and a filtrate compartment for the glomerular filtrate from which resorption and 

transfer to the central compartment can occur. Transfer from the filtrate compartment to the 

central compartment decreases the rate of excretion. The resorption in the model was saturable, 

meaning that there was less resorption and greater excretion at high serum PFOA concentrations 

than at low concentrations. 

The model was parameterized using the body weight and urine output of cynomolgus 

monkeys (Butenhoff et al. 2004b) and a cardiac output of 15 L/h-kg from the literature (Corley et 

al. 1990). A 20% blood flow rate to the kidney was assumed based on data from humans and 

dogs. Other parameters were optimized to fit the data for plasma and urine at lower 

concentrations and then applied for the 20 mg/kg/day dose, which was assumed to represent a 

concentration at which renal resorption was saturated. Based on the data for the dose of 

20 mg/kg/day, the model was able to predict the decline in plasma levels after the cessation of 

dosing. The predictions were fairly adequate for one of the three modeled monkeys; for the other 

two monkeys, the model predicted higher levels than were observed. That result could have 

occurred because the model did not allow for efflux of PFOA into the glomerular filtrate through 

transporters on the basolateral surface of the tubular cells. The authors observed that three 

monkeys had faster CLR of PFOA than the other three monkeys. 

Tan et al. (2008) divided the second compartment in the Andersen et al. model (2006) into a 

liver compartment and a tissue compartment. A storage compartment was added between the 

filtrate compartment and urinary excretion (Figure 2-4) (Tan et al. 2008). 

Perfluorooctanoic acid (PFOA) – May 2016 2-35
Figure 2-4. Physiologically Motivated Pharmacokinetic Model Schematic for

PFOA-Exposed Rats 

The models were parameterized and applied to the Kemper data (2003) for male and female 

CD rats given doses of 1, 5, and 25 mg/kg/day. The model did not provide a satisfactory fit 

between the predictions of plasma concentration or urine + fecal excretion and experimental data 

for either gender. 

Lou et al. (2009) used the data they collected on the serum, liver, and kidney PFOA 

concentration (see section 2.2.1) in CD-1 mice to examine if one- or two-compartment PK 

models would fit the experimental data for 1, 10, and 60 mg/kg/day single gavage doses (see 

Figure 3-5 for the one-compartment model). Both models assumed first order absorption and 

elimination. The two-compartment model included a central compartment that received PFOA 

after absorption and transferred it to a second compartment for excretion. The excretion 

compartment was coupled with bidirectional flow between the two compartments. The net loss 

from the central compartment differed during and after distribution. The models were fit using a 

general nonlinear least squares approach. A likelihood ratio squared approach was applied to 

determine which model achieved the best fit to the data.

Perfluorooctanoic acid (PFOA) – May 2016 2-36
Figure 2-5. Schematic for One-Compartment Model.

The one compartment model performed well for serum, liver, and kidney in this analysis, and 

output was not significantly improved with use of a two-compartment model. The input 

parameters for the one-compartment model included Vd, serum half-life, and absorption rate 

constant (Ka) and elimination rate constant (Ke) for serum, liver, and kidney. There were slight 

differences in the fitted values between males and females for some parameters. The Ke values 

were consistently higher in the female mice (Table 2-24). The quantitative measures for liver and 

kidney were only available for the 1- and 10-mg/kg/day doses. 

Table 2-24. Model Parameters for 1 and 10 mg/kg Single Doses of PFOA to CD1 Mice 

Tissue Parameter (abbreviation) Females Males

Serum Volume of distribution (Vd) 0.135 L/kg 0.266 L/kg

Absorption rate constant (Ka) 0.537 L/hr

Elimination rate constant (Ke) 0.00185 L/hr 0.00133 L/hr

Half-life (T1⁄2) 15.6 days 21.7 days

Liver Volume of distribution (Vd) 0.161 L/kg 0.120 L/kg

Absorption rate constant (Ka) 0.5170 L/hr

Elimination rate constant (Ke) 0.00161 L/hr 0.00129 L/hr

Kidney Volume of distribution (Vd)—1 mg/kg 0.822 L/kg 1.280 L/kg

Volume of distribution (Vd)—10 mg/kg 1.092 L/kg 1.170 L/kg

Absorption rate constant (Ka) 0.527 L/hr

Elimination rate constant (Ke) 0.00151 L/hr 0.00113 L/hr

Source: Lou et al. 2009

The one-compartment model described above was not able to predict serum concentration in 

female mice given a single 60-mg/kg dose, suggesting a change in kinetics with the 60-mg/kg 

dose compared to the 1- and 10-mg/kg doses. This conclusion is supported by comparison of the 

serum measurements made during the 30-day post-dosing period for all three doses. The serum 

PFOA concentration at the 60-mg/kg dose declined more rapidly with time than serum PFOA 

concentrations at the 1- and 10-mg/kg doses. For example, a serum concentration of about 

0.4 mg/L was reached in about 28 days at the 60-mg/kg dose, 61 days at the 10-mg/kg dose, and 

70 days at the 1-mg/kg dose (values estimated from Figure 3, Lou et al. 2009). The oneCentral 

Compartment

 

Oral Dose

Ka

Urinary Loss

Ke

Perfluorooctanoic acid (PFOA) – May 2016 2-37
compartment model also produced a poor fit for the serum level measurements taken 24 hours 

after the cessation of a 17-day exposure to 20 mg/kg/day. The two-compartment model provided 

a better fit with experimental serum concentration data for the single 60-mg/kg dose and the 

repeat 20-mg/kg/day dose, but the fit was still unsatisfactory.

Lou et al. (2009) also tried the Andersen et al. renal-resorption model (2006) to determine if 

it provided an improved fit for the data. The Andersen et al. model (2006) fit to the data was 

superior to that of the one- and two-compartment models of Lou et al. (2009) for the 60-mg/kg 

single-dose and the 20-mg/kg/day repeat-dose scenarios.

The Andersen et al. model (2006) includes a second tissue compartment that articulates with 

the central compartment but not the filtrate compartment. In addition to values for Vd, Ka, and 

Ke, the model includes values for cardiac output, volume for the renal filtrate, renal blood 

filtration rate, intercompartmental CL, transport maximum, transport affinity constant (Kt), and 

the proportion of free PFOA in serum. With the exception of body weight and cardiac output, the 

input parameters for the model were either assumed (i.e., volume of renal filtrate and proportion 

of free serum PFOA) or optimized for the model. The wide confidence bounds around the 

optimized values are indicative of considerable parameter uncertainty.

The Lou et al. parameter estimates (2009) indicate that there may be several biological 

limitations to the Andersen et al. (2006) PK model for adult mice including the fact that it 

requires an unreasonably high portion of the cardiac output to pass through the kidneys to 

optimize fit to the experimental data. It also does not include excretion via export transporters in 

the renal tubular cells or consider that the bound fraction in the serum could vary with the 

magnitude of the dose and duration of dosing. Much of the emerging data is consistent with a 

variety of tubular transporters functioning in both efflux and resorption from the glomerular 

filtrate. In addition, there are opportunities for protein binding within organs that could function 

to retard distribution to the cytosol, especially at low doses. The binding of PFOA with L-FABP

is an example. Once binding sites are saturated, the concentration in the cytosol will increase. 

A model also has been developed that applied to female CD-1 mice during gestation and 

lactation (Rodriguez et al. 2009). The gestational model includes two compartments, one for the 

dam and the other for the litter. They are linked by placental blood flow. The biological data 

used to set the parameters for the two compartments were based on the data from the Lau et al. 

(2006) and Abbott et al. (2007) studies in CD-1 and 129S1/SvlmJ mice, respectively. Exposure 

was assumed to be limited by blood flow, and only the experimental doses that did not impact 

litter size (i.e., 0.1–1.0 mg/kg/day for CD-1 mice and 1–10 mg/kg/day for 129S1/SvlmJ mice) 

were used in model development. 

Lactational exposure was modeled as a dynamic relationship between the dam (n = 10) and 

the litter, and they were connected by a milk compartment. Milk yield information was obtained 

from the literature. Milk was assumed to be consumed as it was produced without any circadian 

impact on consumption patterns. PFOA excreted in pup urine was routed back to the dam. 

Both absorption and excretion were assumed to be first order processes as was lactation 

transfer from the dam to the litter (Figure 2-6) (Rodriguez et al. 2009). Resorption of a portion of 

the PFOA urinary efflux was included in the model. The renal excretion/resorption was 

parameterized for cardiac output, kidney blood flow, GFR, urine flow rate, volume of renal 

plasma (fraction of body weight), and volume of renal filtrate (fraction of body weight). The 

fraction of free PFOA in serum reaching the glomerulus was assumed to be 0.01 based on 

Perfluorooctanoic acid (PFOA) – May 2016 2-38
protein binding information. As was the case with the Lou et al. model (2009), Rodriguez et al. 

(2009) did not include parameters to adjust for transporter-mediated efflux from the renal tubular 

cells into the glomerular filtrate. 

Figure 2-6. PK Model of Gestation and Lactation in Mice

One of the limitations of the Rodriguez et al. modeling effort (2009) was the limited amount 

of laboratory data against which to evaluate projections. Serum measures from the Lau et al. 

(2006) and Abbott et al. (2007) studies were available for only a few time points. Nevertheless, 

the authors reached several conclusions based on the model projections as follows:

• The model had a tendency to overestimate serum levels, suggesting nonlinearity as doses 

increased.

• Gestation and lactation as a source of exposure contributed about equally to the pups of 

129S1/SvlmJ dams exposed only during gestation. 

• The contributions to the pups from gestation exceeded those from lactation in the CD-1 

mice.

• Exposure to the pups via lactation increased over time. 

• Lactation is a CL pathway for the dam. 

A number of uncertainties accompany the model because of the assumptions regarding the 

flow limitation on transport to the fetus and to maternal milk: the first order Ke for the pups and, 

for the milk, the maternal serum partition coefficient and the limited knowledge regarding the 

renal tubular transporters. They caution that the model should not be applied for cross-species or 

high-to-low dose extrapolation.

Loccisano et al. (2011) developed a PFOA PBPK model for monkeys based on the Andersen 

et al (2006) and Tan et al. (2008) models, and then extrapolated it for use in humans (Figure 

2-7). The model reflects saturable renal absorption of urinary PFOA by the proximal tubule of 

Perfluorooctanoic acid (PFOA) – May 2016 2-39
the kidney. This is represented in Figure 2-7 by the interactions between the plasma and kidney 

plus the interaction of the filtrate compartment with both plasma and kidney. 

The fraction of PFOA free in plasma and available for glomerular filtration was based on 

data fit and estimated to be less than 10% because of binding to serum proteins, especially 

albumin. Lacking the kinetic data on tubular resorption, the rate was based on the best fit to the 

plasma/urine data. A storage compartment was added to the model between filtrate and urine. 

Tissue plasma partition coefficients were derived from the data by Kudo et al. (2007) following 

the disposition pattern of a single intravenous (IV) dose to male Wistar rats. 

Notes: 

Tm = transporter maximum, Kt = affinity constant, and Q = flow in and out of tissues. 

Figure 2-7. Structure of the PFOA PBPK Model in Monkeys and Humans

Existing IV and oral data sets from Butenhoff et al. (2004b) for the cynomolgus monkey 

were used to develop the monkey model. In the oral study (section 2.2.1), animals were dosed for 

6 months and followed for 90 days after dosing. Plasma and urine samples were analyzed 

periodically during dosing and recovery. The model projections for the oral study were in good 

agreement with the Butenhoff et al. data (2004b) for the 10-mg/kg dose, showing a rapid rise to 

plasma steady state and a slow terminal half-life. The model performance for the high dose 

(30/20 mg/kg/day) did not fit as well, partially as a consequence of the observed toxicity with the 

initial 30 mg/kg/day dose that necessitated cessation of dosing on study day 12, followed by 

resumption of dosing at 20 mg/kg/day on study day 22. 

The structure of the human model was similar to that used for the monkeys. Human serum 

data (means with standard deviations [SDs] or medians) for PFOA are available for occupational 

and general populations (Bartell et al. 2010; Calafat et al. 2007a, 2007b; Emmett et al. 2006; 

Perfluorooctanoic acid (PFOA) – May 2016 2-40
Hölzer et al. 2008; Olsen et al. 2005; Steenland et al. 2009). The fact that the serum data were 

the results from measurements made following uncertain routes and uncertain exposure durations 

presented a challenge in the assessment of model fit. The human half-lives used for the model 

(3.8 and 2.3 years) came from an occupational study (Olsen et al. 2005) and a study of the Little 

Hocking, Ohio, population after reduction of the PFOA in drinking water as a result of treatment 

(Bartell et al. 2010). See section 2.6.2. Both half-life values were used in evaluating the model’s 

ability to predict serum concentration at the time the serum samples were collected.

The model produced results that can be characterized as fair to good when compared to the 

reported average serum measurements. For the Little Hocking population studied by Emmett et 

al. (2006), the model indicated the need for a 30-year exposure to reach steady-state 

concentrations. The model indicated that both half-life values provided reasonable results when 

compared to the measured serum values. The authors concluded that more data are needed on the 

kinetics of renal transporters and intrahuman variability, plus more definitive information on 

exposures to further refine the human model. 

Fàbrega et al. (2014) adapted the Loccisano et al. model (2011) to include compartments for 

the brain and lung, and to remove the skin. They applied the adjusted model to humans by using 

intake and body burden data from residents in Tarragona County, Spain. Food and drinking 

water were the major sources of exposure. Body burden information came from blood samples 

from 48 residents; tissue burdens came from 99 samples of autopsy tissues. The adjusted model 

overpredicted PFOA serum levels by a factor of about 9, the liver by a factor of 4.5, and the 

kidney by a factor of about 18. Model predictions for PFOS were far more consistent with the 

tissue concentration experimental data.

The authors also looked at the impact of using data for partition coefficients from human 

tissues in place of the Loccisano et al. rat data (2011) for the estimation of steady-state tissue 

concentrations. The PFOA simulation values were closer to the human experimental data when 

using the human partition coefficient values for the brain and lung, but not for the liver. In the 

case of the kidney, the simulated projections were generally equivalent with both the human and 

rat partition coefficients. The authors suggested that both saturable resorption and variations in 

protein binding are important parameters for PK models. With the exception of serum albumin, 

the existing models have not considered protein-binding constants within tissues. Even though 

the use of human partition coefficients improved the steady-state predictions for tissues, overall 

there were still considerable differences between the experimental values and the predictions 

with both models. 

Loccisano et al. (2012a) also used the saturable resorption hypothesis when developing a 

model for adult Sprague-Dawley rats (Figure 2-8). The structure of the model is similar to that 

for the monkey/human model depicted in Figure 2-7, but lacks the fat and skin compartments 

and includes a storage compartment to accommodate fecal loss of unabsorbed dietary PFOA as 

well as loss from biliary secretions. Oral and IV data used in model development came from 

studies by Kemper (2003), Kudo et al. (2007) and Perkins et al. (2004). Partition coefficients for 

liver:plasma, kidney:plasma, and rest of the body:plasma were derived from unpublished data on 

mice by DePierre (2009) through personal communication with the authors (Loccisano et al. 

2012a). Most of the other kinetic parameters were based on values providing the best fit to the 

experimental data. Because a number of the renal transporters involved with PFOA resorption 

are known, available kinetic information was used where appropriate. Model performance was 

evaluated primarily based on its ability to predict plasma and liver concentrations from the 

Perfluorooctanoic acid (PFOA) – May 2016 2-41
studies identified above. Performance was generally good given the limitations in the primary 

data sources, as was the case for the monkey model. 

Figure 2-8. Structure of the PBPK Model for PFOA in the Adult Sprague-Dawley Rat

Loccisano et al. (2012b) expanded the adult Sprague-Dawley rat model to cover gestational 

and lactational exposure to the fetus and pups through their dams. The data from Hinderliter et 

al. (2005) were used in model development for both the gestation and lactation periods. The 

gestational model structure for the dams is similar to the model structure shown in Figure 2-8. 

The model was expanded to include the fetuses linked to the dams by way of the placenta. 

Uptake from the placenta was described by simple diffusion; the fetal plasma compartment was 

separate from the dams as was distribution to fetal tissues and amniotic fluid. Based on the 

transporter data for PFOA, elimination differed for male and female rats and was considered to 

be developmentally regulated, resulting in faster elimination for female rats than for male rats 

after sexual maturation. The lactation model linked the pups to their dams through mammary 

gland secretions. Pup compartments included the gut, liver, kidney, renal filtrate, plasma, and 

rest of the body. 

Model performance was judged by its ability to predict concentrations in maternal and fetal 

plasma, amniotic fluid, and milk. The predictive capability of the model ranged from fair to 

good, depending on the medium. The fit of the projections to the data was weakest for the whole 

embryo during gestation, for which measured levels were greater than projection for two of three 

data points and for neonate plasma during lactation, for which all data points fell below the 

predictions. 

Loccisano et al. (2013) extended their model development to cover humans during pregnancy 

and lactation, building on the work done with rodents and recognizing the limitations of the 

human data available for evaluating the model predictions. Figure 2-9 illustrates the structure of 

Perfluorooctanoic acid (PFOA) – May 2016 2-42
the model used. The basic structure was derived from the rat model discussed above. Following 

are some of the key features of the model: 

• The fetus is exposed via the placenta through simple bidirectional diffusion. 

• Transfer rates to the fetus from the amniotic fluid are governed by bidirectional diffusion. 

• Transfer from the fetal plasma to tissues is flow-limited. 

• Maternal plasma is directly linked to the milk compartment and considered to be flowlimited; only the free fraction in plasma is transferred to maternal milk. 

• The neonate is exposed to PFOA only via maternal milk for the first 6 months 

postpartum. 

• The infant in the model is treated as a single compartment with a Vd. 

Figure 2-9. PBPK Model Structure for Simulating PFOA and PFOS Exposure During 

Pregnancy in Humans (Maternal, Left; Fetal, Right) 

Limitations to the model are acknowledged and attributed primarily to lack of data to support 

a more mechanistic approach. Physiological parameters applicable to a pregnant or lactating 

woman, the fetus, and the nursing infant were obtained from a variety of referenced publications. 

To obtain a plasma value at the time of conception, the model was run until it reached a 

prepregnancy steady-state concentration. The model predicted 30 years as the exposure 

necessary to reach steady state for the general female population (1 E-4 to 2 E-3 μg/kg body 

weight [bw] /day). The model performance simulations for PFOA were run using an exposure of 

1.5 x 10-4 μg/kg bw/day. Projections were developed for maternal plasma, fetal plasma, infant 

plasma, and maternal milk. Agreement between the observed concentrations (μg/L) and the 

predicted values was considered satisfactory if the predicted value was within 1% of the 

Perfluorooctanoic acid (PFOA) – May 2016 2-43
observed value. Model output was compared to maternal and fetal plasma values at delivery or at 

specific time points, and for the infant plasma and milk data where available. Predicted 

maternal:fetal plasma (cord blood) concentration ratios were more consistent for PFOA than for 

PFOS when compared to the published data. The projections for fetal internal dose were 

reasonable, and there was good agreement between the model and the available human lactation 

data. The modeled maternal plasma was 2.4 μg/L at the time of conception; it slowly decreased 

across the gestation period and increased slightly at delivery. For the most part, the modeled 

results fell within ± 1 SD of the observed data. 

During lactation, there was a decline in maternal plasma across the 6 months of lactation (a 

change of 1μg/L). Thereafter, plasma values slowly increased and stabilized at about 1.5 μg/L at 

6 months postpartum. The fetal plasma concentration was about 2.3 μg/L at the start of gestation 

and declined to about 1.8 μg/L at the time of delivery. Maternal plasma values are about the 

same as those for the fetus. During the lactation period, the infant plasma increased in a linear 

fashion to a terminal value of about 5 μg/L. Milk concentrations declined very slightly across the 

lactation period with an initial concentration of 0.07 μg/L and a final value of 0.05 μg/L. These 

concentrations were estimated from the graphic data presentation. Breast milk appears to be an 

important excretory route for the dam. 

The projections for PFOA differed from those for PFOS in several respects. Most 

importantly, maternal and fetal plasma values were similar for PFOA but for PFOS, maternal 

levels were approximately twofold higher than fetal levels. Compared with PFOS, there was a 

much greater decline in maternal PFOA plasma values during lactation accompanied by a 

comparable decline in the PFOA concentration in milk. The increase in infant plasma across the 

lactation period was comparable for PFOA and PFOS, with the concentration at 6 months 

postpartum about 2.5 times higher than at 1 month. 

Loccisano et al. (2011, 2012a, 2012b, 2013) determined that the human pregnancy lactation 

model results, when compared to published data, identified the following important research 

needs:

• Are there differences in the transporter preferences and transfer rates for the individual 

PFASs? Do those differences correlate with half-life differences?

• Are there qualitative or quantitative differences between the transporters favored by 

PFOA and those favored by PFOS?

• What physiological factors influence CL for the mother, the fetus, and the infant during 

gestation and lactation?

• Are placental transport processes active, facilitated, or passive?

These research needs are more pronounced for PFOS than PFOA because the information 

supporting renal resorption and tissue uptake via membrane transporters for PFOS is very 

limited. Most models infer that PFOS and PFOA are similar based on their half-lives rather than 

on published research on PFOS transporter kinetics. 

Building on the work of other researchers, Wambaugh et al. (2013) developed and published 

a PK model to support the development of an EPA RfD for PFOA. The model was applied to 

data from studies conducted in monkeys, rats, and mice that demonstrated an assortment of 

systemic, developmental, reproductive, and immunological effects. A saturable renal resorption 

PK model was used. This concept has played a fundamental role in the design of all of the 

published PFOA models summarized in this section. In this case, an oral dosing version of the 

Perfluorooctanoic acid (PFOA) – May 2016 2-44
original model introduced by Andersen et al. (2006) and summarized early in section 2.6.1 was 

selected for having the fewest number of parameters that would need to be estimated. A unique 

feature of the Wambaugh et al. approach was to use a single model for all species in the 

toxicological studies to examine the consistency in the average serum values associated with 

effects and with no effects from nine animal studies of PFOA. The model structure is depicted in 

Figure 2-3, with minor modifications.

Wambaugh et al. (2013) placed bounds on the estimated values for some parameters of the 

Andersen et al. model (2006) to support the assumption that serum carries a significant portion of 

the total PFOA body load. The Andersen et al. model is a modified two-compartment model in 

which a primary compartment describes the serum and a secondary deep tissue compartment acts 

as a specified tissue reservoir. Wambaugh et al. (2013) constrained the total Vd to a value of not 

more than 100 times that in the serum. As a result, the ratio of the two volumes (serum versus 

total) was estimated in place of establishing a rate of transfer from the tissue to serum.

A nonhierarchical model for parameter values was assumed. Under this assumption, a single 

numeric value represents all individuals of the same species, gender, and strain. The gender 

assumption was applied to monkeys and mice while male and female rats were treated separately 

because of the established gender-related toxicokinetic differences. Body weight, number of 

doses, and magnitude of the doses were the only parameters to vary. In place of external doses, 

serum concentrations as measured at the time of euthanasia were used as the metric for dose 

magnitude. Measurement errors were assumed to be log-normally distributed. Table 2-25 

provides the estimated and assumed PK parameters applied in the Wambaugh et al. model (2013) 

for each of the species evaluated.

The PK data that supported the analysis were derived from two PFOA PK in vivo studies. 

The monkey PK data were derived from Butenhoff et al. (2004b), and the data for the rats (M/F) 

were from Kemper (2003). Two strains of female mice were analyzed separately, with CD1 

information derived from Lou et al. (2009) and C57Bl/6 information derived from DeWitt et al. 

(2008). The data were analyzed within a Bayesian framework using Markov Chain Monte Carlo 

sampler implemented as an R package developed by EPA to allow predictions across species, 

strains, and genders and to identify serum levels associated with the NOAEL and LOAEL 

external doses. The model chose vague, bounded prior distributions on the parameters being 

estimated, allowing them to be significantly informed by the data. The values were assumed to 

be log-normally distributed, constraining each parameter to a positive value.

The model predictions were evaluated by comparing each predicted final serum 

concentration to the serum value in the supporting animal studies. The predictions were generally 

similar to the experimental values. There were no systematic differences between the 

experimental data and the model predictions across species, strain, or gender, and median model 

outputs uniformly appeared to be biologically plausible despite the uncertainty reflected in some 

of the 95th percentile credible intervals. The application of the model outputs in deriving a human 

RfD is the focus of section 4.0 of this document. 

Perfluorooctanoic acid (PFOA) – May 2016 2-45
Table 2-25. Pharmacokinetic Parameters from Wambaugh et al. (2013) Meta-Analysis of 

Literature Data

Parameter Units CD1 Mouse (f)

a 

C57Bl/6 Mouse 

(f)

a 

Sprague-Dawley

Rat (f)

a 

Sprague-Dawley

Rat (m)

a 

Cynomolgus 

Monkey (m/f)

a 

bwb kg 0.02 0.02 0.20 (0.16 – 0.23) 0.24 (0.21 –0.28) 7 (m), 4.5 (f)

Cardiac 

Outputc

L/h/kg0.74 8.68 8.68 12.39 12.39 19.8

ka L/h 290 (0.6 –

73,000)

340 (0.53 –

69,000)

1.7 (1.1 – 3.1) 1.1 (0.83 – 1.3) 230 (0.27 –

73,000)

Vcc L/kg 0.18 (0.16 – 2.0) 0.17 (0.13 – 2.3) 0.14 (0.11 – 0.17) 0.15 (0.13 – 0.16) 0.4 (0.29 – 0.55)

k12 L/h 0.012 (3.1 x e-10 –

38,000)

0.35 (0.058 – 52) 0.098 (0.039 –

0.27)

0.028 (0.0096 –

0.08)

0.0011(2.4 x e-10

– 35,000)

RV2:V1 Unitless 1.07 (0.26 – 5.84) 53 (11 – 97) 9.2 (3.4 – 28) 8.4 (3.1 – 23) 0.98 (0.25 – 3.8)

Tmaxc μmol/h 4.91 (1.75 – 2.96) 2.7 (0.95 – 22) 1.1 (0.25 – 9.6) 190 (5.5 –

50,000)

3.9 (0.65 – 9,700)

KT μmol 0.037 (0.0057 –

0.17)

0.12 (0.033 –

0.24)

1.1 (0.27 – 4.5) 0.092 (3.4 x e-4 –

1.6)

0.043 (4.3 x e-5 –

0.29)

Free Unitless 0.011 (0.0026 –

0.051)

0.034 (0.014 –

0.17)

0.086 (0.031 –

0.23)

0.08 (0.03 – 0.22) 0.01 (0.0026 –

0.038)

Qfilc Unitless 0.077 (0.015 –

0.58)

0.017 (0.01 –

0.081)

0.039 (0.014 –

0.13)

0.22 (0.011 – 58) 0.15 (0.02 – 24)

Vfilc L/kg 0.00097 (3.34 x 

e-9 – 7.21)

7.6 x e-5 (2.7 x 

e-10 – 6.4)

2.6 x e-5 (2.9 x 

e-10 – 28)

0.0082 (1.3 x e-8

– 7.6)

0.0021 (3.3 x e-9

– 6.9)

Notes: 

Means and 95% confidence interval (in parentheses) from Bayesian analysis are reported. For some parameters, the distributions 

are quite wide, indicating uncertainty in that parameter (i.e., the predictions match the data equally well for a wide range of 

values). 

m = male, f = female

a Data sets modeled for the CD1 mouse were from Lou et al. (2009), for the C57Bl/6 mouse were from DeWitt et al. (2008), for 

the rat were from Kemper (2003), and for the monkey from Butenhoff et al. (2004b). b Estimated average body weight for species used except with Kemper study (2003) where individual rat weights were available 

and assumed to be constant. c Cardiac outputs obtained from Davies and Morris (1993).

2.6.2 Half-Life Data

Human. There have been several studies of half-lives in humans and all support a long residence 

time for serum PFOA with estimates measured in years rather than months or weeks. Bartell et 

al. (2010) determined an average half-life of 2.3 years based on a study of the decreases in 

human serum levels after treatment of drinking water for PFOA removal was instituted by the 

Lubeck Public Services District in Washington, West Virginia, and the Little Hocking Water 

Association (LHWA) in Ohio. Source waters for these systems had become contaminated with 

PFAS from the DuPont Works Plant in Washington, West Virginia, between 1951 and 2000. 

The Bartell et al. study (2010) was based on a series of serum measurements (eight over 

4 years) from 200 individuals who agreed to participate in the study. Inclusion criteria for the 

participants included: serum PFOA concentrations ≥ 50 ng/mL, residential water service 

provided by one of the two treatment plants, never employed at the DuPont plant, not growing 

their own vegetables, and signed acceptance of the study consent form. The participants were 

almost equally divided between males and females with an average age of about 50 years (range 

of 18–89 years). Most of the participants consumed public tap water (172) as their primary 

source, but a small number (28) consumed bottled water as their source. 

Perfluorooctanoic acid (PFOA) – May 2016 2-46
The participants were required to report that they primarily used home tap water for cooking, 

bathing, and showering for the years between 2005 and 2007. The tap water users had to report 

public water as their primary source of residential water consumption, and bottled water users 

had to report the use of bottled water as their primary source of residential water consumption. 

The initial blood draw for serum occurred in June 2007, with subsequent samples at 1, 2, 3, 6, 

and 12 months after the initial sample. Samples were analyzed by the Centers for Disease 

Control and Prevention. Nineteen samples from the 2-month blood draw were not analyzed due 

to mislabeling. 

A linear mixed model was used to determine the decline in serum PFOA concentration over 

time. With these models, the decline from baseline by the participants was essentially first order. 

The serum PFOA concentration was the only time-varying measurement entered into the model. 

Serum concentrations were log-normally distributed, as described by the following equation:

lnC = lnC0-kt

where: 

C= serum concentration at time t

C0 = baseline serum concentration

k = elimination rate constant 

t = time point for the measurement

The results of this assessment showed a 26% decrease in PFOA concentration per year after 

adjustment for covariates and a half-life of 2.3 years [confidence interval (CI) = 2.1-2.4]. The 

covariates considered included the water treatment system, the time exposed before and after 

filtration, public versus bottled water, gender, age, consumption of local or homegrown 

vegetables, and exposure to the public water supply at work. The only potential confounders 

determined to be significant were the treatment plant (p = 0.03) and homegrown vegetable 

consumption (p<0.001). 

Identification of consumption of homegrown vegetables as a significant confounder revealed 

a weakness in the study design because it had been an exclusion factor, yet was identified as an 

exposure source at the 12-month interview of the study participants. The researchers concluded 

that this problem was a result of the way the exclusion question was phrased for the original 

interview, “Do you grow your own vegetables?” When the question was asked later in the study, 

it was rephrased, “Do you eat any fruits and vegetables grown at your own home?” Some people 

who answered “no” to the original questions answered “yes” to the second question. 

Changes in the source of drinking water during the study could also have impacted the 

results. When baseline interview data were compared with the results from the 12-month 

interview, 39% of the bottled water group reported using public water at home. Some of the 

public water drinkers (10%) reported using primarily bottled water at the 6-month interview. 

In another study, the drinking water supply was contaminated with a mixture of 

perfluorinated chemicals when a soil-improver mixed with industrial waste was applied upriver 

to agricultural lands in Arnsberg, Germany (Brede et al. 2010). The PFOA levels in the finished 

drinking water were measured as 500–640 ng/L in 2006. PFOS and PFHxS also were present. 

The plasma PFOA levels in the Arnsberg population were 4.5 to 8.3 times higher than those in a 

reference community at the time the problem was discovered. Charcoal filtration was added to 

the potable treatment train and succeeded in reducing concentrations in the drinking water. 

Perfluorooctanoic acid (PFOA) – May 2016 2-47
The authors used the differences in plasma 2008 PFOA measurements from a subset of the 

participants (children and adults) initially exposed in 2006 to determine the PFOA half-life. The 

2008 subjects included 66 males, females, and children from Arnsberg and 73 from the reference 

community in the evaluation. The drinking water concentration monitoring results (nondetects 

estimated as one-half of the limit of detection [LOD], 10 ng/L) and DWI estimates obtained by 

questionnaire and interview were used to estimate PFOA exposures. Plasma PFOA samples were 

collected during a 2-month period in late 2008. Plasma PFOA had declined in the serum for both 

the Arnsberg residents (39.2%) and those from the reference community (13.4%). In Arnsberg, 

the decrease was greater for the exposed females and children than for the males when compared 

to the reference community, an observation that appeared to reflect the reported lower DWIs of 

the Arnsberg females and children (0.3 ± 0.2 and 0.8 ± 0.6 L/day compared to 0.7 ± 0.5 and 

1.6 ± 0.8 L/day, respectively). The estimate for the human half-life was 3.26 years (geometric 

mean; range 1.03–14.67 years). Regression analysis of the data also suggested that the 

elimination rate might have been greater in younger subjects and older subjects. 

Seals et al. (2011) determined half-life estimates for 602 residents of Little Hocking, Ohio, 

and 971 residents of Lubeck, West Virginia, who were part of the C8 study but had relocated to a 

different area of the country. The half-life estimate was based on the decline in serum PFOA 

levels after the time of the initial measurement and the years since the change in residential 

location occurred. A background estimate (5 ng/mL) was subtracted from the serum 

measurements before analysis. On average, the initial serum PFOA concentrations were higher in 

Little Hocking (60.6 ng/mL) than in Lubeck (31.0 ng/mL). Due to the nonlinearity in scatter 

plots of the natural log for adjusted serum PFOA concentrations versus the years elapsed since 

relocation, the authors used a two-segment linear spline regression approach in their analysis of 

the data (i.e., Little Hocking—4 years, Lubeck—9 years). The slope of the line decreased for the 

second time segment compared to the first. In former residents of Little Hocking, a -21.4%-

change in serum PFOA was observed in the first 4 years after leaving Little Hocking, and 

a -7.6%-change was observed beyond 4 years. In former Lubeck residents, the serum PFOA 

change was -7.8% for the first 9 years and 0.2% (a slight increase) afterwards. The half-life 

estimates for Little Hocking ranged from 2.5–3.0 years (average 2.9 years) and for Lubeck 

ranged from 5.9–10.3 years (average 8.5 years).

Based on their analysis, the authors suggested that, if their assumptions were correct, a 

simple first order elimination model might not be appropriate for PFOA given that the rate of 

elimination appeared to be influenced by both concentration and time. There was a difference in 

the CL for the two locations even though the range of years elapsed since relocation was the 

same for both communities. The authors identified three potential limitations of their analysis:

the cross-sectional design, the assumption that exposure was uniform within a water district, and 

a potential bias introduced by exclusion of individuals with serum values <15 ng/mL. 

3M (Burris et al. 2000, 2002) conducted a half-life study on 26 retired fluorochemical 

production workers from their Decatur, Alabama, (n = 24) and Cottage Grove, Minnesota, 

(n = 3) plants. Blood was collected from the subjects between 1998 and 2004, a period during 

which serum samples were drawn every 6 months over a 5-year period, depending on the facility 

at which the subject had worked. Responses on questionnaires determined whether any of the 

retirees had occupational exposures after retirement. The average number of years that 

participants worked was 31 (range 20–36 years) and they had been retired an average of 

2.6 years at study initiation (range 0.4–11.5 years). The mean age of the retirees was 61 years 

(range 55–75) at the beginning of the study. 

Perfluorooctanoic acid (PFOA) – May 2016 2-48
The initial mean serum PFOA concentration of all of the subjects was 0.691 μg/ml (range 

0.072–5.1 μg/mL). At the completion of the study, the mean PFOA concentration was 

0.262 μg/mL (range 0.017–2.435 μg/mL). Two of the retirees died during the study period; 

therefore, they were only followed for 4.2 years. The mean serum elimination half-life of PFOA 

in these workers was 3.8 years (1378 days, 95% CI, 1131-1624 days) and the median was 

3.5 years (Olsen et al. 2005). The range was 1.5–9.1 years (561–3334 days). No association was 

reported between the serum elimination half-life and with initial PFOA concentrations, age, or 

gender of the retirees, the number of years retired or working at the production facility, or 

medication use or health conditions. 

Harada et al. (2005) studied the relationship between age, gender, and serum PFOA 

concentration in residents of Kyoto, Japan. They found that females in the 20–50-year-old age 

group (all with regular menstrual cycles) had serum PFOA concentrations that were significantly 

lower than those in females over age 50 (all postmenopausal). Mean serum PFOA concentration 

in the younger females was 7.89 ± 3.61 ng/ml versus 12.63 ± 2.42 ng/mL in the older females. 

This age difference in serum PFOA concentrations was not seen in males, and serum PFOA 

concentrations in males were comparable to those of the older females.

Harada et al. (2005) also estimated the CLR rate of PFOA in humans and found it to be only 

about 0.001% of the GFR. There was no significant difference in CLR of PFOA with respect to 

gender or age group, and the mean value was 0.03 ± 0.013 ml/day/kg. 

Animal. Kemper (2003) examined the plasma concentration profile of PFOA following gavage 

administration in sexually mature Sprague-Dawley rats. Male and female rats (four per gender 

per group) were administered single doses of PFOA by gavage at DRs of 0.1, 1, 5, and 25 mg 

PFOA/kg. After dosing, plasma was collected for 22 days in males and 5 days in females. 

Plasma concentration versus time data were then analyzed using noncompartmental PK methods 

(see Table 2-26 and Table 2-27). To further characterize plasma elimination kinetics, animals 

were given oral PFOA at a rate of 0.1 mg/kg, and plasma samples were collected until PFOA 

concentrations fell below quantitation limits (extended time).

Plasma elimination curves were linear with respect to time in male rats at all dose levels. In 

males, plasma elimination half-lives were independent of dose level and ranged from 

approximately 138 hours to 202 hours. To further characterize plasma elimination kinetics, 

particularly in male rats, animals were given oral PFOA at a dose of 0.1 mg/kg, and plasma 

samples were collected until PFOA concentrations fell below quantitation limits (2,016 hours in 

males). The estimated plasma elimination half-life in this experiment was approximately 

277 hours (11.5 days) in male rats.

Plasma elimination curves were biphasic in females at the 5-mg/kg and 25-mg/kg dose 

levels. In females, terminal elimination half-lives ranged from approximately 2.8 hours at the 

lowest dose to approximately 16 hours at the high dose. The estimated plasma elimination halflife in the extended time experiment was approximately 3.4 hours in females. 

Perfluorooctanoic acid (PFOA) – May 2016 2-49
Table 2-26. PK Parameters in Male Rats Following Administration of PFOA

Parameter

Dose

0.1 mg/kg 1 mg/kg 5 mg/kg 25 mg/kg 1 mg/kg (IV)

0.1 mg/kg

extended 

time

Tmax (hr) 10.25

(6.45)

9.00

(3.83)

15.0

(10.5)

7.5

(6.2)

NA 5.5

(7.0)

Cmax (μg/mL) 0.598

(0.127)

8.431

(1.161)

44.75

(6.14)

160.0

(12.0)

NA 1.08

(0.42)

Lambda z (1/hr) 0.004

(0.001)

0.005

(0.001)

0.0041

(0.0007)

0.0046

(0.0012)

0.004

(0.000)

0.0026

(0.0007)

T1/2 (hr) 201.774

(37.489)

138.343

(31.972)

174.19

(28.92)

157.47

(38.39)

185.584

(19.558)

277.10

(56.62)

AUCINF (hr.μg/mL) 123.224

(35.476)

1194.463

(215.578)

6733.70

(1392.83)

25,155.61

(7276.96)

1249.817

(113.167)

206.38

(59.03)

AUCINF/D

(hr.μg/ml/mg/kg)

1096.811

(310.491)

1176.009

(206.316)

1221.89

(250.28)

942.65

(284.67)

1123.384

(100.488)

2111.28

(586.77)

Clp (mL/kg.hr) 0.962

(0.240)

0.871

(0.158)

0.85

(0.21)

1.13

(0.31)

0.896

(0.082)

0.51

(0.17)

Source: Kemper 2003

Notes:

Mean (SD)

AUCINF: area under the plasma concentration time curve, extrapolated to infinity; AUCINF/D: AUCINF normalized to dose; Clp: 

plasma clearance; Cmax: maximum plasma concentration; Lambda z: terminal elimination constant; T1/2: terminal elimination 

half-life; Tmax: time to Cmax. 

Table 2-27. PK Parameters in Female Rats Following Administration of PFOA

Parameter

Dose

0.1 mg/kg 1 mg/kg 5 mg/kg 25 mg/kg

1 mg/kg

(IV)

0.1 mg/kg

extended 

time

Tmax (hr) 0.56

(0.31)

1.13

(0.63)

1.50

(0.58)

1.25

(0.87) NA 1.25(0.50)

Cmax (μg/mL) 0.67

(0.07)

4.782

(1.149)

20.36

(1.58)

132.6

(46.0) NA 0.52(0.08)

Lambda z (1/hr) 0.231

(0.066)

0.213

(0.053)

0.15

(0.02)

0.059

(0.037)

0.250

(0.047)

0.22

(0.07)

T1/2 (hr) 3.206

(0.905)

3.457

(1.111)

4.60

(0.64)

16.22

(9.90)

2.844

(0.514)

3.44

(1.26)

AUCINF (hr.μg/mL) 3.584

(0.666)

39.072

(10.172)

114.90

(11.23)

795.76

(187.51)

33.998

(7.601)

3.34

(0.32)

AUCINF/D

(hr.μg/mL/mg/kg)

31.721

(5.880)

38.635

(10.093)

20.78

(2.01)

29.54

(6.92)

30.747

(6.759)

34.39

(3.29)

Clp (mL/kg.hr) 32.359

(6.025)

27.286

(7.159)

48.48

(4.86)

35.06

(.88)

34.040

(9.230)

29.30

(3.06)

Source: Kemper 2003 

Note: Mean (SD)

Gibson and Johnson (1979) administered a single dose of 14C-PFOA averaging 11.4 mg/kg 

by gavage to groups of three male 10-week-old CD rats. The elimination half-life of 14C from the 

plasma was 4.8 days. NRC ([2005], cited in Butenhoff et al. [2004b]) reported half-lives of 4–6 

days for male rats and 2–4 hours for female rats; there was no mention of the strains studied. 

Perfluorooctanoic acid (PFOA) – May 2016 2-50
Kemper (2003) reported half-lives of 6–8 days for male Sprague-Dawley rats (Table 2-26) and 

3–16 hours for females (Table 2-27).

Lou et al. (2009) determined values of 21.7 days (95% confidence interval: 19.5–24.1) for 

male CD1 mice and 15.6 days (95% confidence interval: 14.7–16.5) for females for use in their 

pharmacokinetic model (see section 2.6.1). NRC ([2005], cited in Butenhoff et al. [2004b]) 

provided values of 12 days for males and 20 days for females without any information on strains. 

Butenhoff et al. (2004b) looked at the elimination half-life in monkeys treated for 6 months 

with 0, 3, 10, and 20 mg/kg/day via capsules. Elimination of PFOA from serum after cessation of 

dosing was monitored in recovery monkeys from the 10- and 20-mg/kg dose groups. For the two 

monkeys exposed to 10 mg/kg, serum PFOA elimination half-life was 19.5 (R2=0.98) days and 

indicated first-order elimination kinetics. For three monkeys exposed to 20 mg/kg, serum PFOA 

elimination half-life was 20.8 days (R2

=0.82) and also indicated first-order elimination kinetics, 

although dosing was suspended at different time points because of weight loss. The data from

NRC (2005), which were provided by Butenhoff et al. (2004b), were about 21 days for females 

and 30 days for males. 

2.6.3 Volume of Distribution Data

Several researchers have attempted to characterize PFOA exposure and intake in humans 

through PK modeling (Lorber and Egeghy 2011; Thompson et al. 2010). As an integral part of 

model validation, the parameter for Vd of PFOA within the body was calibrated from the 

available data. In the models discussed below, Vd was defined as the total amount of PFOA in 

the body divided by the blood or serum concentration. 

Two groups of researchers defined a Vd of 170 ml/kg body weight for humans for use in a 

simple, single compartment, first-order PK model (Lorber and Egeghy 2011; Thompson et al. 

2010). The models developed by these groups were designed to estimate intakes of PFOA by 

young children and adults (Lorber and Egeghy 2011) and the general population of urban areas 

on the east coast of Australia (Thompson et al. 2010). In both models, the Vd was calibrated

using human serum concentration and exposure data from the NHANES and assumes that most 

PFOA intake is from contaminated drinking water. Thus, in using the models to derive an intake 

from contaminated water, the Vd was calibrated so that model prediction of elevated blood levels 

of PFOA matched those seen in residents.

Butenhoff et al. (2004b) calculated a Vd from noncompartmental PK analysis of data from 

cynomolgus monkeys. Three males and three females were administered a single IV dose of 

10 mg/kg, and serum PFOA concentrations were measured in samples collected up to 123 days 

post-dosing. The Vd of PFOA at steady state (Vdss) were similar for both genders at 181 ± 12 

ml/kg for males and 198 ± 69 ml/kg for females. 

2.6.4 Toxicokinetic Summary

Uptake and egress of PFOA from cells is largely regulated by transporters in cell membranes. 

It is absorbed from the gastrointestinal tract as indicated by serum measurements in humans and 

treated animals. In serum, PFOA is electrostatically bound to albumin occupying up to nine to 

twelve sites and sometimes displacing other substances that normally would occupy a site. 

Linear PFOA chains display stronger binding than branched chains. PFOA binding causes a 

Perfluorooctanoic acid (PFOA) – May 2016 2-51
change in the conformation of serum albumin, altering its ability to bind with some endogenous 

and exogenous materials it normally transports. 

PFOA is distributed to tissues by a process requiring membrane transporters. Accordingly, 

the tissue levels vary from organ to organ. The highest tissue concentrations are usually in the 

liver. Liver accumulation in males is greater than in females. Other tissues with a tendency to 

accumulate PFOA are the kidneys, lungs, heart, and muscle, plus the testes in males and uterus in 

females. Post-mortem studies in humans have found PFOA in liver, lungs, bone, and kidneys, 

but only low levels in brain. PFOA is not metabolized, thus, any effects observed in toxicological 

studies are the result of parent compound, not metabolites. 

Electrostatic interactions with proteins are an important toxicokinetic feature of PFOA. 

Studies demonstrate binding or interactions with membrane receptors (e.g., PPARα, T3), 

transport proteins, and enzymes. Saturable renal resorption of PFOA from the glomerular filtrate 

via transporters in the kidney tubules is a major contributor to the long half-life of this 

compound. Branched-chain PFOAs are less likely to be resorbed than the linear molecules based 

on half-life information in humans. All toxicokinetic models for PFOA are built on the concept 

of saturable renal resorption first proposed by Andersen et al. (2006). Some PFOA is removed 

from the body with bile, a process that also is transporter-dependent. Accordingly, the levels in 

fecal matter represent both unabsorbed material and that discharged with bile.

During pregnancy, PFOA is present in the placenta and amniotic fluid in both animals and 

humans. Post-delivery, PFOA is transferred to offspring through lactation in a dose-related 

manner. Maternal serum levels decline as those in the pups increase. This also occurs in humans 

as demonstrated in a study of females breast-feeding their infants in Little Hocking, Ohio. 

The half-life in humans for occupationally exposed workers was 3.8 years (95% CI, 1.5-9.1). 

The average half-life was 2.3 years among people in the Lubeck Public Services District in West 

Virginia and the LHWA in Ohio, based on changes in serum levels for the general population 

after treatment of drinking water was implemented. This half-life value reflects humans whose 

exposure came primarily from their public water system. Half-lives from animals included 

21 days (females) and 30 days (males) for monkeys Butenhoff et al. (2004b); 11.5 days (males) 

and 3.4 hours (females) in Sprague-Dawley rats (Kemper 2003); and 27.1 days (male) and 

15.6 days (female) CD1 mice (Lau et al. 2006). The gender difference between male and female 

rats is not seen in mice. In early life, the half-lives are nearly the same for both genders of rats, 

but once the animals reach sexual maturity, resorption increases in males, prolonging the halflife (Hinderliter 2004; Hundley et al. 2006). This change appears to be under the control of 

hormones in both males and females (Cheng et al. 2006; Kudo et al. 2002).

Perfluorooctanoic acid (PFOA) – May 2016 2-52
3 HAZARD IDENTIFICATION

This section provides a summary and synthesis of the data from a large number of human 

epidemiology studies accompanied by studies in laboratory animals designed to identify both the 

dose response and critical effects that result from exposures to PFOA and to examine the MoA 

leading to toxicity.

3.1 Human Studies

Epidemiology studies of effects of PFOA have been conducted in three types of populations: 

workers exposed in chemical plants producing or using PFOA, high-exposure communities 

(i.e., an area in West Virginia and Ohio that experienced water contamination over a period of 

more than 20 years), and general population studies with background exposures. These 

populations differ with respect to exposure levels. The approximate range in serum PFOA 

concentrations is 0.010–> 2.0 (means around 1–4 μg/mL) in the PFOA-exposed workers, and 

0.010–0.100 μg/mL and below LOD to < 0.010 μg/mL in the high-exposure community and 

general population settings, respectively. Although moderate-to-high correlations between PFOA 

and PFOS are often seen in general populations (r > 0.5), the correlation is lower in the West 

Virginia and Ohio high-exposure area (r=0.3). In evaluating and synthesizing results from these 

studies, it is important to consider differences in the exposure range within the study population 

and the exposure level within the referent group, as differences (or inconsistencies) can be 

expected depending on the shape of the exposure-response curve and the exposure range 

encompassed by different studies. In addition, the optimal choice of an exposure metric 

(e.g., cumulative or a time-specific) depends on the specific outcome being examined. 

Occupational studies. Large-scale production of PFOA occurred in the United States for several 

decades. Both 3M (in Alabama and Minnesota) and DuPont (in West Virginia) have been the

primary U.S. producers and users of perfluorinated compounds, and both companies have 

offered voluntary fluorochemical medical surveillance programs to workers at plants that 

produced or used perfluorinated compounds. The monitoring data collected by 3M and DuPont 

were used in conjunction with mortality and health effects information in a number of 

epidemiology studies of cancer and noncancer outcomes in the worker populations. 3M 

discontinued manufacturing PFOA in 2000, but a subsidiary in Europe (Antwerp, Belgium) 

continued to manufacture and sell it through 2008. 

High-exposure community studies. Members of the general population living in the vicinity of 

the DuPont Washington Works PFOA production plant in Parkersburg, West Virginia, are the 

focus of a large-scale, community-based study titled the C8 Health Project. Releases from the 

Washington Works plant, where PFOA (C8) was used as a processing aid in the manufacture of 

fluoropolymers, contaminated the ground water of six water districts near the plant resulting in 

exposures to the general population. The plant began production in the 1950s, with PFOA use 

and emission from the plant increasing in the 1980s. Study participants from the affected areas 

(n = 69,030; 33,242 males, 35,788 females; <10–70+ years) were identified in 2005–2006, and a 

series of studies were conducted. The participants all received compensation and provided a 

blood sample and filled out an extensive questionnaire that included information on drinking 

water sources, use of home-grown produce, and health information. A variety of approaches to 

exposure assessment have been used in these studies, with the most detailed incorporating 

individual residential history and water consumption and source data, emissions data, 

Perfluorooctanoic acid (PFOA) – May 2016 3-1 
environmental characteristics, water pipe installation, PK data, and workplace exposures (Barry 

et al. 2013; Shin et al. 2011; Vieira et al. 2013). The methods allow the estimation of cumulative 

and of current exposure at different time periods or ages for individual study participants. Details 

of the specific analyses undertaken to estimate historical exposures and to ascertain different 

types of outcomes (retrospective and prospective analyses) are described in detail below. 

Drinking water concentrations were based on PFOA releases from the DuPont plant and 

residential address history of the participants (C8 Science Panel 2012).

The C8 Health Project also involved a review of evidence of health effects, considering their 

own studies and studies conducted by others and in other populations. The conclusions for each 

health endpoint assessed—“probable link” or “not a probable link”—are available on the C8 

Science Panel website in a series of reports completed in 2011–2012 (see 

http://www.c8sciencepanel.org/index.html). 

General population studies. Studies investigating the association between PFOA levels and 

health effects in the U.S. general population have been conducted using the NHANES data set. 

NHANES examined representative members of the U.S. population (~5000 adults and 

children/year) through surveys focusing on different health topics. The study consists of an 

interview (demographic, socioeconomic, dietary, and medical questions) and examination 

(medical including blood and urine collection, and dental and physiological parameters). 

Biomonitoring included a number of PFAS, predominantly PFOA and PFOS.

A study by Jain (2014) examined the influence of diet and other factors on the levels of 

serum PFOA and other PFAS using NHANES 2003–2004, 2005–2006, and 2007–2008 data. 

Significantly higher serum PFOA levels were found in males (0.0047 μg/mL) than in females 

(0.035 μg/mL) and in smokers (0.043 μg/mL) than in nonsmokers (0.040 μg/mL). No significant 

differences in PFOA serum concentration were seen during the time periods evaluated. There 

was a positive association of PFOA with increases in serum cholesterol (p<0.001), serum 

albumin (p<0.001), and body mass index (BMI) (p<0.04) based on the 5,591 records used in the 

assessment. Intakes of nonalcoholic beverages were positively associated with serum PFOA 

(P<0.001), but no associations were found for other dietary food groupings. 

The results of these studies along with other population studies are described in the following 

sections. In the studies of worker cohorts, the data collected focused on measures of 

cardiovascular risk, signs of organ damage, standard hematological endpoints, and cancer 

(primarily cancer-related mortality). Within the general population, data were focused on 

cardiovascular risk factors and diabetic or prediabetic conditions as well as reproductive and 

developmental endpoints. The following summary focuses on measures of lipids (e.g., 

cholesterol, LDL); liver, kidney, and thyroid effects; reproductive effects (e.g., pregnancy-related 

outcomes, specifically pregnancy-related hypertension and preeclampsia, measures of fetal 

growth, and pubertal development); and cancer (specifically kidney and testicular cancer). These 

outcomes were selected either because of the availability of studies in a variety of settings with 

some indications of effects (e.g., as noted in the C8 Science Panel reports), or to allow 

comparison with results from studies in animals. Summary tables are included to support 

evaluation of the weight of evidence and facilitate comparison of the serum concentrations in the 

epidemiology studies to those in the animals studies summarized in section 3.2. 

Perfluorooctanoic acid (PFOA) – May 2016 3-2 
3.1.1 Noncancer

3.1.1.1 Serum Lipids and Cardiovascular Diseases

Serum Lipids 

Occupational studies. Four cross-sectional studies are described in this section and in Table 3-1. 

Olsen et al. (2000) analyzed data from voluntary medical surveillance examinations of PFOA 

production workers at a 3M plant in 1993, 1995, and 1997. Cholesterol, LDL, HDL, and 

triglycerides were measured in male workers (n = 111 in 1993, n = 80 in 1995, and n = 74 in 

1997). Multivariable regression analyses, conducted separately by year (cross-sectional), were 

adjusted for age, BMI, alcohol consumption, and cigarette use. Employees’ serum PFOA levels 

were stratified into three categories—<1, 1- <10, and ≥10 μg/mL. The sample size in the highest 

category ranged from 11 to 15 in the three examination years. There was little variation by 

exposure category in mean or median TC, LDL, HDL, or triglycerides across the workers in 

1993, 1995, or 1997. 

Olsen and Zobel (2007) examined data from the 2000 medical surveillance program at the 

three 3M plants, which is an expanded and refined analysis of the data reported in Olsen et al. 

(2003). The fluorochemical workers consisted of males (age 21–67) from the Antwerp, Belgium 

(n = 196); Cottage Grove, Minnesota (n = 122); and Decatur, Alabama (n = 188) production 

facilities who volunteered to participate in the medical surveillance program and did not take 

cholesterol-lowering medication. Blood was collected for fluorochemical concentration 

determination and serum lipid parameters including cholesterol, LDL, HDL, and triglycerides. 

Analysis of variance (ANOVA), analysis of covariance, logistic regression, and multiple 

regression models were used to analyze the data with age, BMI, and alcohol consumption as 

covariates. Potential associations with PFOS levels were not evaluated because a previous 

analysis had shown no association between PFOS and the selected outcomes. Serum PFOA 

concentrations ranged from 0.01 to 92.03 μg/mL for the male workers (all sites combined), with 

a mean serum PFOA concentration of 2.21, 1.02, 4.63, and 1.89 μg/mL for all sites combined, 

and the Antwerp, Cottage Grove, and Decatur sites, respectively. Serum PFOA (all sites 

combined) was not associated with TC or LDL. A negative association was observed between 

serum PFOA concentration (all sites combined) and HDL. Serum triglyceride was positively 

associated with serum PFOA at all sites combined and independently at the Antwerp site. 

Nonadherence to the fasting requirement for blood collection, especially for night-shift workers, 

and potential binding of PFOA to albumin and LDL, were identified by the authors as possible 

factors that influenced the triglyceride results. 

Sakr et al. (2007a) conducted a cross-sectional analysis of PFOA and lipids among active 

employees at the DuPont Washington Works fluoropolymer production plant in West Virginia. 

The employees who volunteered to participate in the study (n = 1025, 782 males, 243 females) 

each had a physical examination, provided a fasting blood sample, and answered a medical and 

occupation history questionnaire in 2004. The association between PFOA and lipid levels was 

evaluated by ANOVA, χ2 test, student’s t-test, and linear regression models. Confounders 

including age, BMI, gender, alcohol consumption, and parental heart attack were considered in 

the models. Mean serum PFOA concentration in the workers was 0.428 ± 0.189 μg/mL 

(interquartile range 0.099–0.381). For those with current occupational exposure to PFOA, the 

range was 0.0174–9.55 μg/mL and for workers with intermittent occupational exposure, the 

range was 0.0081–2.07 μg/mL. The range was 0.0086–2.59 μg/mL for workers with past 

occupational exposure and the 0.0046–0.963 μg/mL for workers with no occupational exposure. 

Perfluorooctanoic acid (PFOA) – May 2016 3-3 
Serum PFOA was positively associated with cholesterol, very low-density lipoprotein (VLDL), 

and LDL (p<0.03) in the participating workers, whether or not they were taking lipid-lowering 

medication. No association was observed between serum PFOA and HDL or triglycerides. PFOS 

was not included in the study. 

Costa et al. (2009) examined serum lipid data using 30 years of medical surveillance data 

from workers of a PFOA production plant in Italy. The workers (n = 53 males, 20–63 years of 

age) participated in the medical surveillance program yearly from 1978 to 2007. The length of 

work exposure was 0.5–32.5 years. In 2007, 37 males were active workers and 16 males were 

retired or had transferred to other departments and were no longer being exposed. Unexposed 

male workers (n = 107, 12 executives and 95 blue collar workers) from different departments 

also participated in the medical surveillance program and served as controls. Beginning in 2000, 

serum PFOA was monitored yearly except in 2005. Serum PFOA concentrations in the workers 

decreased after plant renovations partially automated the PFOA production process and 

procedures for the use of protective devices were instituted in 2002. In 2007, the geometric mean 

serum PFOA was 4.02 and 3.76 μg/mL, respectively, in currently exposed and retired workers. 

Three analyses were conducted: a t-test comparing 34 exposed workers matched to 34 unexposed 

workers by age, work seniority, day/shift work, and living conditions; linear regression with 34 

exposed workers and 107 unexposed workers adjusting for age, work seniority, BMI, smoking, 

and alcohol consumption; and a repeated measures analysis with a total of 56 individuals with 

more than one measure, adjusting for age, work seniority, BMI, smoking, alcohol consumption, 

and year of observation. TC and uric acid were significantly increased (p<0.05) in relation to 

PFOA exposure in each of these analyses. No correlations were observed between serum PFOA 

concentration and Apo-A (HDL-associated) or Apo-B (LDL-associated) proteins, HDL, or 

triglycerides in any of the analyses. PFOS was not included in this study. 

Three other studies included analyses with multiple measures over time (Table 3-1). Olsen et 

al. (2003) conducted a longitudinal analysis of the 2000 medical surveillance data from the 3M 

workers in conjunction with 1995 and 1997 data. This analysis included 175 male employees 

with data from 2000 and at least one of the other survey dates. Only 41 employees were 

participants in all three surveillance periods. Mean serum levels for the group sampled in 1995 

and 2000 (n = 64) were 1.36 μg/mL and 1.59 μg/mL, respectively. Mean serum levels for the 

group sampled in 1997 and 2000 (n = 69) were 1.22 μg/mL and 1.49 μg/mL, respectively. 

Finally, mean serum levels for the group sampled in 1995, 1997, and 2000 (n = 41) were 

1.41 μg/mL, 1.90 μg/mL, and 1.77 μg/mL, respectively. When serum levels were analyzed using 

a repeated measures mixed-model multivariable regression, adjusting for age, BMI, smoking, 

alcohol consumption, location, year at first entry, years worked (at baseline), and years of 

follow-up, there was a statistically significant positive association between PFOA and serum 

cholesterol (Beta1

 = 0.032, 95% CI 0.013, 0.015) and triglycerides (Beta = 0.094, 95% CI 0.045, 

0.144) (p = 0.0002). PFOS levels were not associated with changes in serum lipids over time. 

Sakr et al. (2007b) conducted a longitudinal analysis among the workers at the DuPont 

Washington Works plant in West Virginia using data from 1979 to 2004. Employee medical 

records from the medical surveillance program were used to obtain blood lipid (e.g., TC, LDL, 

HDL, triglycerides), height, and weight data. As part of the medical surveillance program, each 

employee gave a detailed medical history and had a physical examination at least every 3 years. 

Serum PFOA concentration was measured every 1–2 years in PFOA-exposed workers and every 

1

 The beta coefficient from the regression analysis.

Perfluorooctanoic acid (PFOA) – May 2016 3-4 

 
3–5 years in non-PFOA-exposed workers on a volunteer basis. This study included 454 workers 

who had two or more serum PFOA measurements. The study population included 334 males and 

120 females ranging in age from 24 to 66 years who had worked at the plant for at least 1 year 

since 1979. A linear mixed effects regression model was used to analyze the data and accounted 

for age (and age-squared), gender, BMI, and decade of hire as potential confounders. Serum PFOA 

concentrations ranged from 0 to 22.66 μg/mL, with a mean of 1.13 μg/mL over the 23-year 

monitoring period in the study population. For employees with two or more PFOA measurements, 

the mean of the first and last sample was 1.04 μg/mL and 1.16 μg/mL, respectively, with an 

average of 10.8 years between samples. Serum PFOA concentration was positively associated with 

TC after age, BMI, gender, and decade of hire adjustment in the model (Beta = 1.06, 95% CI 0.24, 

1.88) per ppm increase in PFOA. Information on lipid-lowering medications and alcohol intake by 

the participants was not available. PFOS was not included in this study.

Steenland et al. (2015) conducted an analysis of the incidence of several conditions, 

including high cholesterol (based on prescription medication use) among 3,713 workers at the 

Washington Works plant in West Virginia who participated in the C8 Health Project. Yearly 

serum estimates were modeled from work history information and job-specific concentrations. 

Cox proportional hazard models, stratified by birth year, were used to assess self-reported 

incidence of high cholesterol in relation to time-varying cumulative estimated PFOA serum 

concentration, controlling for gender, race, education, smoking, and alcohol consumption. No 

association was seen when analyzed without a lag (HRs by quartile 1.0, 1.11, 1.06, 1.05; trend 

p = 0.56 for log cumulative exposure), or when using a 10-year lag (HRs by quartile 1.0, 0.93, 

1.01, 0.96; trend p = 0.62). 

High-exposure community studies. Several studies examined serum lipids in populations 

serviced by water districts contaminated by the Washington Works PFOA production plant in 

Ohio and West Virginia (Table 3-2). Emmett et al. (2006) is a small study (n = 371) with limited 

analysis (t-tests comparing PFOA levels in people with abnormal versus normal TC); the larger 

studies were conducted as part of the C8 Health Project. This collection of studies includes 

analyses of current serum PFOA levels in relation to serum lipids in adults (Steenland et al. 

2009) and children (Frisbee et al. 2010), longitudinal analysis of the change in lipids seen in 

relation to a change in serum PFOA (Fitz-Simon et al. 2013), and analyses of the incidence of 

hypercholesteremia in relation to modeled exposure (Winquist and Steenland 2014a). With the 

exception of one set of analyses within the Winquist and Steenland study (2014a), these data 

provide consistent evidence of positive associations between PFOA exposure (measured directly 

in blood or modeled based on environmental and drinking water data) and TC. 

Emmett et al. (2006) examined the association of serum PFOA concentration with serum TC

in residents of the Little Hocking water district in Ohio. The study population (n = 371, 2–>60 

years of age) was a random sample of the population served by LHWA. The subjects completed 

questionnaires (e.g., demographic, occupational, health conditions, and so forth) and provided 

blood samples. PFOA concentration was determined by HPLC/MS/MS; no other PFASs were 

measured. Regression models were used to analyze the data. The median serum PFOA 

concentration was 0.354 μg/mL. No association was observed between serum PFOA and TC. 

Steenland et al. (2009) examined the association of PFOA with serum lipids in adult 

participants of the C8 Health Project (n = 46,294; 18–≥80 years). Serum samples were separated 

into deciles or quartiles for analysis. TC, HDL, triglycerides, LDL, and non-HDL (TC minus 

HDL cholesterol) were measured or calculated from blood samples. The data were analyzed by 

linear regression using the log-transformed values for all variables. Covariates of the model 

Perfluorooctanoic acid (PFOA) – May 2016 3-5 
included age, gender, quartile BMI, education, smoking, regular exercise, and alcohol 

consumption. A logistic model was used to analyze high cholesterol and serum PFOA 

concentration (quartiles). The mean serum PFOA concentration was 0.080 μg/mL. All lipid 

outcomes, except for HDL, showed significant increasing trends with increasing serum PFOA 

decile. There was a positive association between mean levels of serum PFOA and TC, LDL 

cholesterol, triglycerides, TC/HDL ratio, and non-HDL. The predicted increase in TC from 

lowest to highest serum PFOA concentration decile was 11–12 milligrams per deciliter (mg/dL). 

The odds ratio (OR) for high cholesterol (≥240 mg/dL) increased from the lowest to the highest 

quartile of serum PFOA concentrations: 1.00, 1.21 (95% CI: 1.12–1.31), 1.33 (95% CI: 1.23–

1.43), and 1.38 (95% CI: 1.28–1.51). No association was observed between mean level of serum 

PFOA and HDL cholesterol. PFOS also was positively associated with TC, LDL cholesterol, and 

triglycerides. The results of the study were consistent with occupational studies that found a 

positive association between PFOA exposure and serum lipids. 

The study by Frisbee et al. (2010) used a design and analysis strategy similar to that of 

Steenland et al. (2009), but it was conducted among children (n = 6536; 1–11.9 years) and 

adolescent (n = 5934; 12.0–17.9 years) participants of the C8 Health Project. The mean serum 

PFOA concentration was 0.0777 μg/mL and 0.0618 μg/mL, respectively, for children and 

adolescents. TC, LDL, and triglycerides were positively associated (p≤0.02) with serum PFOA 

concentration, adjusting for age, gender, BMI, exercise, and length of fast. Assessment of the 

quintile trends showed significant differences (p≤0.02) between the first and fifth quintile for TC

and LDL for children and adolescents of both genders combined and separated. A significant 

difference (p = 0.04) was observed for fasting triglycerides in female children only. An increased 

risk of abnormal TC and LDL were positively associated with serum PFOA. The ORs were 

1.0 first (reference), 1.1 (95% CI: 1.0–1.3, second), 1.2 (95% CI: 1.0–1.4, third), and 1.2 (95% 

CI: 1.1–1.4, fourth and fifth) for TC, and 1.0 (reference, first), 1.2 (95% CI: 1.0–1.5, second), 

1.2 (95% CI: 1.0–1.4, third and fourth), and 1.4 (95% CI: 1.2–1.7, fifth) for LDL. An increased 

risk of abnormal fasting triglyceride and HDL was not associated with serum PFOA. PFOS also 

was positively associated with TC, LDL cholesterol, and HDL cholesterol. 

The C8 Science Panel (2012) used data from the C8 general population cohorts as well as 

from combined general population and worker cohorts to evaluate the association between PFOA 

and a medical diagnosis of high cholesterol. Despite inconsistent evidence between studies, they 

concluded that there is a probable link between PFOA and diagnosed high cholesterol. The 

worker cohort was not evaluated separately in this analysis.

A cohort of 521 members of the C8 Health Project was evaluated for an association between 

changes in serum PFOA levels and changes in serum LDL-cholesterol, HDL-cholesterol, TC, 

and triglycerides over a 4.4-year period (Fitz-Simon et al. 2013). Linear regression models were 

fit to the logarithm (base 10) of ratio change in each serum lipid measurement in relation to the 

logarithm of ratio change in PFOA. Mean serum PFOA concentration decreased by approximately 

one-half between baseline (0.140 ± 0.209 μg/mL) and follow-up (0.068 ± 0.144 μg/mL). No 

corresponding changes in serum lipids were found. However, those individuals with the greatest 

declines in serum PFOA had a larger decrease in LDL cholesterol.

Perfluorooctanoic acid (PFOA) – May 2016 3-6 
Table 3-1. Summary of PFOA Occupational Exposure Studies of PFOA and Serum Lipids

Reference and Study Details PFOA Level TC LDL HDL Triglycerides

Cross-sectional

Olsen et al. 2000

n = 111 in 1993, 80 in 1995, 74 in 1997; 

50-70% participation rate

Mean age: ~ 40 yrs

Mean duration: not reported

ANOVA based on three exposure 

categories, adjusted

(1) 0 to < 1, mean ~ 

0.4 μg/mL

(2) 1 to < 10, mean ~ 

3 μg/mL

(3) ≥ 10, mean ~ 30 

μg/mL

1993: mean 215, 219, 232 

mg/dl (p = 0.45)

1995: mean 207, 212, 221 

mg/dl (p = 0.48)

1997: mean 199, 213, 217 

mg/dl (p = 0.08) 

1993: mean 138, 143, 140 

mg/dl (p = 0.84)

1995: mean 131, 133, 130 

mg/dl (p = 0.96) 

1997: mean 114, 134, 134 

mg/dl (p = 0.11) 

1993: mean 43, 47, 48 

mg/dl (p = 0.32)

1995: mean 42, 43, 41 

mg/dl (p = 0.70)

1997: mean 41, 44, 45 

mg/dl (p = 0.40) 

1993: mean 171, 205, 

221 mg/dl (p = 0.77)

1995: mean 152, 123, 

183 mg/dl (p = 0.07)

1997: mean 219, 176, 

251 mg/dl (p = 0.13)

Olsen and Zobel 2007

3M. Antwerp, Cottage Grove, Decatur 

combined; 50-65% participation rate

n = 506 (men, not taking lipid-lowering 

medications) 

Mean age: 40 yrs

Mean duration: not reported

Linear regression, adjusted

[Related reference: Olsen et al. 2003]

Mean 2.21 μg/mL

range 0.01–92.03 

μg/mL

Beta = 0.0076

(SE 0.0059)

(p = 0.20)

[log-transformed PFOA 

and cholesterol]

Beta = 0.0021

(SE 0.0090)

(p = 0.81)

[log-transformed PFOA 

and cholesterol]

Beta = -0.0183

(SE 0.0069) 

(p = 0.01)

[log-transformed PFOA 

and cholesterol]

Beta = 0.0711

(SE 0.0169)

(p = 0.0001)

[log-transformed PFOA 

and cholesterol]

Sakr et al. 2007a

Washington Works (West Virginia)

n = 1025 (782 men, 243 women), 55% 

participation rate

Mean age: 46.5 and 44.4 yrs, respectively 

for men and women

Mean duration: 19.6 and 15.9 yrs, 

respectively for men and women

Linear regression, adjusted

Mean 0.428 μg/mL

range 0.005– 9.55 

μg/mL

All: Beta = 4.036

(SE 1.284) (p = 0.002)

Excluding workers taking 

lipid-lowering 

medications: 

Beta = 5.519

(SE 1.467)

p = < 0.001)

All: Beta = 2.834

(SE 1.062) (p = 0.008)

Excluding workers taking 

lipid-lowering 

medications: 

Beta = 3.561

(SE 1.213) (p = 0.003)

All: Beta = -0.178

(SE 0.432) (p = 0.68)

Excluding workers taking 

lipid-lowering 

medications: 

Beta = 0.023

(SE 0.058) (p = 0.96)

All: Beta = 018

(SE 0.021) (p = 0.38)

Excluding workers 

taking lipid-lowering 

medications: 

Beta = 0.030

(SE 0.024 (p = 0.21)

[TG log-transformed]

Costa et al. 2009

Italy PFOA production plant

n = 37 currently exposed, 16 formerly 

exposed, 107 controls (different areas in 

the plant)

Mean age: 42 yrs—currently exposed and 

controls; 52 yrs—formerly exposed

Mean duration (in 2007): 14-16 yrs

Analysis 1: t-test, 34 currently exposed 

matched to 34 controls, adjusted

Analysis 2: Linear regression, 34 currently 

exposed and 107 controls, adjusted

Analysis 3: linear regression (generalized 

estimating equation [GEE] modeling), 56 

total with concurrent PFOA and lipid 

measure, adjusted

Currently exposed:

mean 12.9, geometric 

mean 4.02, range 

0.2–47 μg/mL

Formerly exposed:

mean 6.81 geometric 

mean 3.76, range 

0.53–18 μg/mL

Analysis 1:

Currently exposed:

237.0 mg/dl

Controls: 206.4 mg/dl

(p = 0.003)

Analysis 2:

Beta = 21.7

(95% CI 6.83, 36.6)

(p = 0.005)

Analysis 3:

Beta = 0.028

(95% CI 0.002, 0.055)

(p < 0.05)

Not measured Analysis 1:

Currently exposed:

56.68 mg/dl

Controls: 57.82 mg/dl

(p => 0.05)

Analysis 2:

Beta = 2.42

(95% CI -2.30, 7.13)

(p > 0.05)

Analysis 3:

Beta = -0.018 

(95% CI -0.047, 0.012)

(p > 0.05)

Analysis 1:

Currently exposed:

150.03 mg/dl

Controls: 155.35 mg/dl

(p > 0.05)

Analysis 2:

Beta = -0. 15

(95% CI -34.6, 34.3)

(p > 0.005)

Analysis 3:

Beta = 0.055

(95% CI -0.036, 0.147)

(p > 0.05)

Perfluorooctanoic acid (PFOA) – May 2016 3-7 
Reference and Study Details PFOA Level TC LDL HDL Triglycerides

Longitudinal

Olsen et al. 2003

3M, Antwerp and Decatur combined

~5 yr follow-up period

n = 174 (measure in 1995 or 1997, and in 

2000) 

Mean age: not reported

Mean duration: not reported

Linear mixed effects regression for 

repeated measures, adjusted

1995 baseline: 1.36-

1.41 μg/mL

2000 follow-up: 

1.49-1.77 μg/mL

Beta = 0.032

(95% CI 0.013, 0.051)

[and statistically 

significant PFOA-years 

follow-up interaction, 

Beta = -0.0004] 

Not measured Reported as “no 

significant changes”

Beta = 0.094

(95% CI 0.045, 0.144)

Sakr et al. 2007b

Washington Works (West Virginia) 

n = 454

23-yr follow-up (mean 3.7 PFOA 

measures)

Mean age: 27 yrs (at hire)

Mean duration: 27 yrs

Linear mixed effects regression for 

repeated measures, adjusted

1.04 μg/mL (first)

1.16 μg/mL (last)

Declined since 1980 

(mean 4.78 μg/mL in 

1980 to 1.00 μg/mL 

in 2001–2004) 

Beta = 1.06

(95% CI 0.24, 1.88)

Beta = 0.46

(95% CI -0.87, 1.79)

Beta = 0.16

(95% CI -0.39, 0.71)

Beta = 0.79

(95% CI -5.99, 7.57)

Steenland et al. 2015

n=3,713 workers

Data collected in 2005-2006 and 2008-

2011 

n=1,298 cases

In 2005-2006: mean 

0.325 μg/mL, median 

0.113 μg/mL

HR (95% CI), for self-reported use of cholesterol-lowering medications (incidence based on year of diagnosis).

Cumulative exposure quartile, no lags:

1.00 (referent) 

1.11 (0.94, 1.30)

1.06 (0.89, 1.27)

1.05 (0.87, 1.27) (Ptrend = 0.56)

Perfluorooctanoic acid (PFOA) – May 2016 3-8 
Table 3-2. Summary of High-Exposure Community Studies of PFOA and Serum Lipids

Reference and Study Details PFOA level TC LDL HDL Triglycerides

Cross-sectional

Emmett et al. 2006

n = 371, aged 2–89 yrs, median 50 yrs

(317 from stratified random population sample)

Linear regression (continuous PFOA) and t-test, 

PFOA, abnormal versus normal cholesterol

0.354 μg/mL Regression: Beta = 

0.00551 (p = 0.27)

p-value of t-test = 0.79

[n = 182, 49% 

abnormal]

Not measured Not measured Not measured

Steenland et al. 2009

n = 46,294, aged 18-80 yrs (not taking 

cholesterol-lowering medications)

Linear regression, quartiles PFOA and continuous 

PFOA

Mean 0.08 

μg/mL

Quartiles 

0- 0.0131 

0.0132-0.0265 

0.0266-0.0669 

≥ 0.067

Beta = 0.01112

(SE 0.00076)

[log PFOA and lipids]

By quartiles (OR):

1.0 (referent)

1.21 (1.12, 1.31)

1.33 (1.23, 1.43)

1.38 (1.28, 1.50)

Beta = 0.01499

(SE 0.00121)

[log PFOA and lipids]

Beta = 0.00276

(SE 0.00094)

[log PFOA and lipids]

Beta = 0.00169

(SE 0.00219)

[log PFOA and lipids]

Frisbee et al. 2010

6,536 children 1-< 12 yrs

5,934 adolescents, 12-18 yrs

Linear regression, adjusted

Mean μg/mL

Children 0.0777

Adolescents 

0.0618 

Difference between 1st

and 5th quintile PFOA 

(trend p):

Children 5.8 mg/dl

 (p < 0.0001)

Adolescents 4.2 mg/dl

 (p < 0.0001)

Difference between 1st

and 5th quintile PFOA 

(trend p):

Children 4.9 mg/dl

 (p = 0.001)

Adolescents 3.2mg/dl

 (p = 0.004)

Difference between 1st

and 5th quintile PFOA 

(trend p):

Children 5.8 mg/dl

 (p = 0.88)

Adolescents 4.2 mg/dl

 (p = 0.20)

Difference between 1st

and 5th quintile PFOA 

(trend p):

Children 2.0 mg/dl

 (p = 0.10)

Adolescents 3.8 mg/dl

 (p = 0.10)

Longitudinal (Change in Lipid in Relation to Change in PFOA)

Fitz-Simon et al. 2013

Longitudinal; 4.4 yrs n = 521

Linear regression of log of ratio change in serum 

lipid to log of ratio change in PFOA, adjusted for 

age, gender, interval between measures, fasting 

status (change in lipid in relation to change in 

PFOA)

0.140 μg/mL 

(baseline)

0.068 μg/mL 

(follow-up) 

Percent decrease (95% 

CI) in lipid per halving 

PFOA:

1.65 (0.32, 2.97);

with additional 

adjustment for PFOS:

0.63 (-0.88, 2.12)

Percent decrease (95% 

CI) in lipid per halving 

PFOA:

3.58 (1.47, 5.66);

with additional 

adjustment for PFOS:

2.92 (0.71, 5.09)

Percent decrease (95% 

CI) in lipid per halving 

PFOA:

1.33 (-0.21, 2.85);

with additional 

adjustment for PFOS:

1.24 (-0.34, 2.79)

Percent decrease (95% 

CI) in lipid per halving 

PFOA:

-0.78 (-5.34, 3.58);

with additional 

adjustment for PFOS:

-1.16 (-5.85, 3.33)

Incidence of Hypercholesterolemia

Winquist and Steenland 2014a

n = 32,254 (including 3,713 workers)

Data collected in 2005-2006 and 2008-2011 

n = 9,653 cases in primary analysis (all diagnoses)

n = 1,825 cases in prospective analysis (diagnoses 

after 2005-2006) 

In 2005-2006: 

mean 0.0866 

μg/mL, median 

0.0261 μg/mL

HR (95% CI), self-reported use of cholesterol-lowering medications, primary analysis

Cumulative exposure quintiles: Year exposure quintiles:

1.00 (referent) 1.00 (referent)

1.24 (1.15, 1.33) 1.07 (1.01, 1.15)

1.17 (1.09, 1.26) 1.11 (1.04, 1.19)

1.19 (1.11, 1.27) 1.05 (0.99, 1.13)

1.19 (1.11, 1.28) (Ptrend = 0.005) 1.20 (1.12, 1.28) (Ptrend = 0.001)

Diagnoses after 2005: no association with PFOA with either exposure metric

Perfluorooctanoic acid (PFOA) – May 2016 3-9 
More recently, participants in the C8 Health Project were examined for an association 

between PFOA levels and incidence of several conditions, including high cholesterol (based on 

prescription medication use) (Winquist and Steenland 2014a). The cohort included 28,541 

community members and 3,713 workers who had completed study questionnaires during 2008–

2011. The median serum PFOA level at enrollment in 2005–2006 was 0.0261 μg/mL for the 

combined cohort, 0.0242 μg/mL for the community members, and 0.1127 μg/mL for the 

workers. Retrospective serum levels for the community cohort were estimated from air and water 

concentrations, residential history, and water consumption rates. For the workers, yearly serum 

estimates were modeled from work history information and job-specific concentrations. Cox 

proportional hazard models, stratified by birth year, were used to assess self-reported adult heart

disease hazard in relation to time-varying yearly or cumulative (sum of yearly estimates) 

estimated PFOA serum concentration, controlling for gender, race, education, smoking, and 

alcohol consumption. Using the cumulative exposure metric, the HRs for hypercholesterolemia 

for quintiles 2–5 versus quintile 1 were 1.24, 1.17, 1.19, and 1.19 (Ptrend = 0.005). Using the 

yearly exposure metric, the HRs for high cholesterol for quintiles 2–5 versus quintile 1 were 

1.07, 1.11, 1.05, and 1.20 (Ptrend = 0.001). The strongest association was in males aged 40–59. 

No associations were found between PFOA level and hypertension or coronary artery disease 

incidence. (The analysis of these data restricted to the worker population by Steenland et al. 

[2015] is described in the previous section).

A subset of 290 individuals in the C8 Health Project was evaluated for evidence that PFOA 

exposure can influence the transcript expression of genes involved in cholesterol metabolism, 

mobilization, or transport (Fletcher et al. 2013). RNA was extracted from whole blood samples 

taken from 144 males and 146 females aged 20–60 years; serum collected at the same time was 

used to measure PFOA concentration. The association between candidate gene expression levels 

and PFOA levels was assessed by multivariable linear regression with adjustments for 

confounders. Inverse associations were found between PFOA levels and expressions of 

transcripts involved in cholesterol transport (NR1H2, NPC1, and ABCG1; p = 0.002, 0.026, and 

0.014, respectively). When genders were analyzed separately, PFOA was negatively associated 

with expression of genes involved in cholesterol transport in males (NPC1, ABCG1, PPARα) 

and females (NCEH1). Similar associations were found with PFOS. 

General population studies. Several studies examined serum lipids in the general population 

(Table 3-3). Nelson et al. (2010) examined the relationship between polyfluoroalkyl chemical 

serum concentration, including PFOA, and lipid and weight outcomes in the general population 

of the United States by analyzing data from the 2003–2004 NHANES. The population (n = 860) 

included persons aged 20–80 years with no missing covariate information who were not 

pregnant, breast-feeding, taking insulin or cholesterol medicine, or undergoing dialysis. 

Cholesterol (TC, HDL, LDL) was measured from serum samples. Data for covariates predicting 

cholesterol and body weight including age, gender, race/ethnicity, socioeconomic status, 

saturated fat intake, exercise, alcohol consumption at ≥ 20 years of age, smoking, and parity 

were obtained from the questionnaires. Regression analyses were performed for gender and the 

age groups 12–19 years, 20–59 years, and 60–80 years. The mean PFOA concentration was 

0.0046 ± 0.003 μg/mL. A positive association was found between TC and non-HDL (TC-HDL, 

~70–80% TC) cholesterol and serum PFOA (effect estimate 9.8; 95% CI, -0.2–19.7). No 

association was found between serum PFOA concentration and HDL, or LDL. No association 

was found between serum PFOA concentration and body weight. Similar results were found 

with PFOS. A similar analysis using 1999–2008 NHANES data for 815 adolescents (aged 12–18 

years) by Geiger et al. (2014a) found an association between serum PFOA and TC

Perfluorooctanoic acid (PFOA) – May 2016 3-10
(Beta 4.55, 95% CI 0.90, 8.20, per ln-unit increase in PFOA) and LDL (Beta 5.75, 95% CI 2.16, 

9.33, per ln-unit increase in PFOA). 

Eriksen et al. (2013) examined the association between plasma PFOA (and PFOS) levels and 

TC levels in a middle-aged Danish population. This cross-sectional study included 663 males 

and 90 females aged 50–65 years who were enrolled in the Danish Diet, Cancer and Health 

cohort. Generalized linear models were used to analyze the association between PFOA and TC 

levels, adjusted for age, gender, education, BMI, smoking, alcohol consumption, egg intake, 

animal fat intake, and physical activity. The mean plasma PFOA level was 0.0071 μg/mL. A 

significant, positive association was found between PFOA (and PFOS) and TC such that, in the 

fully adjusted model, a 4.4-mg/dL (95% CI 0.8, 8.5) higher concentration of TC was found per 

interquartile range of plasma PFOA (quartile cut-points were not reported). 

Fisher et al. (2013) examined the association of plasma PFAS levels, including PFOA, with 

metabolic function and plasma lipid levels. This population-based sample included 2,700 

participants aged 18–74 years (~50% male) in the Canadian Health Measures Survey. The 

geometric mean PFOA concentration was 0.0025 ± 0.0018 μg/mL. In analyses that included 

sampling weights, no associations were found between PFOA (or PFOS) and TC, HDL- and 

LDL-cholesterol, and metabolic syndrome and glucose homeostasis parameters. Covariates 

considered included age, gender, marital status, income adequacy, race, education, BMI, physical 

activity, smoking, and alcohol consumption. 

Starling et al. (2014) examined the association between PFOA (and six other PFASs) and 

serum lipids in pregnant females in the Norwegian Mother and Child Cohort Study. Most of the 

blood samples were drawn during weeks 14–26 of gestation. Weighted multiple linear regression 

was used to estimate the association between PFOA level and each lipid level. Covariates 

considered included age, prepregnancy BMI, nulliparous or interpregnancy interval, breastfeeding duration, education, current smoking, gestation week at blood draw, oily fish 

consumption, and weight gain during pregnancy. The median plasma PFOA level was 

0.00225 μg/mL. No association was observed between PFOA and triglycerides, TC, or LDLcholesterol. PFOA was positively associated with HDL-cholesterol, although the CI was large 

for the association. With HDL-cholesterol, each interquartile range- (IQR-) unit increase in lnPFOA was associated with an increase of 1.28 mg/dL (95% CI: -0.15, 2.71). Five of the seven 

PFASs studied were positively associated with HDL cholesterol and all seven had elevated HDL 

associated with the highest quartile.

Perfluorooctanoic acid (PFOA) – May 2016 3-11
Table 3-3. Summary of General Population Epidemiology Studies of PFOA with Serum Lipids

Reference and Study Details PFOA level TC LDL HDL Triglycerides

All Adults

Nelson et al. 2010

United States, NHANES (2003–2004) 

n = 860, aged 20-80 yrs

(451 men, 409 women) 

Linear regression, PFOA in quartiles and 

continuous PFOA, adjusted

PFOA-PFOS correlation Spearman r = 0.65

Mean 0.0046 μg/mL, 

median 0.0038 μg/mL

Beta = 1.22

(95% CI 0.04, 2.40)

9.8 mg/dl increase in top 

versus bottom quartile

(PFOS results similar)

Beta = -0.21

(95% CI -1.91, 1.49)

(PFOS results similar)

Beta = -0.12

(95% CI -0.41, 0.16)

Different pattern seen with 

PFOS

Not measured

Eriksen et al. 2013

Denmark 

n = 753, aged 50-65 yrs

(663 men, 90 women) 

Linear regression, continuous PFOA, adjusted

PFOA-PFOS correlation not reported

Mean 0.0071 μg/mL 4.4 mg/dl increase per 

IQR

(PFOS results similar)

Not measured Not measured Not measured

Fisher et al. 2013

Canada, Canadian Health Measures Survey

n = 2,700, aged 18-74 yrs

Linear regression, continuous PFOA (logtransformed PFOA and lipids)

PFOS correlation r = 0.36

Mean 0.0025 μg/mL

Quartiles:

0.00015-0.00185

0.00186-0.00258

0.00259-0.00355

≥ 0.0036

Beta = 0.03

(95% CI -0.017, 0.07)

[log PFOA and lipids]

Beta = 0.02

(95% CI -0.06, 0.091)

[log PFOA and lipids]

Beta = 0.0009

(95% CI -0.04, 0.04)

[log PFOA and lipids]

Not measured

Pregnant Women

Starling et al. 2014

Norway

n = 891 pregnant women

Plasma PFOA (collected in 2nd trimester)

Linear regression, continuous PFOA (logtransformed PFOA), adjusted

PFOS correlation Spearman r = 0.64

Median 0.00225 

μg/mL

Beta = 2.58

(95% CI -4.32, 9.47)

[per ln-unit increase in 

PFOA]

Beta = 2.25

(95% CI -3.97, 8.48) 

[per ln-unit increase in 

PFOA]

Beta = 2.13

(95% CI -0.26, 4.51)

[per ln-unit increase in 

PFOA]

Beta = 0.00

(95% CI -0.07, 

0.06) 

[per ln-unit increase 

in PFOA]

Adolescents

Geiger et al. 2014a

United States, NHANES (1999-2008) 

n = 815, aged 12-18 yrs

Linear regression, continuous PFOA (logtransformed PFOA), adjusted

Mean 0.0042 μg/mL Beta = 4.55

(95% CI 0.90, 8.20)

[per ln-unit increase in 

PFOA]

(PFOS results similar) 

Beta = 5.75

(95% CI 2.16, 9.33)

[per ln-unit increase in 

PFOA]

(PFOS results similar) 

Beta = -1.52

(95% CI -3.02, -0.03 

[per ln-unit increase in 

PFOA]

Attenuated results when 

adjusted for PFOS

Beta = 1.74

(95% CI -2.88, 

6.36) 

Different pattern 

seen with PFOS

Perfluorooctanoic acid (PFOA) – May 2016 3-12
The association between PFOA and serum lipids has been examined in several studies in 

different populations. Cross-sectional and longitudinal studies in occupational settings (Costa et 

al. 2009; Olsen et al. 2000, 2003; Olsen and Zobel 2007; Sakr et al. 2007a, 2007b; Steenland et 

al. 2015) and in the high-exposure community (the C8 Health Project study population) (FitzSimon et al. 2013; Frisbee et al 2010; Steenland et al. 2009; Winquist and Steenland 2014a) 

generally observed positive associations between serum PFOA and TC in adults and children 

(aged 1–< 18 yrs); most of the effect estimates were statistically significant. Although exceptions 

to this pattern are present (i.e., some of the analyses examining incidence of self-reported high 

cholesterol based on medication use in Winquist and Steenland [2014a] and in Steenland et al. 

[2015]), the results are relatively consistent and robust. Similar associations were seen in 

analyses of LDL, but were not seen with HDL. The range of exposure in occupational studies is 

large (with means varying between 0.4 and > 12 μg/mL), and the mean serum levels in the C8 

population studies were around 0.08 μg/mL. Positive associations between serum PFOA and TC 

(i.e., increasing lipid level with increasing PFOA) were observed in most of the general 

population studies at mean exposure levels of 0.002–0.007 μg/mL (Eriksen et al. 2013; Fisher et 

al. 2013; Geiger et al. 2014a; Nelson et al. 2010; Starling et al. 2014). The interpretation of these 

general population results is limited, however, by the moderately strong correlations (Spearman 

r > 0.6) and similarity in results seen for PFOS and PFOA. 

3.1.1.2 Cardiovascular Diseases

Occupational exposure studies. Several studies examined cardiovascular-related cause of death 

among PFOA-exposed workers at the West Virginia Washington Works plant (Leonard et al. 

2008; Sakr et al. 2009; Steenland and Woskie 2012) and the 3M Cottage Grove plant in 

Minnesota (Lundin et al. 2009; Gilliland and Mandel 1993). This type of mortality is of interest 

because of the relation between lipid profiles (e.g., LDL) and the risk of cardiovascular disease. 

The most recent West Virginia study included 5,791 individuals who had worked at the plant for 

at least 1 year between 1948 and 2002, with mortality follow-up through 2008. No associations 

were found between cumulative PFOA levels and ischemic heart disease (IHD) mortality 

(standardized mortality ratio [SMR] 1.07, 1.02, 0.87, and 0.93 across four quartiles of cumulative 

exposure, compared to U.S. referent group). Based on these data from the worker cohorts, the C8 

Science Panel (2012) concluded that there is no probable link between PFOA and stroke and 

coronary artery disease.

The analysis of the Minnesota plant (n = 3,993 workers who began work between 1983 and 

1997, with follow-up through 2002) also found no association between cumulative PFOA 

exposure and IHD risk, but an increased risk of cerebrovascular disease mortality was seen in the 

highest exposure category (HR 2.1, 95% CI 1.0, 4.6). These studies are limited by the reliance on 

mortality (rather than incidence) data, which can result in a substantial degree of under 

ascertainment and misclassification. 

3.1.1.3 Liver Enzymes and Liver Disease

Cross-sectional studies and longitudinal studies of PFOA and liver enzymes in various 

populations are described in this section and summarized in Table 3-4. 

Perfluorooctanoic acid (PFOA) – May 2016 3-13
Table 3-4. Summary of Epidemiology Studies of PFOA and Liver Enzymes

Reference and Study Details PFOA level Results

Cross-sectional: Occupational Exposure Studies

Olsen et al. 2000

n = 111 in 1993, 80 in 1995, 74 in 1997; 50-70% 

participation rate

Mean age: ~ 40 yrs

Mean duration: not reported

ANOVA and linear regression adjusted for age, 

BMI, and alcohol and cigarette use

Mean (range) μg/mL

1993: 5 (0-80)

1995: 6.8 (0-114) 

1997: 6.4 (0.1-81) 

ALT: Year β (±SE) (p value)

1993: 0.89 (2.88) (p = 0.76) 

1995: 0.81 (2.62) (p = 0.75) 

1997: 2.77 (1.27) (p = 0.03) 

Change per 10 μg/mL increase in serum PFOA; stronger association in individuals with 

BMI< 30. No associations for ALP, GGT, AST, total bilirubin, direct bilirubin.

Olsen and Zobel 2007

3M. Antwerp, Cottage Grove, Decatur combined; 

50-65% participation rate

n = 506 (men, not taking lipid-lowering medications)

Mean age: 40 yrs

Mean duration: not reported

Linear regression adjusting for ln age, ln BMI, ln 

alcohol 

[Related reference: Olsen et al. 2003)

Mean (range)

2.21 (0.01 – 92.03) 

μg/mL

β change per ln PFOA (±SE) p value

ln ALP: All 0.009 (± 0.008) (p = 0.25) Decatur: 0.08 (0.34) (p = 0.02)

ln AST: All -0.005 (± 0.009) (p = 0.55) Decatur: 0.011 (0.02) (p = 0.57)

ln ALT: All 0.025 (± 0.013) (p = 0.06) Decatur: 0.08 (0.034) (p = 0.02)

ln GGT: All 0.033 (± 0.017) (p = 0.05) Decatur: 0.08 (0.034) (p = 0.02)

ln total bilirubin: All -0.033 (± 0.01) (p = 0.001)Decatur: -0.054 (± 0.021) (p = 0.01)

Replacement of ln BMI with triglycerides in the model resulted in reduced associations for 

ALT and GGT.

Sakr et al. 2007a

Washington Works (West Virginia)

n = 1025 (782 men, 243 women), 55% participation 

rate 

Mean age: 46.5 and 44.4 yrs, respectively for men 

and women

Mean duration: 19.6 and 15.9 yrs, respectively for 

men and women

Linear regression, adjusting for age, BMI, alcohol 

consumption, gender, history of heart attack in 

parent, use of lipid-lowering medications

0.428 μg/ml

LOQ 0.0005 μg/ml 

range 0.005 – 9.55 

μg/mL

β (±SE) p value: Full sample Excluding 178 men on lipid-lowering medications

ln AST: 0.012 (± 0.012) (p = 0.317) ln AST: 0.023 (±0.013) (p = 0.079)

ln ALT: 0.023 (± 0.015) (p = 0.124) ln ALT: 0.031 (±0.017) (p = 0.071)

ln GGT: 0.048 (± 0.02) (p = 0.016) ln GGT: 0.05 (±0.023) (p = 0.03)

ln bilirubin: 0.008 (± 0.014) (p = 0.59) ln bilirubin: 0.1 (±0.017) (p = 0.637)

Costa et al. 2009

Italy

Cross-sectional

56 male workers (currently and formerly exposed 

and unexposed) with concurrent serum PFOA and 

clinical parameters measured in last 7 yrs

GEE models adjusting for age, years of exposure,

year of PFOA sampling, BMI, smoking, and alcohol 

consumption

Currently exposed:

mean 12.9, geometric 

mean 4.02, range 0.2-47 

Formerly exposed:

mean 6.81 geometric 

mean 3.76, range 0.53-

18 μg/mL

β change per μg PFOA/mL (95% CI)

AST: 0.038 (-0.003, 0.080)

ALT: 0.116 (0.054, 0.177)

GGT: 0.177 (0.076, 0.278)

ALP: 0.057 (0.007, 0.107)

Total bilirubin: -0.080 (-0.137, -0.024) 

Conj. bilirubin: -0.034 (-0.09, 0.031)

Perfluorooctanoic acid (PFOA) – May 2016 3-14
Reference and Study Details PFOA level Results

Longitudinal: Occupational Exposure Studies

Olsen et al. 2003

3M, Antwerp and Decatur combined

~5 yr follow-up period

n = 174 (measure in 1995 or 1997, and in 2000)

Mean age: not reported

Mean duration: not reported

Linear mixed effects regression for repeated 

measures, adjusted

1995 baseline:

1.36-1.41 μg/mL

2000 follow-up: 1.49-

1.77 μg/mL

No associations observed; however, data not provided

Sakr et al. 2007b

Washington Works (West Virginia)

n = 454 

23-yr follow-up (mean 3.7 PFOA measures)

Mean age: 27 yrs (at hire)

Mean duration: 27 yrs

Linear mixed effects regression for repeated 

measures, adjusted

1.04 μg/mL (first)

1.16 μg/mL (last)

Used PFOA 

measurement from same 

year as biomarker test or 

interpolated using two 

surrounding values

β IU/L change per 1 μg/mL PFOA (95% CI)

ALP: (n = 1327) -0.21 (-0.60, 0.18)

AST: (n = 1326) 0.35 (0.10, 0.60)

ALT: (n = 231) 0.54 (-0.46, 1.54)

GGT: (n = 233) 1.24 (-1.09, 3.57)

Total bilirubin: (n = 1327) 0.008 (-0.0139, -0.0021) 

Cross-sectional: High-Exposure Community Studies

Emmett et al. 2006

n= 371, aged 2–89 yrs, median 50 yrs

(317 from stratified random population sample)

Linear regression (continuous PFOA) and t-test, 

PFOA, abnormal versus normal enzyme levels

Median 0.354 μg/mL 

(IQR 0.184 – 0.571 

μg/mL); nonfasting 

blood sample

Linear regression, Beta (p-value) n (%) abnormal,(t-test p-value)

ALP: -0.00416 (p = 0.65) 6 (2%) (p = 0.63)

AST: -0.0007586 (p = 0.76) 9 (2%) (p = 0.03)

ALT: -0.00183 (p = 0.65) 28 (8%) (p = 0.30)

GGT: 0.00057711 (p = 0.89) 11 (3%) (p = 0.50)

Gallo et al. 2012

West Virginia, United States; C8 Health Project,

46,452 of 56,554 (82.1%) adults

Adjusting for age, gender, physical activity, BMI, 

average household income, educational level, fasting 

status, month of blood sample collection, race, 

insulin resistance, alcohol consumption, and cigarette 

smoking 

Median 0.028 μg/mL 

(IQR 0.135 – 0.71 

μg/mL) nonfasting 

blood sample; LOD 

0.0005 μg/mL, n = 32 

below LOD

Linear regression, Logistic regression of abnormal values

β per 1 unit increase PFOA (95% CI) OR (95% CI)

Ln ALT: 0.022 (0.018 – 0.025) ALT: 1.10 (1.07, 1.13)

ln GGT: 0.015 (0.01 – 0.019) GGT: 1.01 (0.99, 1.04)

n Direct (conjugated) bilirubin: 0.001 (-0.002 – 0.004) Direct bilirubin 0.97 (0.90, 1.05)

Analysis of Ln ALT or Ln GGT by decile showed increase from 0.005 to 0.030 μg/mL, then 

leveling; p value for trend < 0.001; Direct bilirubin showed a U-shaped relation increasing to 

0.030 μg/mL, then declining.

Cross-sectional: General Population Studies

Lin et al. 2010

United States, NHANES (1999-2000; 2003-2004) 

1,076 men, 1,140 women of 10,224 enrolled, 

excluding < 6-hr fast n = 1,802), and missing 

covariate or no serum PFOA or liver function of 

metabolic syndrome data

Adjusting for age, gender, race/ethnicity, smoking, 

alcohol consumption, education level, BMI, 

metabolic syndrome, and iron saturation status

Geometric mean

0.00505 μg/mL

0.00406 μg/mL; 0.4% 

of samples below LOD 

(LOD 0.0002 and 

0.0001 μg/mL in 1999-

2000 and 2003-2004) 

β per unit (ng/ml) increase in log serum PFOA (95% CI)

ALT: (U/l) 1.86 (1.24, 2.48)

Log GGT: (U/l) 0.08 (0.05, 0.11)

β per unit (ng/ml) increase in log serum PFOA (95% CI); Same model as above, also 

controlling for other PFASs

ALT: (U/l) 2.19 (1.4, 2.98)

Log GGT: (U/l) 0.15 (0.11, 0.19)

Perfluorooctanoic acid (PFOA) – May 2016 3-15
Liver Enzymes

Occupational exposure studies. Olsen et al. (2000) analyzed alkaline phosphatase (ALP), GGT, 

aspartate aminotransferase (AST), ALT, and total- and direct bilirubin data from voluntary 

medical surveillance examinations of PFOA production workers at a 3M plant in 1993, 1995, 

and 1997. No association was observed between serum PFOA concentration and the parameters 

explored in cross-sectional analyses in the workers in 1993 or 1995; although in 1997 increases 

in AST per unit increase in serum PFOA concentration were observed. When measurements for 

all years were combined in longitudinal analyses (Olsen et al. 2003), the authors reported that no 

associations were observed with serum PFOA levels. Other than the analyses of AST, however, 

quantitative results were not provided.

A subsequent analysis involving these fluorochemical workers and an additional plant 

(Cottage Grove, Minnesota) that used medical surveillance data collected in 2000 examined the 

association between serum PFOA concentration and liver enzymes (Olsen and Zobel 2007). 

Serum samples were analyzed for ALP, GGT, AST, ALT, and bilirubin concentrations. Ln 

serum PFOA was marginally associated with ln ALT and ln GGT in regression models adjusting 

for ln age, ln BMI, and ln alcohol consumption, although the association was reduced when ln 

triglycerides replaced ln BMI in the model. An inverse association between total bilirubin and 

serum PFOA concentration (p<0.05) was observed at all sites combined. 

Sakr et al. (2007a) examined the relationship between serum PFOA and several clinical 

chemistry parameters in workers at the Washington Works plant in West Virginia. A complete 

blood count, metabolic panel (glucose, blood urea nitrogen [BUN], creatinine, iron, uric acid, 

electrolytes, creatinine kinase, lactic dehydrogenase [LDH], ALP, protein, albumin, C-reactive 

protein), liver enzyme panel (AST, ALT, GGT, bilirubin), and serum PFOA concentration were 

determined from the blood samples. Serum PFOA was associated (p<0.05) with increasing GGT 

in all of the participating workers. It was stated that an association was observed between serum 

PFOA concentration and iron, LDH, calcium, and potassium, but quantitative results were not 

included and the direction of association was not specified. 

Costa et al. (2009) also examined associations between serum PFOA concentration and liver 

enzymes in workers at a fluorochemical production plant in Italy. Serum PFOA concentration 

was associated with increasing ALT, GGT, and ALP levels (p<0.05), and inversely associated 

with total bilirubin (p<0.01) in 56 workers with PFOA and liver enzymes measured concurrently 

over the last 7 years. This subset of 56 workers included currently, formerly, and never exposed. 

Sakr et al. (2007b) also conducted a longitudinal study of liver enzymes among workers at 

the Washington Works plant with two or more PFOA measurements as described previously. 

Hepatic clinical chemistry (GGT, AST, ALT, ALP, total bilirubin), height, and weight data were 

analyzed. Serum PFOA concentration was associated in the model with increasing AST levels 

(p = 0.009) and inversely associated with total bilirubin (p = 0.006) after adjustment for age, 

BMI, gender, and decade of hire. No association was observed between serum PFOA 

concentration and GGT, ALT, and ALP. The regression models did not adjust for alcohol 

consumption, a potential limitation. 

High-exposure community studies. A small study (n = 371) of residents of the Little Hocking 

water district in Ohio found inconsistent results in different analyses of liver enzymes: An 

association with AST but not with ALP or ALT was seen when comparing serum PFOA levels 

between groups with abnormal compared to normal enzyme levels, but no association with any 

Perfluorooctanoic acid (PFOA) – May 2016 3-16
enzyme was seen in regression analyses with PFOA as a continuous variable (Emmett et al. 

2006). A subsequent study, which included a wider set of communities in the contaminated area, 

investigated the correlation between serum PFOA levels and liver enzymes in a total of 47,092 

samples collected from members enrolled in the C8 Health Project (Gallo et al. 2012). The 

association of ALT, GGT, and direct bilirubin with PFOA was assessed using linear regression 

models adjusted for various confounders. The median PFOA level was 0.028 μg/mL. The lntransformed values of ALT were significantly associated with ln-PFOA (and PFOS). There was a 

steady increase in fitted levels of ALT per decile of PFOA, leveling off after approximately 

0.030 μg PFOA/mL. Fitted values of GGT by deciles of PFOA showed a slight positive trend 

when adjusted for insulin resistance and BMI, but this was not confirmed in the logistic model 

analysis of elevated enzyme levels. Direct bilirubin levels appeared to increase at lower 

concentrations and then decline in a U-shaped pattern at 0.030 μg PFOA/mL. 

General population studies. Lin et al. (2010) investigated the association between serum PFOA 

(plus three other PFASs) and liver enzymes in the adult population of the United States by 

analyzing data from the 1999–2000 and 2003–2004 NHANES. The study population included 

2,216 adults (1076 males, 1140 females) older than 20 years who were not pregnant or nursing; 

had fasted more than 6 hours at the time of examination; were negative for hepatitis B or C virus; 

had body weight, height, educational attainment, and smoking status data available; and had serum 

tests for PFAS, liver function, or the five physiological measures associated with metabolic 

syndrome. Regression models were used to analyze the data and adjust for confounders. Mean 

PFOA levels were 0.00505 μg/mL and 0.00406 μg/mL for males and females, respectively. Serum 

PFOA concentration was divided into quartiles (Q1 = ≤ 0.0029; Q2 = ≤ 0.0042; Q3 = ≤ 0.00595; 

Q4 = > 0.00595 μg/ml). In the univariate regression models, liver enzymes, serum ALT, and logGGT increased across quartiles of PFOA (p ≤ 0.012), but total bilirubin showed no trend. The 

linear regression models were adjusted for (1) age, gender, and race/ethnicity; (2) age, gender, 

race/ethnicity, and lifestyle (smoking status, drinking status, education level), and (3) additional 

data for BMI, metabolic syndrome biomarkers, iron saturation status, and insulin resistance. An 

association was found between serum log-PFOA concentration and increasing serum ALT and logGGT. One unit increase in serum log-PFOA was associated with an increase of 1.86 units in serum 

ALT measurements and a 0.08-unit increase in log-GGT measurements. Effect modification was 

seen: For example, stronger associations between serum PFOA concentration and serum ALT (or 

GGT) were found among non-Hispanic Caucasians. PFOS also was positively associated with 

ALT in the fully adjusted model.

The results of the occupational studies provide evidence of an association with increases in 

serum AST, ALT, and GGT, with the most consistent results seen for ALT. The associations 

were not large and they might depend on the covariates in the models such as BMI, use of lipidlowering medications, and triglycerides (Costa et al. 2009; Olsen et al. 2000, 2003; Olsen and 

Zobel 2007; Sakr et al. 2007a, 2007b). Two population-based studies of highly exposed residents 

in contaminated regions near a fluorochemical industry in West Virginia have evaluated 

associations with liver enzymes, and the larger of the two studies reported associations of 

increasing serum ln ALT and ln GGT levels with increasing serum PFOA concentrations 

(Emmett et al. 2006; Gallo et al. 2012). A cross-sectional analysis of data from NHANES, 

representative of the U.S. national population, also found associations with ln PFOA 

concentration with increasing serum ALT and ln GGT levels. Serum bilirubin was inversely 

associated with serum PFOA in the occupational studies. A U-shaped exposure-response pattern 

for serum bilirubin was observed among the participants in the C8 Health Project, which might 

explain the inverse associations reported for occupational cohorts. Overall, an association of 

Perfluorooctanoic acid (PFOA) – May 2016 3-17
serum PFOA concentration with elevations in serum levels of ALT and GGT has been 

consistently observed in occupational and highly exposed residential communities, and the U.S. 

general population. The associations are not large in magnitude, but indicate the potential of 

PFOA to affect liver function.

Liver Diseases

High-exposure community studies. Few studies of the relationship between PFOA and liver 

disease are available, but the C8 Health Project did not observe associations with hepatitis, fatty 

liver disease, or other types of liver disease in their initial studies. The most recent update of 

disease incidence in the workers identified 35 cases of nonhepatitis liver disease (with medical 

validation) (Steenland et al. 2015); no association was seen with cumulative exposure when 

analyzed without a lag (HRs by quartile 1.0, 0.58, 1.43, 1.20; trend p = 0.86 for log cumulative 

exposure), but there was a possible trend in the analysis using a 10-year lag (HRs by quartile 1.0, 

1.46, 2.13, and 2.02; trend p = 0.40). 

3.1.1.4 Biomarkers of Kidney Function and Kidney Disease

Kidney Function 

PFOA has the potential to affect the kidney’s function of tubular resorption because of it uses

tubular transporters for excretion and resorption (see section 2.4). Since PFOA is removed from 

the blood by the kidney, cross-sectional analyses using serum PFOA as the exposure measure are 

problematic if individuals with compromised kidney function are included: PFOA concentrations 

could be increased in those individuals and an apparent association with GFR would be 

observed, even if one did not exist. Studies examining measures of kidney function are described 

in this section and summarized in Table 3-5. 

Table 3-5. Summary of Epidemiology Studies of PFOA and Measures of Kidney Function

Reference and Study Details PFOA Level Results

Sakr et al. 2007a

Washington Works plant

Cross-sectional; all active, 

nonpregnant employees 

enrolled over 12 days in 2004

1,025 of 1,863 eligible (55%)

0.428 μg/ml

LOQ 0.0005 μg/ml range 

0.005–9.55 μg/mL

Reported association with uric acid but quantified 

results were not provided

Costa et al. 2009

Italy

Cross-sectional

56 male workers (currently and 

formerly exposed and 

unexposed) with concurrent 

serum PFOA and clinical 

parameters measured in last 7 

yrs

Currently exposed:

mean 12.9, geometric 

mean 4.02, range 0.2–47 

μg/mL

Formerly exposed:

mean 6.81 geometric mean 

3.76, range 0.53–18 μg/mL

β change per μg PFOA/mL (95% CI)

Uric acid 0.026 (0.001, 0.053)

GEE models adjusting for age, years of exposure, 

year of PFOA sampling, BMI, smoking, and 

alcohol consumption

Perfluorooctanoic acid (PFOA) – May 2016 3-18
Reference and Study Details PFOA Level Results

Steenland et al. 2010

C8 Health Project, West 

Virginia

Cross-section; adult subjects (n

= 53,458; 20–≥80 years of age) 

participating in the C8 Health 

Project from 2005-2006

Subjects had consumed water 

for at least 1 year prior to 2004 

0.0864 μg/mL, measured 

in 2005–2006

Increased predicted uric acid of 0.2–0.3 μg/dL with 

increasing deciles of PFOA or PFOS

Shankar et al. 2011

United States, NHANES

Uric acid analysis:

1999–2000, 2003–2004 and 

2005–2006 cycles

3,883 out of 3,974 participants 

≥ 20 years of age with serum 

PFOA measurements; excluded 

subjects with missing data (n = 

91); 48.3% male, mean age 

46.4 years

eGFR analysis:

1999–2000, 2003–2004, 2005–

2006, and 2007–2008 cycles

4,587 out of 5,717 (80%) 

eligible 20 years or older with 

PFOA measures; excluded selfreported CVD (n = 572), 

missing data on serum 

creatinine or covariates (n = 

558)

0.0059 μg/mL; LOD 0.1 

ng/mL

Quartiles, μg/mL, n

1 < 0.0028 μg/mL, 1,176

2 0.0028–0.0041 μg/mL, 

1,141

3 0.0042–0.0059 μg/mL, 

1,141

4 > 0.0059 μg/mL, 1,129

Mean change in uric acid, mg/dL (95% CI) by 

quartile

1 referent

2 0.14 (0.04–0.25)

3 0.37 (0.25–0.49)

4 0.44 (0.32–0.56), p trend 0.0001

Mean change in uric acid, mg/dL (95% CI) by ln 

PFOA: 0.22 (0.15–0.30)

Multivariate regression adjusting for age, gender, 

race/ethnicity, education, smoking, alcohol 

consumption, BMI, hypertension, diabetes, and 

serum total cholesterol

Hyperuricemia risk by quartile, OR (95% CI)

1 referent

2 1.14 (0.78–1.67)

3 1.90 (1.35–2.69)

4 1.97 (1.44–2.70), p trend 0.0001

Hyperuricemia risk per unit increase in ln PFOA, 

OR (95% CI): 1.43 (1.16–1.76)

Logistic regression adjusting for age, gender, 

race/ethnicity, education, smoking, alcohol 

consumption, BMI, hypertension, diabetes, and 

serum total cholesterol

Chronic kidney disease defined as eGFR < 60 

mL/min/1.73 m2

Quartile, OR (95% CI)

1 referent

2 0.83 (0.55–1.24)

3 1.24 (0.75–2.05)

4 1.73 (1.04–2.88)

Logistic regression adjusting for age, gender, 

race/ethnicity, education, smoking, alcohol 

consumption, BMI, systolic blood pressure, 

diastolic blood pressure, diabetes, serum TC, and 

glycohemoglobin

Adjustment for PFOS did not alter association with 

PFOA

Multivariate regression of association PFOA with 

eGFR among subjects with and without chronic 

kidney disease

β (SE) with -1.6 (0.8) and without -2.8 (0.6) chronic 

kidney disease

Perfluorooctanoic acid (PFOA) – May 2016 3-19
Reference and Study Details PFOA Level Results

Watkins et al. 2013

West Virginia

Cross-sectional populationbased survey, residents near the 

Washington Works plant (C8 

Health Project)

9,660 (children < 18 yrs) out of 

9,783 eligible with complete 

data for serum creatinine, 

height, and serum PFOA

Median measured PFOA 

0.0283 μg/mL; range 

0.0007–2.071; yearly 

serum PFOA estimated for 

each individual from 

model used to predict 

serum PFOA at time of 

enrollment, historical 

serum PFOA during the 

first 10 years of life, 3 

years before enrollment or 

at birth

β (95% CI) change in unit eGFR (mL/min/1.73 m2 ) 

per ln serum PFOA, -0.75 (-1.41–-0.010)

Linear regression adjusting for age, gender, race, 

smoking, and household income; additional 

adjustment for regular exercise, BMI z-score, and 

TC did not alter association

No associations of predicted serum PFOA 

(modeled) with eGFR

Uric Acid (risk factor for hypertension)

Occupational exposure studies. Costa et al. (2009) examined associations between serum 

PFOA concentration and uric acid levels in serum in workers at a fluorochemical production 

plant in Italy. Serum PFOA concentration was associated with uric acid levels (p<0.05) in 56 

workers assessed concurrently over the previous 7 years. This subset of 56 workers included 

currently, formerly, and never exposed with relatively high serum PFOA concentrations. 

High-exposure community studies. Steenland et al. (2010) examined the association of serum 

PFOA concentrations with uric acid levels in adult subjects (n = 53,458; 20–≥80 years) 

participating in the C8 Health Project from 2005–2006. The reference range for uric acid is 

2.0–8.5 mg/dL. Serum samples were separated into deciles or quintiles for analysis. The data 

were analyzed by linear and logistic regression with uric acid as the outcome and PFOA as the 

exposure variable. Covariates of the model included age, gender, BMI, education, smoking, 

alcohol consumption, and serum creatinine. The mean serum PFOA concentration was 0.0864 

μg/mL. The mean uric acid level was 5.58 mg/dL with an IQR of 4.5–6.6 mg/dL. The increase in 

uric acid from lowest to highest serum PFOA concentration decile was 0.2–0.3 mg/dL. The OR

for high serum uric acid levels increased from the lowest to the highest quintile of PFOA serum 

concentrations: 1.00, 1.33 (95% CI: 1.24–1.43), 1.35 (95% CI: 1.26–1.45), 1.47 (95% CI: 1.37–

1.58), and 1.47 (95% CI: 1.37–1.58). The study showed that higher serum PFOA concentrations 

were associated with higher incidence of high serum uric acid levels. The serum of C8 study 

participants included several PFASs; PFOA appeared to have a greater influence on uric acid 

trends than PFOS in the models employed by Steenland et al. (2010). 

The C8 Science Panel (2012) combined the data from the C8 general population cohort with 

data from worker cohorts and concluded that there is no probable link between PFOA and stroke, 

hypertension, and coronary artery disease. The general population cohorts were not evaluated 

separately in these analyses.

General population studies. Shankar et al. (2011) investigated the association between serum 

PFOA (and PFOS) and uric acid concentration in the adult population of the United States by 

analyzing data from the 1999–2000, 2003–2004, and 2005–2006 NHANES evaluations. The 

study population included 3,883 adults (48.3% male) older than 20 years with data available for 

serum PFOA, plasma uric acid, and important covariates. Regression models were used to 

analyze associations with serum PFOA as a continuous variable and in quartiles. Logistic 

regression models analyzed risk for hyperuricemia defined as plasma uric acid > 6.8 mg/dL in 

males and > 6.0 mg/dL in females. Ln PFOA concentration was associated with increasing uric 

Perfluorooctanoic acid (PFOA) – May 2016 3-20
acid concentration in multivariate models adjusting for age, gender, race/ethnicity, education, 

smoking, alcohol consumption, BMI, hypertension, diabetes, and serum TC. Mean uric acid 

concentration increased by 0.22 (95% CI 0.15–0.30) mg/dL per unit change in ln PFOA. A 

concentration-response relationship was indicated across all quartiles. In addition, an elevated 

hyeruricemia risk was observed with increasing serum PFOA concentration (OR 1.43, 95% CI 

1.16–1.76). 

Glomerular Filtration Rate

High-exposure community studies. Watkins et al. (2013) evaluated the cross-sectional 

association between PFOA exposure and kidney function among children aged 1<18 years (mean 

12.4 ± 3.8 years) enrolled in the C8 Health Project. A total of 9,660 participants had data 

available on serum PFOA (median 0.0283 μg/mL), as well as serum creatinine and height, which 

were used to calculate an estimated glomerular filtration rate (eGFR). Linear regression was used 

to evaluate the association between quartiles of measured serum PFOA concentration and eGFR. 

A shift from the lowest to the highest quartile of measured, natural log-transformed 

concentrations of PFOA in serum [IQR ln (PFOA) = 1.63] was associated with a decrease in 

eGFR of 0.75 mL/min/1.73 m2 (95% CI: –1.41, –0.1; p = 0.02) adjusting for age, gender, race, 

smoking status, and household income. With increasing quartile of serum PFOA concentrations, 

eGFR decreased monotonically, although the change was slight and did not attain statistical 

significance (p for trend across quartiles = 0.30). PFOS also was associated with a decrease in 

eGFR and showed a dose-related trend. Modeled predicted serum PFOA and PFOS 

concentrations were not associated with eGFR.

General population studies. Shankar et al. (2011) also used data from the NHANES to 

determine whether there was a relationship between serum PFOA levels and chronic kidney 

disease defined as eGFR (determined from serum creatinine) of less than 60 mL/min/1.73 m2

. 

Serum PFOA levels were categorized into quartiles: Q1 = <0.0028 μg/mL; Q2 = 0.0028–0.0041 

μg/mL; Q3 = 0.0042–0.0059 μg/mL; Q4 = >0.0059 μg/mL. The adjusted OR for chronic kidney 

disease for individuals in Q4 was 1.73 (95% CI: 1.04, 2.88; p for trend = 0.015) compared with 

individuals in Q1. The logistic regression model was adjusted for age, gender, race/ethnicity, 

education, smoking, alcohol consumption, BMI, systolic blood pressure, diastolic blood pressure, 

diabetes, serum TC, and glycohemoglobin. Although a similar increase in OR was seen for 

PFOS, additional adjustment for serum PFOS did not alter the association with PFOA. In 

addition, the inverse association of eGFR with serum PFOA was observed over all quartiles of 

PFOA, as well as among individuals both with and without chronic kidney disease. Although the 

possibility of reverse causality could not be excluded, the association between serum PFOA and 

eGFR among participants without chronic kidney disease suggests a PFOA-related effect on 

kidney function. 

Overall, studies of occupational cohorts (Costa et al. 2009), a highly exposed community 

(Steenland et al. 2010; Watkins et al. 2013), and the U.S. general population (Shankar et al.

2011) that evaluated uric acid levels or eGFR as a measure of kidney function found associations 

with decreased function, although reverse causality as an explanation cannot be ruled out. Since 

the URAT transporter functions in the renal resorption of PFOA, the increase in serum uric acid 

could be a reflection of systemic transport pharmacodynamics rather than formation 

biochemistry. 

Perfluorooctanoic acid (PFOA) – May 2016 3-21
Kidney Disease

The occupational mortality studies have produced generally negative results with respect to 

the association between PFOA and mortality due to chronic kidney disease (Steenland et al. 

2015; Steenland and Woskie 2012; Raleigh et al. 2014). The most recent update of incidence of 

chronic kidney disease in the workers in the C8 West Virginia community identified 43 cases 

(with medical validation) (Steenland et al. 2015); no association was seen with cumulative 

exposure when analyzed without a lag (HRs by quartile 1.0, 0.50, 0.69, 0.67; trend p = 0.92 for 

log cumulative exposure), or using a 10-year lag (HRs by quartile 1.0, 1.32, 0.50, and 0.67; trend 

p = 0.99). 

In 2012, the C8 Science Panel concluded that there is no probable link between PFOA and 

chronic kidney disease. Their conclusion was based on findings in combined general population 

and worker cohorts, data on children enrolled in the C8 Health Project, and published data from 

NHANES.

3.1.1.5 Immunotoxicity

Immune suppression 

Immune function—specifically immune system suppression—can affect numerous health 

outcomes, including risk of common infectious diseases (e.g., colds, flu, otitis media) and some 

types of cancer. The World Health Organization (WHO) guidelines for immunotoxicity risk 

assessment recommend measures of vaccine response as a measure of immune effects, with 

potentially important public health implications (WHO 2012). 

Associations between prenatal PFOA exposure and risk of infectious diseases (as a marker of 

immune suppression) were not seen in two studies, although there was some indication of effect 

modification by gender (i.e., associations seen in females but not in males). Fei et al. (2010a) 

examined hospitalizations for infectious diseases in early childhood in a Danish birth cohort. 

Mean maternal PFOA concentration was 0.0056 μg/mL. A slightly higher risk for 

hospitalizations was observed in females with higher maternal PFOA concentrations (incidence 

rate ration [IRR] = 1.00, 1.20, 1.63, 1.74 for Q1, Q2, Q3, and Q4, respectively), and the risk for 

males was below 1.0 for each quartile. Overall, there was no association between hospitalizations 

due to infectious diseases and maternal PFOA exposure; similar results were found with PFOS. 

Okada et al. (2012) examined history of otitis media (and of allergic conditions) in children 

up to the age of 18 months. Mean maternal PFOA concentration was 0.0014 μg/mL. Cord blood 

immunoglobulin E (IgE) level decreased significantly with high maternal PFOA concentration 

among female infants, but not male infants. No significant associations were observed between 

maternal PFOA levels (and PFOS) with the incidence of otitis media (or specific types of 

allergies or wheeze). Two other studies, described below, examined reported history of colds and 

gastroenteritis in children up to age 3 years (Granum et al. 2013) or colds and flu in adults 

(Looker et al. 2014). Granum et al. (2013) observed associations between prenatal PFOA 

exposure and frequency of colds or gastroenteritis episodes, but not with a variable based on 

“ever had” this condition in the past year. Looker et al. (2014) did not observe associations 

between serum PFOA and “ever had” or frequency of colds or flu in the past year. 

In 2012, the C8 Science Panel (2012) concluded that there is no probable link between PFOA 

and common infections, including influenza, in children or adults. The panel based their 

Perfluorooctanoic acid (PFOA) – May 2016 3-22
conclusions on a subset of adult members of the cohort, a subset of mother-child pairs, and 

published data from other researchers.

Three studies have examined response to one or more vaccine (e.g., measured by antibody 

titer) in relation to higher exposure to PFOA in children (Grandjean et al. 2012; Granum et al. 

2013) or adults (Looker et al. 2014); the latter study was conducted in the high-exposure C8 

community population (Table 3-6). 

Table 3-6. Summary of Epidemiology Studies of PFOA and Immune Suppression

(Vaccine Response)

Reference and Study Details PFOA Level Results

High-Exposure Community: Adults

Looker et al. 2014

C8 Health Project, West Virginia

2005-2006 enrollment and baseline 

blood sample and questionnaires; 2010 

follow-up n = 411 with prevaccination 

blood sample – flu vaccination – 21-day 

post vaccination blood sample

Linear regression: antibody titer rise

Logistic regression: seroconversion and 

seroprotection 

Median 0.032

μg/mL

Q1: 0.0025-0.0137

Q2: 0.0138 – 0.0315

Q3: 0.0316 – 0.0903

Q4: 0.0904 – 2.14 

(Percentage positive) OR (95% CI), by influenza strain:

Seroconversion Seroprotection

(fourfold increase in antibody 

titer)

(antibody titer 1:40 following 

vaccine)

Influenza B (62%) (66%)

PFOA continuous 0.80 (0.53, 1.21) 1.04 (0.68, 1.60)

Q1 1.0 (referent) 1.0 (referent)

Q2 1.43 (0.76, 2.70) 0.76 (0.40, 1.45)

Q3 1.39 (0.73, 2.66) 1.13 (0.57, 2.23)

Q4 0.71 (0.38, 1.36) 0.77 (0.39, 1.50)

A/H1N1 (84%) (96%)

PFOA continuous (84%) (96%)

Q1 1.0 (referent) 1.0 (referent)

Q2 0.74 (0.34, 2.70) 0.74 (0.17, 3.28)

Q3 1.11 (0.73, 2.66) 1.59 (0.33, 7.70)

Q4 2.23 (0.38, 1.36) 6.47 (0.91, 45.9)

A/H3N2 (65%) (84%)

PFOA continuous 0.76 (0.51, 1.15) 0.66 (0.39, 1.12)

Q1 1.0 (referent) 1.0 (referent

Q2 0.90 (0.48, 1.68) 0.34 (0.14, 0.83)

Q3 1.13 (0.59, 2.17) 0.28 (0.11, 0.70)

Q4 0.62 (0.33, 1.66) 0.36 (0.15, 0.99)

General Population: Children

Grandjean et al. 2012

Faroe Islands

Birth cohort, follow-up to age 7 yrs

n = 587 

Age 5 prebooster (e.g., tetanus, 

diphtheria) and 4 weeks after booster 

and age 7

PFOA in 3rd trimester blood sample and 

in child (age 5)

Linear regression, adjusted for gender, 

age, birth weight, maternal smoking, 

breast-feeding, and PCBs (and time 

since booster for post-booster analysis) 

Geometric mean

Maternal sample

0.0032 μg/mL

Child’s sample

0.004 μg/mL 

Log PFOA and Log antibody Beta (95% CI) [% change in antibody 

titer per twofold increase in PFOA]

Maternal PFOA Tetanus Diphtheria

Prebooster -10.5 (-28.2, 11.7) -16.2 (-34.2, 6.7)

Postbooster 14.5 (-10.4, 46.4) -6.2 (-22.4, 13.3)

Year 7

(adjusted for age 5) 12.3 (-8.6, 38.1) -16.8 (-32.9, 3.3)

Child’s PFOA Tetanus Diphtheria

Prebooster -13.3 (-31.6, 9.9) -6.8 (-28.3, 21.0)

Postbooster - 9.7 (-30.7, 17.7) -6.1 (-23.6, 15.5)

Year 7 -28.2 (-42.7, -10.1) -23.4 (-39.3, -3.4)

(adjusted for age 5)

Similar results seen with PFOS

Granum et al. 2013

Norway

Birth cohort, Norwegian Mother and 

Child Cohort Study

n = 56 with maternal blood at delivery 

and child blood samples at 3 yrs

Linear regression, considered potential 

confounders

Mean 0.001 μg/mL Beta (95% CI ) (p-value), PFOA and antibody titer

Rubella -0.40 (-0.64, -0.17) (p = 0.001)

Measles -0.13 (-0.35, 0.09) (p = 0.24)

Tetanus 0.01 (-0.009, 0.10) (p = 0.92)

Hib -0.05 (-3.85, 3.74) (p = 0.98)

Simlar results for other PFAS

Perfluorooctanoic acid (PFOA) – May 2016 3-23
A cohort of 411 adult members of the C8 Health Project was evaluated in 2010 for an 

association between serum PFOA levels and antibody response following vaccination with an 

inactivated trivalent influenza vaccine (Looker et al. 2014). A prevaccination serum sample was 

collected at the time of vaccination and the postvaccination serum sample was collected 21 ± 3 

days later. The geometric mean serum PFOA level was 0.0337 μg/mL (95% CI 0.0298, 0.0382) 

and participants were divided into quintiles for analyses. PFOA was negatively associated with 

geometric mean A/H3N2 antibody titer rise, but no association was found with antibody titers for 

A/H1N1 and influenza type B. No association was found between antibody titers and PFOS levels.

Antibody responses to diphtheria and tetanus toxoids following childhood vaccinations were 

assessed in context of exposure to five perfluorinated compounds (Grandjean et al. 2012). The 

prospective study included a birth cohort of 587 singleton births during 1999–2001 from the 

National Hospital in the Faroe Islands. Serum antibody concentrations were measured in children 

at age 5 years prebooster, approximately 4 weeks after the booster, and at age 7 years. Prenatal

exposures to perfluorinated compounds were assessed by analysis of serum collected from the 

mother during week 32 of pregnancy (PFOA geometric mean 0.0032 μg/mL; IQR 0.00256–

0.00401); postnatal exposure was assessed from serum collected from the child at 5 years of age 

(PFOA geometric mean 0.00406 μg/mL; IQR 0.00333–0.00496). Multiple regression analyses 

with covariate adjustments were used to estimate the percent difference in specific antibody 

concentrations per twofold increase in PFOA concentration in both maternal and 5-year serum. 

Maternal PFOA serum concentration was negatively associated with antidiphtheria antibody 

concentration (-16.2%) at age 5 before booster. The biggest effect was found in comparison of 

antibody concentrations at age 7 with serum PFOA concentrations at age 5 where a twofold 

increase in PFOA was associated with differences of -36% (95% CI, -52%–-14%) and -25% 

(95% CI, -43%–-2%) for tetanus and diphtheria, respectively. Additionally at age 7, a small 

percentage of children had antibody concentrations below the clinically protective level of 

0.1 international unit (IU) /mL. The ORs of antibody concentrations falling below this level were 

4.20 (95% CI, 1.54–11.44) for tetanus and 3.27 (95% CI, 1.43–7.51) for diphtheria when age 7 

antibody levels were correlated with age 5 PFOA serum concentrations. Maternal and child 

PFOS levels also were negatively associated with antibody titers in children.

The effects of prenatal exposure to perfluorinated compounds on vaccination responses and 

clinical health outcomes in early childhood were investigated in a subcohort of the Norwegian 

Mother and Child Cohort Study (Granum et al. 2013). A total of 56 mother-child pairs, for whom 

both maternal blood samples at delivery and blood samples from the children at 3 years of age, 

were evaluated. Antibody titers specific to measles, rubella, tetanus, and influenza were 

measured as these vaccines are part of the Norwegian Childhood Vaccination Program. Serum 

IgE levels also were measured. Clinical health outcomes, including common colds and 

gastroenteritis, at ages 1, 2, and 3 years were assessed by means of a questionnaire sent to 

participants. Mean maternal plasma PFOA concentration was 0.0011 μg/mL at delivery; the 

PFOS level was 0.0056 μg/mL and PFNA and PFHxS were below the LOQ. PFOA levels in the 

children were not measured. No associations were found with PFOA or any perfluorinated 

compound and antibody levels to the vaccines with one exception. A slight, but significant, 

inverse relationship between maternal PFOS level and anti-rubella antibodies in children at 

3 years was found (β = -0.8 [95% CI -0.14, -0.02]). Maternal PFOA levels were not associated 

with adverse childhood health outcomes. 

In summary, three studies have reported decreases in response to one or more vaccines 

(e.g., measured by antibody titer) in relation to higher exposure to PFOA in children (Grandjean 

Perfluorooctanoic acid (PFOA) – May 2016 3-24
et al. 2012; Granum et al. 2013) and adults (Looker et al. 2014). In the two studies examining 

exposures in the background range (i.e., general population exposures, < 0.010 μg/ml), the 

associations with PFOA also were seen with other correlated PFASs. This limitation was not 

present in the study in adults in the high-exposure C8 community population. Serum PFOA 

levels in this study population were approximately 0.014–0.090 μg/mL. 

Asthma

The association between serum levels of perfluorinated compounds and childhood asthma 

was investigated by Dong et al. (2013). The cross-sectional study included a total of 231 children 

aged 10–15 years with physician-diagnosed asthma and 225 age-matched nonasthmatic controls. 

Between 2009 and 2010, asthmatic children were recruited from two hospitals in Northern 

Taiwan, while the controls were part of a cohort population in seven public schools in Northern 

Taiwan. Serum was collected for measurement of 10 perfluorinated compounds, absolute 

eosinophil counts, total IgE, and eosinophilic cationic protein. A questionnaire was administered 

to asthmatic children to assess asthma control and to calculate an asthma severity score 

(e.g., frequency of attacks, use of medicine, and hospitalization) during the previous 4 weeks. 

Associations of perflourinated compound quartiles with concentrations of immunological 

markers and asthma outcomes were estimated using multivariable regression models. Nine of 

10 perfluorinated compounds were detectable in ≥84.4% of all children with levels generally 

higher in asthmatic children than in nonasthmatics. Serum concentrations of PFOA in asthmatic 

and nonasthmatic children were 0.0015 ± 0.0013 μg/mL and 0.0010 ± 0.0011 μg/mL, 

respectively; four other compounds were measured at higher concentrations with the highest 

levels for PFOS and perfluorotetradecanoic acid. The adjusted ORs for asthma association with 

the highest versus lowest quartile levels were significantly elevated for seven of the compounds. 

For PFOA, the OR was 4.05 (95% CI: 2.21, 7.42). In asthmatic children, absolute eosinophil 

counts, total IgE, and eosinophilic cationic protein concentration were positively associated with 

PFOA levels with a significant monotonic trend with increasing serum concentration. None of 

these biomarkers were significantly associated with PFOA levels in nonasthmatic children. 

Serum PFOA levels were not significantly associated with asthma severity scores among the 

children with asthma, although four other compounds did show an association. 

Humblet et al. (2014) evaluated a cohort from NHANES to investigate children’s PFAS 

serum levels, including PFOA, and their association with asthma-related outcomes. Sera were 

analyzed for 12 PFASs with focus on PFOA, PFOS, PFHxS, and PFNA. A total of 1,877 

children aged 12–19 years with at least one serum sample available were included. Asthma and 

related outcomes were self-reported. Median serum PFOA levels were 0.0043 μg/mL for those 

ever having asthma and 0.0040 μg/mL for children without asthma. In the multivariable adjusted 

model, a doubling of PFOA level was associated with an increased odds of ever having asthma 

(OR=1.18, 95% CI 1.01, 1.39). PFOS was inversely associated with asthma and no associations 

were found between the other PFAS and outcome. 

On the basis of epidemiological and other data available, the C8 Science Panel (2012) found 

no probable link between PFOA and asthma in children and adults and chronic obstructive 

pulmonary disease (COPD) in adults. 

Autoimmune conditions 

The most recent report on the worker cohort initially described by Leonard et al. (2008) 

included 6,026 workers evaluated for disease incidence, not just mortality (Steenland et al. 

Perfluorooctanoic acid (PFOA) – May 2016 3-25
2015). Lifetime serum cumulative dose was estimated by combining occupational and 

nonoccupational exposures. Median measured serum level was 0.113 μg/mL based on samples 

collected in 2005. Statistically significant positive trends were found between log of cumulative 

exposure and ulcerative colitis and rheumatoid arthritis. Rate ratios for the highest quartile 

compared to the lowest quartile were 2.74 (95% CI 0.78, 9.65) for ulcerative colitis and 4.45 

(95% CI 0.99, 19.9) for rheumatoid arthritis. 

The C8 Science Panel (2012) combined these data with findings from the C8 general 

population cohort and concluded that there is a probable link between PFOA and ulcerative 

colitis. Using historical estimates for serum PFOA, the C8 Science Panel found a significant 

positive, dose-response trend with a relative risk (RR) for the highest quartile compared to the 

lowest of 3.18 (95% CI 1.84, 5.51). The panel concluded that there was no probable link between 

PFOA and autoimmune diseases, including rheumatoid arthritis, lupus, type1 diabetes, Crohn’s 

disease, or multiple sclerosis. The C8 Science Panel also concluded that there is no probable link 

between PFOA and osteoarthritis. These analyses by the panel included both worker and general 

population cohorts. 

3.1.1.6 Thyroid Effects

Several epidemiology studies have evaluated thyroid function and/or thyroid disease and its 

association with serum PFOA concentrations. Thyroid disease is more common in females than 

in males. Among the PFOA studies, the three most highly powered studies with the largest 

number of participants are one from the general U.S. population (Melzer et al. 2010) and two 

from highly exposed individuals within the C8 population (Lopez-Espinosa et al. 2012; Winquist 

and Steenland 2014b). Two of these studies are of adults (Melzer et al. 2010; Winquist and 

Steenland 2014b) and one is of children/adolescents (Lopez-Espinosa et al. 2012). 

Hypothyroidism is characterized by elevated thyroid stimulating hormone (TSH) and low T4; 

elevated TSH in conjunction with normal T4 and triiodothyronine (T3) is defined as subclinical 

hypothyroidism. Hyperthyroidisim is characterized by elevated T4 and low TSH; low levels of 

TSH in conjunction with normal T4 and T3 is defined as subclinical hyperthyroidism. Some 

studies focused on the prevalence of clinically defined disease (or the subclinical state), and 

others examined variations in TSH, T4, and T3 measurements among people who have not been 

diagnosed with a thyroid disease. Both hypothyroidism and hyperthyroidism can result from an 

autoimmune pathogenesis involving destruction of thyroid tissue. A summary of the studies on 

PFOA’s association with thyroid disease or changes in thyroid hormones follows, and is depicted 

in Table 3-7 (studies in adults) and Table 3-8 (studies in special populations—children and 

pregnant females).

Occupational exposure studies. Serum PFOA levels were obtained from volunteer workers of 

the Cottage Grove, Minnesota, PFOA plant in 1993 (n = 111) and 1995 (n = 80) as part of the 

medical surveillance program and analyzed to determine a relationship between TSH and PFOA 

concentration (Olsen et al. 1998). Employees were placed into four exposure categories based on 

their serum PFOA levels: 0–1 μg/mL, 1– < 10 μg/mL, 10– < 30 μg/mL, and >30 μg/mL. 

Statistical methods used to compare PFOA levels and hormone values included multivariable 

regression analysis, ANOVA, and Pearson correlation coefficients. TSH was significantly 

(p = 0.002) elevated in 10–<30 μg/mL exposure category for 1995 only (mean serum TSH level 

was 2.9 ppm). However, mean TSH levels for the other exposure categories, including the 

≥30 μg/mL category, were all the same (1.7 ppm). In 1993, TSH was elevated in this same 

exposure category, but was not statistically significant (p = 0.09) when compared to the other 

exposure categories.

Perfluorooctanoic acid (PFOA) – May 2016 3-26
Table 3-7. Summary of Epidemiology Studies of PFOA and Thyroid Effects in Adults

Reference and Study Details PFOA Level TSH T3 T4

Occupational Exposure Studies

Olsen and Zobel 2007

3M. Antwerp, Cottage Grove, Decatur combined; 

50-65% participation rate

n= 506

Mean age: 40 yrs

Mean duration: not reported

Linear regression adjusting for ln age, ln BMI, ln 

alcohol 

(Related references: Olsen et al. 1998, 2003)

Mean (range)

2.21 (0.01–92.03) 

μg/mL

Beta (±SE) (p-value), ln 

PFOA and ln TSH:

0.0360 (± 0.0207) (p = 0.08)

Beta (±SE) (p-value), ln 

PFOA and ln T3:

0.0105 (± 0.0053) (p = 0.05)

Beta (±SE) (p-value), ln PFOA 

and ln T4:

-0.0057 (± 0.0054) (p = 0.29)

Beta (±SE) (p-value), ln PFOA 

and ln FT4:

-0.0117 (± 0.0043) (p = 0.01)

Steenland et al. 2015

n = 3,713 workers

Data collected in 2005-2006 and 2008-2011 

n = 82 cases in men, 77 cases in women

In 2005-2006: mean 

0.325 μg/mL, median 

0.113 μg/mL

HR (95% CI), for self-reported thyroid disease, with medical record validation (incidence based on 

year of diagnosis). Cumulative exposure quartile, no lag

In men: In women:

1.0 (referent) 1.0 (referent)

1.64 (0.82, 3.29) 1.00 (0.54, 1.87)

1.13 (0.50, 2.54) 1.02 (0.48, 2.17)

2.16 (0.98, 4.77) 0.33 (0.08, 1.26)

(Ptrend = 0.98) (Ptrend = 0.97)

Adults: High-Exposure Community Studies

Emmett et al. 2006

n = 371, aged 2–89 yrs, median 50 yrs

(317 from stratified random population sample)

t-test, PFOA in abnormal vs normal TSH levels

0.354 μg/mL 6% abnormal; p-value of ttest comparing PFOA in 

abnormal and normal = 0.59

Not measured Not measured

Winquist and Steenland 2014b

n = 32,254 (including 3,713 workers)

Data collected in 2005-2006 and 2008-2011 

n = 2,008 cases in primary analysis

n = 454 cases in prospective analysis (diagnoses 

after 2005-2006) 

Stratified by gender; also conducted separate 

analyses for hyperthyroidism and hypothyroidism

In 2005-2006: mean 

0.0866 μg/mL, 

median 0.0261 

μg/mL

HR (95% CI), incident thyroid disease (with medical record validation), primary analysis:

Cumulative exposure quintiles Year exposure quintiles:

Full sample Men Women Full sample Men Women

1.0 (referent) 1.0 1.0 1.0 (referent) 1.0 1.0

1.21 1.12 1.24 1.23 1.13 1.26

1.17 0.83 1.27 1.24 1.11 1.28

1.27 1.01 1.36 1.10 1.06 1.11

1.2 1.05 1.37 1.28 1.04 1.38

(Ptrend = 0.03) (Ptrend = 0.85) (Ptrend = 0.03) (Ptrend = 0.04) (Ptrend = 0.97) (Ptrend = 0.008)

Perfluorooctanoic acid (PFOA) – May 2016 3-27
Reference and Study Details PFOA Level TSH T3 T4

Diagnoses after 2005:

Cumulative exposure quintiles Year exposure quintiles:

Full sample Men Women Full sample Men Women

1.0 (referent) 1.0 1.0 1.0 (referent) 1.0 1.0

1.23 1.35 1.23 0.80 1.32 0.74

1.00 1.37 0.93 0.91 2.09 0.76

1.06 1.44 1.00 0.93 1.83 0.82

1.12 1.85 0.96 0.91 1.76 0.80

(Ptrend = 0.73) (Ptrend = 0.09) (Ptrend = 0.55) (Ptrend = 0.86) (Ptrend = 0.54) (Ptrend = 0.53)

Adults: General Population Studies

Bloom et al. 2010

United States (New York; New York State Anglers 

Cohort Study)

n = 31 (4 women)

Mean age: 39 yrs (31–45 years)

Linear regression, adjusted

PFOA-PFOS correlation r = 0.35

Geometric mean

0.0013 μg/mL

Log-PFOA and log-TSH:

Beta = -0.06 (-0.78, 0.67)

(p = 0.87)

Not measured Log-PFOA and log-T4:

Beta = -0.01 (-0.16, 0.14)

(p = 0.89)

Shrestha et al. 2015

United States (Upper Hudson River Valley)

n = 87 (with serum for analyses); excluded if taking 

thyroid medicine

Aged: 55–74 yrs

PFOA-PFOS correlation r = 0.52

Linear regression, adjusted 

Geometric mean 

(IQR) 

0.0092 

(0.0071–0.0131) 

μg/mL

Log-PFOA and log-TSH:

Beta = 0.102 (-0.047, 0.25)

(p = 0.18)

Log-PFOA and log-T3:

Beta 3.03 (-1.73, 7.79)

(p = 0.21)

Log-PFOA and log-T4:

Beta = 0.38 (-0.07, 0.83)

(p = 0.97)

Log-PFOA and log-FT4:

Beta = 0.016 (-0.036, 0.069)

(p = 0.54)

Melzer et al. 2010

United States, NHANES 1999–2000, 2003–2004, 

and 2005–2006 

n = 3,974 adults, ages ≥ 20 yrs

Linear regression, stratified by gender, adjusted

Men (μg/mL)

Q1: 0.0001–0.0036

Q2: 0.0037–0.0052

Q3: 0.0053–0.0072

Q4: 0.0073–0.0459

Similar cut-points in 

women

Thyroid Disease, self-reported, with medication use:

Men Women

1 (referent) 1 (referent)

1.17 (0.64–2.15) 0.98 (0.65–1.50)

0.58 (0.21–1.59) 1.09 (0.66–1.81)

1.58 (0.79–3.16) 1.63 (1.07–2.47)

Wen et al. 2013

United States, NHANES 2007-2008, 2009-2010 

n = 1181, adults, aged ≥ 20 yrs

Linear regression, adjusted, with sampling weights

Mean 0.004 μg/mL Beta (95% CI) (p-value)

Ln-PFOA and ln-TSH:

Men

0004 (-0.081, 0.090)

(p = 0.92)

Women

-0.030 (-0.2157, 0.154) (p = 

0.73)

Beta (95% CI) (p-value)

Ln-PFOA and ln-T3:

Men

0.775 (-3.048, 4.598) (p = 

0.67)

Women

6.628 (0.545, 12.7) (p = 0.035)

Ln-PFOA and ln-FT3:

Men

0.016 (0.001, 0.031) (p = 0.04)

Women

0.027 (0.009, 0.044) (p = 

0.002)

Beta (95% CI) (p-value)

Ln-PFOA and ln-T4:

Men

0.000 (-0.28, 0.28)

(p = 1.0)

Women

0.082 (-0.369, 0.532) (p = 0.71)

Ln-PFOA and ln-FT4:

Men

-0.010 (-0.041, 0.022) (p = 0.53)

Women

-0.004 (-0.047,0.039)

(p = 0.83)

Perfluorooctanoic acid (PFOA) – May 2016 3-28
Table 3-8. Summary of Epidemiology Studies of PFOA and Thyroid Effects in Special Populations

Reference and Study Details PFOA Level TSH T3 T4

Children: High-Exposure Community

Lopez-Espinosa et al. 2012

n = 10,725 children, aged 1-17 yrs C8 Health Project

Reported thyroid disease based on parent- report of 

health care provider diagnosis of thyroid disease 

(and specific types); also included current use of 

thyroid medications

Subclinical disease based on hormone levels 

excluding people with self-reported thyroid disease 

or taking thyroid medication (subclinical 

hypothyroidism = above age-specific reference 

range for TSH and total T4 within reference range; 

subclinical hyperthyroidism = below age-specific 

reference range for TSH and total T4 within or 

above reference range)

Modeled in utero

PFOA: median 0.012 

μg/mL

Measured in children: 

median 0.0293 μg/mL

Beta (95% CI) for % change 

in TSH per IQR ln-PFOA:

in utero -0.5 (-2.4, 1.5)

in child 1.0 (-0.5, 2.7)

Not measured Beta (95% CI) for % change in 

total T4 per IQR ln-PFOA:

in utero -0.1 (-0.8, 0.6)

in child 0.1 (-0.5, 0.6)

OR (95% CI) for thyroid disease per IQR ln-PFOA:

in utero PFOA: Child’s PFOA:

Any thyroid disease (n = 27) 1.45 (0.95, 2.27) (n = 61) 1.44 (1.02, 2.03)

Hypothyroidism (n = 20) 1.61(0.96, 2.63) (n = 39) 1.54 (1.00, 2.37)

Subclinical hypothyroidism (n = 155) 0.94 (0.76, 1.16) (n = 365) 0.98 (0.86, 1.15)

Subclinical hyperthyroidism (n = 31) 1.10 (0.69, 1.74) (n = 78) 0.81 (0.58, 1.15)

Associations with any thyroid disease and hypothyroidism were not seen with PFOS 

Children: General Population Studies

de Cock et al. 2014

Netherlands 

n = 83 newborns

PFOA in cord blood samples

T4 in heel prick blood

Linear regression, stratified by gender; PFOA 

quartiles, adjusted 

0.943 μg/L

0.000943 μg/mL

Not measured Not measured Beta T4 (nmol/L) (95% CI)

Boys

Q1: Referent

Q2: 7.9 (-18.04, 33.92)

Q3: -2.1 (-20.94, 16.7)

Q4: 6.2 (-16.08, 28.50)

Girls

Q1: Referent 

Q2: -5.9 (-26.75, 14.94)

Q3: 11.8 (-19.08, 42.72)

Q4: 38.6 (13.34, 63.83)

Lin et al. 2013

Taiwan, Young Taiwanese Cardiovascular Cohort 

Study

n = 545 (45 with elevated blood pressure); n = 18 

hypothyroid- TSH > normal range

Aged: 12-30 yrs

Geometric mean 

0.00267 μg/mL

Q1: <0.00364

Q2: 0.00364 – 

≤0.00066

Q3: 0.00666 - 

≤0.00971

Q4: >0.009.71

Mean (±SE) Ln TSH by 

PFOA quartile (adjusted)

Q1: 0.48 (± 0.08)

Q2: 0.45 (± 0.09)

Q3: 0.36 (± 0.11)

Q4: 0.41 (± 0.12)

No association with risk of 

hypothyroidism

Not measured Mean (±SE) free T4 by PFOA 

quartile (adjusted)

Q1: 1.07 (± 0.01)

Q2: 1.08 (± 0.02)

Q3: 1.10 (± 0.02)

Q4: 1.06 (± 0.02)

Perfluorooctanoic acid (PFOA) – May 2016 3-29
Reference and Study Details PFOA Level TSH T3 T4

Pregnant Women: General Population Studies

Chan et al. 2011

Canada 

n = 96 hypothyroxinemia cases (normal TSH with 

decreased free T4 – below 10th percentile) and 175 

controls (normal TSK and free T4 in 50th–90th 

percentile; matching based on referring physician 

and maternal age

2nd trimester blood sample (mean 18 weeks)

Conditional logistic regression, adjusted

PFOA-PFOS correlation r = 0.5

Geometric mean 

0.00135 μg/mL

Ln PFOA OR (95% CI):

0.94 (0.74–1.18) 

With additional adjustment for PFOS and PFHxS:

0.87 (0.63–1.19) 

Wang et al. 2013

Norway (from case-control study of subfecundity in 

the Norwegian Mother and Child Cohort Study; 

cases and controls combined)

n = 903 women

2nd trimester blood sample (mean 18 weeks)

Linear regression, adjusted

Median 0.00215 

μg/mL

PFOA and ln-TSH

Beta (95% CI)

−0.0001 (−0.045, 0.044)

Not measured Not measured

Berg et al. 2015

Norway, Northern Norway Mother and Child 

Contaminant Cohort Study

n = 375

2nd trimester blood sample (18 weeks)

Thyroid hormones and anti-TPO antibodies 

measured at 18 weeks gestation and at day 3 and 

week 6 after delivery

Mixed effects linear models

Repeated measures of thyroid hormone levels were 

used in model

PFOA-PFOS correlation r = 0.65

Median 0.00153 

μg/mL

Highest quartile PFOA 

associated with higher TSH, 

but not significant when 

adjusted for PFOS 

(quantitative results not 

reported) 

Quantitative results not 

reported (noted as no 

association) 

Quantitative results not reported 

(noted as no association)

Webster et al. 2014

Canada (Vancouver Chemicals Health and 

Pregnancy Study)

n = 152, not taking thyroid medicine

2nd trimester blood samples (15 and 18 weeks)

Mixed effects linear models, stratified by TPO 

antibody levels

PFOA-PFOS correlation r = 0.71

Median 0.0017

μg/mL

Beta per IQR PFOA and TSH, 

(95% CI) (p-value)

Normal TPO antibody 

 0.07 (-0.1, 0.2) (p = 0.41)

High TPO antibody

0.7 (0.09, 1) (p = 0.02)

Similar results for PFOS

[IQR PFOA = 0.0014 μg/mL]

not measured Beta per IQR PFOA and FT4, 

(95% CI) (p-value)

Normal TPO antibody

 -0.03 (-0.3, 0.2) (p = 0.82)

High TPO antibody

 -0.4 (-1, 0.5) (p = 0.35)

[IQR PFOA = 0.0014 μg/mL]

Perfluorooctanoic acid (PFOA) – May 2016 3-30
In an expanded and refined analysis of the data reported in Olsen et al. 2003, Olsen and 

Zobel (2007) looked at the relationship between serum PFOA concentration and TSH, serum and 

free T4, and T3 levels in workers at the Decatur, Antwerp, and Cottage Grove production plants. 

The fluorochemical workers consisted of males (aged 21–67) from the Antwerp, Belgium 

(n = 196); Cottage Grove, Minnesota (n = 122); and Decatur, Alabama (n = 188) production 

facilities who volunteered to participate in the medical surveillance program in 2000. The mean 

serum PFOA concentration was 2.21 μg/mL for all sites combined. No association between 

TSH, serum T4, and PFOA concentration was observed. A negative association (p<0.01) 

between free T4 and serum PFOA concentration was observed in the unadjusted and adjusted 

(age, BMI, and alcohol consumption) models for all locations combined; no association was 

observed for the individual locations. A positive association (p<0.05) was observed between T3 

and serum PFOA concentration in the unadjusted and adjusted models for all locations 

combined, the Antwerp plant, and the Decatur plant. The authors noted that the results were not 

considered clinically relevant because the results were within normal reference range. Steenland 

et al. (2015) did not find an association between self-reported thyroid disease and PFOA levels 

among 3,713 workers at the Washington Works plant in West Virginia who participated in the 

C8 Health Project.

Two studies measured thyroid hormones in PFOA-exposed workers, but did not present an 

analysis of the relation between PFOA exposure and hormone levels. Both studies noted that the 

thyroid hormone values were in the normal range (Costa et al. 2009; Sakr et al. 2007a). 

High-exposure community studies. Emmett et al. (2006) examined the association of serum 

PFOA with thyroid disease in 371 residents of the Little Hocking, Ohio, water district as 

described previously. No association was observed between serum PFOA and thyroid disease. 

Serum PFOA was decreased (not significantly different) in subjects with self-reported disease 

(e.g., hyperthyroidism, goiter or enlarged thyroid, hypothyroidism) (0.387 μg/mL; n = 40) 

compared to subjects without thyroid disease (0.451 μg/mL; n = 331). No association was seen 

between serum PFOA and TSH when analyzed with linear regression or by t-test comparison of 

PFOA in the abnormal TSH (n = 24, 6%) and normal TSH groups (p = 0.59). 

Participants in the C8 Health Project were examined for an association between PFOA levels 

and thyroid disease (Winquist and Steenland 2014b). The cohort included 28,541 community 

members and 3,713 workers who had completed study questionnaires during 2008–2011. The 

median serum PFOA level at enrollment in 2005–2006 was 0.0261 μg/mL for the combined 

cohort, 0.0242 μg/mL for the community members, and 0.1127 μg/mL for the workers. 

Retrospective serum levels for the community cohort, estimated from air and water 

concentrations, residential history, and water consumption rates, were used to estimate yearly 

intakes. For the workers, yearly serum estimates were modeled from work history information 

and job-specific concentrations. Cox proportional hazard models, stratified by birth year, were 

used to assess self-reported adult thyroid disease hazard in relation to time-varying yearly or 

cumulative (sum of yearly estimates) estimated PFOA serum concentration, controlling for 

gender, race, education, smoking, and alcohol consumption. For the combined cohort, quintiles 

for yearly exposure were 0.00011–<0.0047, 0.0047–<0.00849, 0.00849–<0.0216, 0.0216–

<0.100, and 0.100–3.303 μg/mL; quintiles for cumulative exposure were 0.0001–<0.115, 0.115–

<0.202, 0.202–<0.497, 0.497–2.676, and 2.676–97.396 μg/mL·year. As expected, the number of 

thyroid disease cases was higher among females than among males. Positive associations were 

seen with the cumulative exposure and the per-year exposure metrics for incidence of all thyroid 

disease (as well as for specific subtypes), with the observations seen primarily in females

Perfluorooctanoic acid (PFOA) – May 2016 3-31
(Table 3-7). When limited to disease occurring after the 2005–2006 serum collection, the number 

of incident cases was reduced from 2,008 to 454, and the patterns of associations were more 

variable. No associations between estimated serum PFOA level and thyroid disease were found

in the analysis limited to workers in this study population (Steenland et al. 2015). 

The C8 Science Panel (2012) used data from the C8 general population cohort and concluded 

that there is a probable link between PFOA and thyroid disease. 

General population studies. Bloom et al. (2010) investigated the associations between serum 

PFAS, including PFOA, and TSH and free thyroxine (FT4). The serum samples came from 31 

participants (27 males, 4 females; mean age 39 years) of the 1995–1997 New York State Angler 

Cohort Study Dioxin Exposure Substudy. The study subjects each completed a questionnaire and 

provided a blood sample for serum analysis. The questionnaire contained questions about sportfish and game consumption, lifestyle, demographic factors, and medical history. The serum 

samples were analyzed for TSH and FT4 in 2003 by immunometric chemiluminescent sandwich 

assay and for PFAS in 2006 by ion pair extraction high-performance LC-MS/MS. Regression 

models were used to analyze the data and adjust for confounders. No subjects reported use of 

thyroid medication or physician-diagnosed goiter or thyroid conditions. Mean TSH concentration 

(range 0.43–15.70 μIU/mL) was within normal range (0.40–5.00 μIU/mL) with the exception of 

one subject. Mean FT4 (0.90–1.55 ng/dL) was within normal range (0.80–1.80 ng/dL) for all 

subjects. The mean serum PFOA concentration was 0.00133 μg/mL and ranged from 0.00057 to 

0.00258 μg/mL. The males had a significantly higher serum PFOA concentration than the 

females (0.00147 μg/mL versus 0.00105 μg/mL; p = 0.047). There was no association between 

serum PFOA concentration (or PFOS) and TSH or FT4. 

The relationship between serum levels of PFOA, PFOS and other persistent organic 

pollutants and thyroid biomarkers was investigated in older adults (Shrestha et al. 2015). Levels 

of TSH, FT4, T4, and T3 were measured in 51 males and 36 females with a mean age of 63.6 

years. None of the participants had thyroid disease or were taking thyroid medication. Covariates 

in the analysis included age, gender, education level, the sum of polychlorinated biphenyls 

(∑PCBs) and polybrominated diphenyl ethers (∑PBDEs), smoking status, and alcohol 

consumption. The mean PFOA serum level was 0.0104 ± 0.0057 μg/mL for all participants. In 

both unadjusted and adjusted models, PFOA was significantly (p<0.05 or 0.01) and positively 

associated with T4 and T3; a possible dose-response was not evaluated in this small sample. A 

statistical interaction was detected between age and PFOA for effects on FT4 and T4 suggesting 

that the positive associations of PFOA were potentiated by age. PFOS was also positively 

associated with FT4 and T4.

Melzer et al. (2010) examined the association between serum PFOA concentration and 

thyroid disease in the general population of the United States by analyzing data from the 1999–

2000, 2003–2004, and 2005–2006 NHANES The population included 3,966 adults (2,066 

females, 1,900 males) older than 18 years. Each of the participants answered a questionnaire, had 

a physical examination, and provided blood and urine samples for analysis. Serum samples were 

analyzed for PFOA concentration by solid-phase extraction coupled to isotope dilution/highperformance LC-MS/MS. Data on diseases diagnosed by a physician and confounding factors, 

including year of NHANES, age, gender, race/ethnicity, education, smoking status, BMI, and 

alcohol consumption were obtained from the questionnaire. Individuals were considered to have 

thyroid disease if they responded on the questionnaire to having a physician-diagnosed disease or 

if they were taking medication for either hypothyroidism or hyperthyroidism. 

Perfluorooctanoic acid (PFOA) – May 2016 3-32
Regression models were used to analyze the data and adjust for confounders. Serum PFOA 

concentration was divided into quartiles for each gender. In females, serum PFOA concentration 

ranged from 0.0001–0.123 μg/mL (Q1 = 0.0001–0.0026; Q2 = 0.0027–0.004; Q3 = 0.0041–

0.0057; Q4 = 0.0057–0.123), and in males, serum PFOA concentration ranged from 0.0001–

0.0459 μg/mL (Q1 = 0.0001–0.0036; Q2 = 0.0037–0.0052; Q3 = 0.0053–0.0072; Q4 = 0.0073–

0.0459). Females in PFOA Q4 were more likely to report current thyroid disease [OR = 2.24, 

95% CI: 1.38–3.65, p = 0.002] compared to females in Q1 and Q2. No association between 

serum PFOA concentration and thyroid disease was observed in males. With PFOS, the opposite 

was found, with males in the highest quartile, but not females, more likely to report thyroid 

disease. Data interpretation was limited by the cross-sectional study design, lack of information 

on the specific thyroid disorder diagnosis in the questionnaire responses, and single serum 

samples for PFOA measurements taken at the same time disease status was ascertained through 

the questionnaire. Thus, the possibility of reverse causality cannot be eliminated. 

Another study of 1,181 members of NHANES for survey years 2007–2008 and 2009–2010 

examined the association between serum PFOA levels (and 12 other PFASs) and thyroid 

hormone levels (Wen et al. 2013). Multivariable linear regression models were used with serum 

thyroid measures as the dependent variable and individual natural log-transformed PFAS

concentration as a predictor along with confounders. The geometric mean serum PFOA level was 

0.00415 μg/mL. A positive association between PFOA level and free T3 (FT3) was found in 

females as a 1-unit increase in natural log-serum PFOA increased serum total T3 concentration 

by 6.628 ng/dL (95% CI 0.545, 12.712, p = .035). However, the association was no longer 

significant when PFOS, PFNA, and PFHxS levels were included in the model. 

A different type of examination was undertaken by Pirali et al. (2009). The study measured 

intrathyroidal levels of PFOA (and PFOS) in thyroid surgical specimens to determine if a 

relationship existed between PFOA and the clinical, biochemical, and histological phenotype of 

thyroid disease patients. Serum PFOA concentration also was measured to determine if a 

relationship existed between thyroid tissue and serum PFOA levels. Patients (n = 28; 8 males, 

20 females; 33–79 years) with benign multinodular goiters (n = 15), Graves’ disease (n = 7), 

malignant papillary carcinoma (n = 5), and malignant follicular carcinoma (n = 1) were included 

in the study. Informed consent, clinical examination, work history, thyroid hormone and 

antibody measurements, thyroid ultrasound, fine-needle aspiration of nodules greater than 1 cm, 

and serum samples (n = 21) were performed or collected prior to surgery. The control group 

consisted of thyroid tissues collected at autopsy from subjects with no history of thyroid disease 

(n = 7; 5 males, 3 females; 12–83 years) and serum samples from 10 subjects with no evidence 

of thyroid disease. The student’s t-test, Mann-Whitney U-test, Pearson and Spearman’s 

correlation tests, and chi-square test with Fisher’s correction were used to compare group results. 

Regression analysis was used to test the effect of different variables independently of a covariate.

The median concentration of PFOA in thyroid tissue was 2.0 ng/g (range = 0.4–4.6 ng/g). 

The patients were divided into three different groups: group I (toxic and nontoxic multinodular 

goiter, n = 12), group II (differentiated thyroid cancer, n = 6), and group III (Hashimoto’s 

thyroiditis or Graves’ disease, n = 10). Thyroid PFOA concentration for the control group, group 

I, group II, and group III ranged from 1.0–6.0, 0.4–4.4, 1.4–4.0, and 1.0–4.6 ng/g, respectively. 

Serum PFOA concentration for the control group, group I, group II, and group III ranged from 

0.004–0.0137, 0.0012–0.0166, 0.0051–0.0096, and 0.0039–0.0125 μg/ml, respectively. The 

concentration of PFOA in the thyroid and serum was similar between control and thyroid 

patients at the time of measurement. Age, gender, residence, working activity, malignant / 

Perfluorooctanoic acid (PFOA) – May 2016 3-33
nonmalignant conditions, antibodies, thyroid hormone concentrations, and ultrasound parameters 

were not associated with thyroid or serum PFOA concentration. There also was no correlation 

between serum and thyroid PFOA concentration. Similar results were obtained with PFOS.

Children. Three studies evaluated thyroid function in children (or children and young adults) 

(Table 3-8). In the children from the C8 cohort who were highly exposed to PFOA, LopezEspinosa et al. (2012) observed positive associations between prenatal PFOA (modeled maternal 

levels) and any thyroid disease or clinical hypothyroidism; similar results were seen with the 

child’s PFOA level. Associations were not seen with subclinical hypothyroidism or 

hyperthyroidism, or TSH or total T4 levels among children without thyroid disease. In a study 

from the Netherlands of 52 males and 31 females, increasing T4 levels in females were 

associated with increasing prenatal PFOA concentrations (as measured in cord blood samples) 

(de Cock et al. 2014); no associations were reported in males. A study of adolescents and young 

adults (aged 12–30 years) from Taiwan did not observe associations between serum PFOA 

concentrations and TSH or T4 levels (Lin et al. 2013). 

Pregnant females. Several studies of thyroid have been conducted in pregnant females (Table 

3-8), mostly reporting null associations between maternal PFOA concentration and thyroid status 

during pregnancy (Berg et al. 2015; Chan et al. 2011; Wang et al. 2013). The exception to these 

results is the only study that included an analysis stratified by presence of antithyroid peroxidise 

(anti-TPO) antibodies (Webster et al. 2014), in which associations between PFOA and TSH were 

seen only among females with high autoantibody levels. This finding supports the importance of 

further research into the association between PFOA and autoimmunity and autoimmune 

conditions. 

Chan et al. (2011) examined the association between hypothyroxinemia and serum PFOA 

concentration (and PFOS) in pregnant Canadian females (n = 271; 20.1–45.1 years of age, 

≥22 weeks of gestation) in a matched case-control study. Maternal serum from the second 

trimester was collected between December 15, 2005, and June 22, 2006, as part of an elective 

prenatal screen for birth defects. Serum samples were analyzed for TSH and FT4 concentrations 

and PFOA. The cases of hypothyroxinemia (n = 96) had normal TSH concentrations and FT4

concentrations in the lowest 10th percentile (≤8.8 pmol/L). The controls (n = 175) had normal 

TSH concentrations and FT4 concentrations between the 50th and 90th percentiles (12–14.1 

pmol/L). Maternal age, weight, and gestational age at blood draw and dichotomized at 50th

percentiles were included as confounders, and race was included as a covariate. Chi-square tests 

and regression models were used to analyze the data. Overall, the geometric mean PFOA level 

was 0.00135 μg/mL. Statistical comparisons used the geometric mean serum PFOA 

concentration in the cases of 3.10 nmol/L and 3.32 nmol/L in the controls. There was no 

association between serum PFOA concentration (or PFOS) and hypothyroxinemia in pregnant 

females. 

A cross-sectional study of 903 pregnant females evaluated the association between plasma 

PFOA levels and plasma TSH (Wang et al. 2013). Twelve other PFASs also were quantified and 

evaluated. The females were a cohort of the Norwegian Mother and Child Cohort Study and the 

blood samples were drawn at approximately week 18 of gestation. The median PFOA 

concentration was 0.0022 μg/mL with an interquartile range of 0.00157–0.00295 μg/mL. No 

association was found between plasma levels of PFOA and TSH. PFOS was associated with 

higher TSH levels, but plasma levels of other PFASs were unrelated to TSH.

Perfluorooctanoic acid (PFOA) – May 2016 3-34
Expanding on the above study, Berg et al. (2015) investigated the association between a 

number of PFASs, including PFOA, and TSH, T3, T4, FT3, and FT4. A subset of 375 females in 

the Norwegian Mother and Child Cohort Study with blood samples at about gestational week 18 

and at 3 days and 6 weeks after delivery were included. Seven compounds were detected in more 

than 80% of the blood samples, with PFOS present in the highest concentration followed by 

PFOA. The median PFOA level was 0.00153 μg/mL, and the females were assigned to quartiles 

based on the first blood sample at week 18 of gestation. Females in the highest quartile had 

significantly higher mean TSH than females in the first quartile; however, when PFOS 

concentration was included as a covariate, the association was not significant.

A study of Canadian females (n = 152) evaluated maternal serum PFOA levels (and PFHxS, 

PFNA, and PFOS) for associations with thyroid hormone levels during the early second trimester 

of pregnancy, weeks 15–18 (Webster et al. 2014). Mixed effects linear models were used to 

examine associations between PFOA levels and FT4, total T4, and TSH; associations were made 

for all females and separately for females with high levels of TPO antibody, a marker of 

autoimmune hypothyroidism. Median serum PFOA was 0.0017 μg/mL. No associations were 

found between levels of PFOA (or PFOS and PFHxS), and thyroid hormone levels in females

with normal antibody levels. PFNA was positively associated with TSH. Clinically elevated TPO

antibody levels were found in 14 (9%) of the study population. In the females with high antibody 

levels, PFOA, as well as PFNA and PFOS, was strongly and positively associated with TSH. An 

IQR increase in maternal PFOA concentrations was associated with a 54% increase in maternal 

TSH compared to the median TSH level. PFNA and PFOS concentrations were associated with 

46% and 69% increases, respectively, in maternal TSH.

As illustrated above, numerous epidemiology studies have evaluated thyroid function and/or 

thyroid disease in association with serum PFOA concentrations (Tables 3-7 and 3-8). As noted 

previously, thyroid disease is more common in females. Several studies provide support for an 

association between PFOA exposure and incidence or prevalence of thyroid disease, and include 

large studies of representative samples of the general U.S. population (Melzer et al. 2010) and 

the high-exposure C8 community population (Lopez-Espinosa et al. 2012; Winquist and 

Steenland 2014b). Two of these studies are of adults (Melzer et al. 2010; Winquist and Steenland 

2014b) and one is of children/adolescents (Lopez-Espinosa et al. 2012). The trend for an 

association with thyroid disease was seen in females in the C8 population (Winquist and 

Steenland 2014b) and the general population (Melzer et al. 2010), and in children (LopezEspinosa et al. 2012); this was most often hypothyroidism. Association between PFOA and TSH 

also was seen in pregnant females with anti-TPO antibodies (Webster et al. 2014). In contrast, 

generally null associations were found between PFOA and TSH or thyroid hormones (T3 or T4) 

in people who have not been diagnosed with thyroid disease. 

3.1.1.7 Diabetes and Related Endpoints

Occupational exposure studies. Leonard et al. (2008) examined cause of death among former 

workers at the DuPont Washington Works plant in West Virginia. The cohort consisted of 6,027 

employees (4,872 males and 1,155 females) who had worked at the plant from 1948 through 

2002. The DuPont Epidemiology Registry and U.S. National Death Index were used to obtain 

causes of death. SMRs were estimated using three reference populations; the populations of the 

United States and West Virginia and the DuPont regional worker reference population excluding 

workers at the Washington Works plant. A significant increase in diabetes mortality was 

observed for Washington Works plant workers compared to the DuPont regional worker 

Perfluorooctanoic acid (PFOA) – May 2016 3-35
reference population [SMR = 197, 95% CI: 123, 298]. However, no regression analyses were 

done with PFOA levels. 

The Leonard et al. study (2008) was updated in a cohort mortality study conducted by 

Steenland and Woskie (2012) to include 5,791 individuals who had worked at the DuPont West 

Virginia plant for at least 1 year between 1948 and 2002. Mean duration of employment was 19 

years. Deaths through 2008 were ascertained through either the National Death Index or death 

certificate data. Exposure quartiles were assessed by estimated cumulative annual serum levels 

based on blood samples from 1,308 workers taken during 1979–2004 and time spent in various 

job categories (ppm-years). Referent groups included both nonexposed DuPont workers in the 

same region and the U.S. population. Overall, the mean cumulative exposure was 7.8 ppm-years 

and the estimated average annual serum level was 0.35 μg/mL. Compared to the referent rates 

from other DuPont workers, cause-specific mortality rates were elevated for diabetes (n = 38; 

SMR=1.90; 95% CI 1.35, 2.61). These data are limited by the small number of cases and the 

restriction to mortality as an outcome.

The most recent report on the above cohort included 6,026 workers evaluated for disease 

incidence, not just mortality (Steenland et al. 2015). Lifetime serum cumulative dose was 

estimated by combining occupational and nonoccupational exposures. Median measured serum 

level was 0.113 μg/mL based on samples collected in 2005. No association was found between 

PFOA level and type II diabetes incidence rate.

High-exposure community studies. MacNeil et al. (2009) examined the association of PFOA 

with type II diabetes in adult participants of the C8 Health Project (n = 54,468; age 20 to >80 

years). Serum PFOA concentration was divided into deciles using the population distribution. 

Other PFAS were not evaluated in this study. Serum PFOA (deciles), BMI, gender, family 

history of diabetes, race, use of cholesterol-lowering medicine, and use of blood pressurelowering medicine were used to analyze the data in categorical and logistic regression models for 

the outcome of type II diabetes. Serum fasting glucose levels were the focus for a linear 

regression analysis of the study population (n = 21,643) excluding type II diabetics and those 

who had provided nonfasting blood samples. The mean serum PFOA concentration for the entire 

study population was 0.0868 μg/mL and 0.0913 μg/mL for subjects with type II diabetes 

validated by medical review (n = 3,539). 

There was no association between serum PFOA concentration and fasting serum glucose 

level in subjects characterized as nondiabetic. The mean serum PFOA concentration was 

0.0929 μg/mL in subjects who self-reported type II diabetes (n = 4,278) and 0.1227 μg/mL in 

subjects diagnosed in the last 10 years (n = 1,055). No association was observed between type II 

diabetes and serum PFOA concentration. The OR by decile was 1.00, 0.71, 0.60, 0.72, 0.65, 

0.65, 0.87, 0.58, 0.62, and 0.72. The results of the analysis indicated that PFOA exposure is not 

associated with type II diabetes among the population studied. Data interpretation was limited by 

the cross-sectional study design, which made it difficult to determine if PFOA exposure preceded 

disease.

The C8 Science Panel (2012) combined these data from the C8 general population cohort 

with follow-up data and data from worker cohorts, and concluded that there is no probable link 

between PFOA and type II diabetes.

General population studies. Preconception serum levels of PFOA (and other PFASs) were 

evaluated in females attempting pregnancy in relation to risk of developing gestational diabetes 

Perfluorooctanoic acid (PFOA) – May 2016 3-36
(Zhang et al. 2015). The 258 participants were members of the Longitudinal Investigation of 

Fertility and the Environment (LIFE) study with blood samples taken during 2005–2009. The 

ORs and 95% CIs of gestational diabetes associated with each SD increment of preconception 

serum PFOA concentration (log-transformed) (and six other PFASs) were estimated with the use 

of logistic regression after adjusting for confounders. Preconception mean serum PFOA levels 

were 0.0033 μg/mL for the entire cohort, 0.00394 μg/mL in females with gestational diabetes 

and 0.00307 μg/mL in females without gestational diabetes. A significant positive association 

was found between PFOA and risk of gestational diabetes in the fully adjusted model (OR=1.86; 

95% CI 1.14, 3.02). Associations for six other PFAS were slightly increased (e.g., PFOS 

OR=1.13), but did not attain statistical significance. 

Metabolic syndrome is a combination of medical disorders and risk factors that increase the 

risk of developing cardiovascular disease and diabetes. Lin et al. (2009) investigated the 

association between serum PFOA (plus three other PFASs) and glucose homeostasis and 

metabolic syndrome in adolescents (aged 12–20 years) and adults (aged >20 years) by analyzing 

the 1999–2000 and 2003–2004 NHANES data. The National Cholesterol Education Program 

Adult Treatment Panel III guidelines were used to define adult metabolic syndrome and the 

modified guidelines were used to define adolescent metabolic syndrome. The study population 

included 1,443 subjects (474 adolescents, 969 adults) at least 12 years of age who had a morning 

examination and triglyceride measurement. There were 266 male and 208 female adolescents 

and 475 male and 493 female adults. Multiple linear regression and logistic regression models 

were used to analyze the data. Covariates included age, gender, race, smoking status, alcohol 

intake, and household income. Log-transformed PFOA concentration was 1.51 and 1.48 ng/mL 

for adolescents and adults, respectively. In adults, serum PFOA concentration was associated 

with increased β-cell function (β coefficient 0.07, p<0.05). Serum PFOA concentration was not 

associated with metabolic syndrome, metabolic syndrome waist circumference, glucose 

concentration, homeostasis model of insulin resistance, or insulin levels in adults or adolescents. 

Both PFOS and PFNA were positively associated with some of the endpoints associated with 

metabolic syndrome. 

Nelson et al. (2010) examined the relationship between polyfluoroalkyl chemical serum 

concentration, including PFOA, and insulin resistance as previously described for data from 

NHANES. Fasting insulin and fasting glucose were used to determine the homeostatic model 

assessment for insulin resistance. No association was found between serum PFOA concentration, 

or any other PFAS, and insulin resistance. 

Overall, these studies show a lack of association of PFOA with diabetes, metabolic 

syndrome, and related endpoints. 

3.1.1.8 Reproductive and Developmental Endpoints

Several studies have examined the relationship between PFOA exposures and reproductive, 

gestational, and developmental endpoints as well as postnatal growth and maturation in humans. 

Pregnancy-related endpoints include gestational age (Nolan et al. 2009), measures of fetal 

growth (Apelberg et al. 2007; Fei et al. 2007, 2008a; Monroy et al. 2008; Nolan et al. 2009; Stein 

et al. 2009; Washino et al. 2009), miscarriage or preterm birth (Stein et al. 2009), birth defects 

(Stein et al. 2009), hypertension and preeclampsia (Darrow et al. 2013; Savitz et al. 2012a, 

2012b; Stein et al. 2009), and fecundity (Fei et al. 2009; Vélez et al. 2015). Infant growth and 

development during the first 7 years (Andersen et al. 2010, 2013; Fei et al. 2008b, 2010a, 2010b; 

Høyer et al. 2015b) and postnatal growth and maturation, including neurodevelopment (Fei and 

Perfluorooctanoic acid (PFOA) – May 2016 3-37
Olsen 2011; Hoffman et al. 2010; Høyer et al. 2015a; Liew et al. 2014; Stein et al. 2013) and risk 

of adult obesity (Halldorsson et al. 2012) also have been studied. Male reproductive endpoints 

evaluated in humans include sperm count and semen quality (Buck Louis et al. 2015; Joensen et 

al. 2009, 2013; Vested et al. 2013). Female pubertal development was examined in three studies 

(Christensen et al. 2011; Kristensen et al. 2013; Lopez-Espinosa et al. 2011). As noted 

previously, the focus of this review is on pregnancy-related outcomes, specifically pregnancyrelated hypertension and preeclampsia, measures of fetal growth, and pubertal development. 

Within each section, the discussion is divided into occupational exposure studies (if applicable), 

the C8 high-exposure community studies, and general population studies. 

Several analyses are based on the Danish National Birth Cohort (Andersen et al. 2010, 2013; 

Fei et al. 2007, 2008a, 2008b, 2009, 2010a, 2010b; Fei and Olsen 2011). The females (n = 1,400) 

and their infants were randomly selected, and the study included those who provided their first 

blood samples between gestational weeks 4 and 14 and gave birth to a single live-born child 

without congenital malformation. The females participated in telephone interviews—at 12 and 

30 weeks gestation, when the children were 6 and 18 months of age, and when the children were 

7 years of age—and filled out a food frequency questionnaire. As the children aged, more 

questionnaires were completed by the mothers with regard to behavioral health and motor 

coordination. Highly structured questionnaires were used to gather information on possible 

confounders, including infant gender, maternal age, parity, socio-occupational status, 

prepregnancy BMI, and smoking during pregnancy. The National Hospital Discharge Register 

was used to obtain birth weight, gestational age, placental weight, birth length, head and 

abdominal circumference data, Apgar scores based on heart rate, respiratory effort, reflex, 

irritability, muscle tone, and skin color. Plasma PFOA concentration was determined from the 

first blood samples of 1,399 females, from the second blood samples of 200 females, and from 

cord blood samples of 50 infants by solid-phase extraction high-performance LC-MS/MS. PFOA 

concentrations were divided into quartiles (Fei et al. 2009, 2010b), with the lowest quartile 

designated as the reference group, as follows: <lower limit of quantification– (LLOQ–) 0.00390, 

0.00391–0.00520, 0.00521–0.00696, and ≥ 0.00697 μg/mL. Regression models were used to 

analyze the data. Results of these studies are included in the following discussion of results for 

specific endpoints. 

Pregnancy-related hypertension and preeclampsia. There are no occupational exposure and 

general population studies examining pregnancy-related hypertension and preeclampsia in 

relation to PFOA exposure. The only data available come from the high-exposure C8 Health 

Project study population (Table 3-9). 

Several studies, using different designs and exposure measures, have examined birth 

outcomes, including pregnancy-induced hypertension or preeclampsia in infants born to mothers 

in the high-exposure C8 community population in West Virginia and Ohio (information obtained 

from questionnaire-based pregnancy histories or from linkage to birth records) (Table 3-9). Stein 

et al. (2009) used an exposure measure based on individual serum PFOA levels obtained in the 

2005–2006 baseline survey to examine birth outcomes (based on self-report) in 1,845 births in 

the 5 years preceding the PFOA measurement. Savitz et al. (2012a, 2012b) included births from 

1990 to 2004, modeling exposure based on the serum measurements in 2005, information 

obtained in the 2005 baseline questionnaire regarding residential history, information on 

historical environmental releases, and PKs. In one of the analyses (study II in Savitz et al. 

2012b), linkage with birth records was used to verify the preeclampsia outcome. 

Perfluorooctanoic acid (PFOA) – May 2016 3-38
Table 3-9. Summary of Epidemiology Studies of PFOA and Pregnancy-Induced 

Hypertension or Preeclampsia

Study PFOA Level Results

Stein et al. 2009

United States (C8 Health Project)

n = 1,845 pregnancies in the 5 years 

before enrollment

Exposure based on serum collected in 

2005

Outcome based on pregnancy history 

collected in 2005

Median 0.0212 

μg/mL

OR (95% CI), preeclampsia

 per IQR (lnPFOA) increase in PFOA:

 1.1 (0.9, 1.3) [IQR(lnPFOA)=0.0395 μg/mL]

< 50th percentile 1.0 (referent)

≥ 50th 1.3 (0.9, 1.9)

< 50th percentile 1.0 (referent)

50-75th 1.5 (1.0, 2.3)

75-90th 1.2 (0.7, 2.1)

> 90th 0.9 (0.5, 1.8)

Savitz et al. 2012a

United States (C8 Health Project)

n = 11,737 pregnancies in 1990-2006

Modeled exposure based on serum 

collected in 2005, residential history, and 

other data

Outcome based on pregnancy history 

collected in 2005

OR (95% CI), preeclampsia

 per IQR(lnPFOA) increase in PFOA:

 1.13 (1.00–1.28) [IQR (lnPFOA)=0.00219 μg/mL]

 per 0.100 μg/mL increase in PFOA:

 1.08 (1.01–1.15)

< 40th percentile 1.0 (referent)

40-60th 1.2 (1.0, 1.5)

60-80th 1.1 (0.9, 1.4)

≥ 80th 1.2 (1.0, 1.6)

Also noted stronger associations with Bayesian 

calibration of exposure and among women with 

highest quality residential history

Savitz et al. 2012b

United States (C8 Health Project)

n = 4,547 pregnancies in 1990-2004

Modeled exposure based on serum 

collected in 2005, residential history, and 

other data

Outcome based on linkage to birth 

records

Median 0.0134 

μg/mL

With Bayesian calibration of exposure:

OR (95% CI)

per IQR (lnPFOA) increase in PFOA:

1.13 (0.92, 1.37) [IQR (lnPFOA)=0.00192 μg/mL]

per 0.100 μg/mL increase in PFOA:

0.97 (0.85, 1.11)

< 40th percentile 1.0 (referent)

40-60th 1.0 (0.7, 1.4)

60-80th 1.5 (1.1, 2.1)

≥ 80th 1.2 (0.8, 1.7)

Darrow et al. 2013

United States (C8 Health Project)

n = 1,330 pregnancies in 2005-2010; 770 

first pregnancies after PFOA measures; 

947 (pregnancies in 2005-2007)

Exposure based on serum collected in 

2005

Geometric mean 

0.016 μg/mL

Mean 0.031 

OR (95% CI) per log unit increase in PFOA

Full analysis: 1.27 (1.05, 1.55)

(adjusted for PFOS) 1.22 (0.99, 1.51)

By quintile:

Q1 up to 0.0069 μg/mL 1.0 (referent)

Q2 0.0069 – < 0.0111 2.39 (1.05, 5.46)

Q3 0.0111 – < 0.0189 3.43 (1.50 (7.82)

Q4 0.0189 - < 0.0372 3.12 (1.35, 7.18)

Q5 ≥ 0.0372 3.16 (1.35, 7.38)

First pregnancy after

 PFOA measure 1.23 (0.92, 1.64)

Pregnancies in 2005-2007: 1.35 (1.04, 1.76)

Darrow et al. (2013) examined birth outcomes in births that occurred in the 5 years after the 

PFOA measurement. In this study, reproductive history in a follow-up interview in 2010 was 

collected from females who had provided serum for PFOA measurement in 2005–2006. 

Singleton live births among 1,330 females after January 1, 2005, were linked to birth records to 

identify outcomes of pregnancy-induced hypertension and other outcomes (e.g., preterm birth, 

low birth weight, and birth weight among full-term infants). Thus there is a progressively greater 

refinement and reduction in misclassification (or exposure and outcome) among this set of 

Perfluorooctanoic acid (PFOA) – May 2016 3-39
studies. Each of these studies provides some evidence of an association between PFOA exposure 

and risk of pregnancy-induced hypertension or preeclampsia, with the most robust findings from 

the methodologically strongest study (Darrow et al. 2013). Maternal serum PFOA levels were 

positively associated with pregnancy-induced hypertension, with an adjusted OR per log unit 

increase in PFOA of 1.27 (95% CI: 1.15, 1.55). PFOS also was positively associated with 

pregnancy-induced hypertension. 

The C8 Science Panel (2012) considered both hypertension and preeclampsia together in 

determining a link between PFOA and pregnancy-induced hypertension. Some studies conducted 

by the panel found no associations while others showed positive associations. Among the studies 

with positive associations, no clear dose response was indicated. However, the panel decided that 

the evidence was sufficient to conclude that PFOA has a probable link to pregnancy-induced 

hypertension. 

Fetal growth. Many different measures of fetal growth can be used in epidemiology studies. 

Birth weight is widely available (as it is routinely collected in medical records and birth 

certificates). Low birth weight (defined as < 2,500 g) can be a proxy measure for preterm birth 

(particularly when accurate gestational age dating is not available). Other measures of fetal 

growth such as small for gestation age might more accurately reflect fetal growth retardation.

Both birth weight and gestational age are characterized as two-part distributions, with a 

larger Gaussian portion representing term births and a longer tail representing preterm births. 

Increased risks of complications, including infant mortality, are seen in preterm births (or low 

birth-weight births). When analyzed as a continuous measure, changes in birth weight might not 

be clinically significant, as small changes in the distribution among term infants do not result in a 

shift into the distribution seen in preterm infants (Savitz 2007; Wilcox 2010). This consideration 

differs from that of some other types of continuous measures, such as neurodevelopment scales, 

blood pressure, or cholesterol, in which shifts in the distribution are expected to move a greater 

proportion of the population into an “at risk” or “abnormal” level. 

High-exposure community studies. As noted in the previous discussion of preeclampsia, 

several studies using different designs and exposure measures have examined birth outcomes in 

infants born to mothers in the high-exposure C8 community population in West Virginia and 

Ohio (Darrow et al. 2013; Nolan et al. 2009; Savitz et al. 2012a, 2012b; Stein et al. 2009). These 

studies include analyses of birth weight and of low birth weight, and have not observed 

associations between PFOA and either birth weight among term births or the risk of low birth 

weight among all (singleton) births (Table 3-10). 

Based on these data, as well as continued follow-up of the community cohort, the C8 Science 

Panel (2012) concluded that there is no probable link between PFOA and low birth weight. 

General population studies. Two studies examined associations between maternal PFOA levels 

and birth weight among term infants (Fei et al. 2007; Monroy et al. 2008). The larger of these is 

from the Danish National Birth Cohort by Fei et al. (2007) (Table 3-10). In this study of 1,207 

term births, the change in birth weight per log unit increase in PFOA was -9 g (95% CI: -20, 

2 g). 

Perfluorooctanoic acid (PFOA) – May 2016 3-40
Table 3-10. Summary of Epidemiology Studies of PFOA and Birth Weight

Study PFOA Level Results

High-Exposure Community

Darrow et al. 2013

United States (C8 Health 

Project)

n = 1,629 pregnancies in 2005-

2010; 770 first pregnancies 

after PFOA measures; 947 

(pregnancies in 2005-2007)

Exposure based on serum 

collected in 2005

Geometric mean 

0.0162 μg/mL

Mean 0.031 μg/mL

Change in birth weight per log unit increase (95% CI)

Full analysis: -8 (-28, 12) g

(adjusted for PFOS) -4 (-25, 17) g

First pregnancy after

 PFOA measure 5 (-22, 33) g

Pregnancies in 2005-2007: -10 (-34, 14) g

OR (95% CI) for low birth weight (< 2500 g) per log unit 

increase

Full analysis: 0.94 (0.75, 1.17)

(adjusted for PFOS) 0.91 (0.72, 1.16)

First pregnancy after

 PFOA measure 1.07 (0.78, 1.47)

Pregnancies in 2005-2007: 0.91 (0.70, 1.17)

[Similar results in Nolan et al. 2009; Savitz et al. 2012a, 

2012b; Stein et al. 2009]

General Population: Birth Weight Among Term Births

Fei et al. 2007

Denmark

n = 1,207 (term births)

Blood sample at 4-14 weeks

0.0056 μg/mL Change in birth weight per unit increase (95% CI)

-8.7 (-19.5, 2.1)

Monroy et al. 2008

Canada

n = 101

Cord blood sample

0.0019 μg/mL (cord 

blood)

Change in PFOA per g change in birth weight:

Beta = 0.000171 (p = 0.65)

General Population: Birth Weight or Low Birth Weight Among All Births (by time of blood collection)

Fei et al. 2007

Denmark

Blood sample at 4-14 weeks

n = 1,399 (full sample)

3.8% preterm

0.0056 μg/mL Change in birth weight per unit increase (95% CI)

-10.6 (-20.8, -0.47) g

OR (95% CI) for low birth weight (< 2500 g) by quartile

Q1 up to 0.00390 μg/mL 1.0 (referent)

Q2 0.00391-0.00520 4.3 (0.51, 37)

Q3 0.00521-0.00696 3.7 (0.42, 32)

Q4 ≥ 0.00697 2.4 (0.27, 22) (Trend p = 0.94)

Hamm et al. 2010

Canada

n = 252

Blood sample at 15-16 weeks

8.3% preterm

0.0021 μg/mL Change in birth weight per ln unit increase (95% CI)

-37.4 (-86.0, 11.2) g

Whitworth et al. 2012

Norway

n = 849

Blood sample at around 17 

weeks

3.9% preterm

0.0021 μg/mL Change in birth weight z-score per unit increase (95% CI)

-0.03 (-0.10, 0.04)

Maisonet et al. 2012

United Kingdom

n = 395

Blood sample at 10-28 weeks

3.1% preterm

0.0037 μg/mL Change in birth weight per log unit increase

-34.2 (-54.8, -13) g

Washino et al. 2009

Japan

n = 428

Blood sample at 23-35 weeks

% preterm not reported

0.0014 μg/mL Change in birth weight per log unit increase (95% CI)

-75 (-191, 42) g

Perfluorooctanoic acid (PFOA) – May 2016 3-41
Study PFOA Level Results

Apelberg et al. 2007

United States (Baltimore)

n = 293

Cord blood sample

13% preterm

0.0016 μg/mL (cord 

blood)

Change in birth weight per log unit increase (95% CI)

-104 (-213, 5) g

Chen et al. 2012

Taiwan

n = 429

Cord blood sample

9.3% preterm

0.0018 μg/mL

(cord blood)

Change in birth weight per log unit increase (95% CI)

-19 (-63, 25) g

Fei et al. (2007, 2008a), and other studies in the general population have examined PFOA in 

relation to birth weight or risk of low birth weight (or other measures of fetal growth), without 

restriction to term births (Table 3-10). These studies vary in size from approximately 250 to 

1,400 births, and also in terms of timing of exposure measure. Fei et al. (2007, 2008a) used blood 

samples collected early in pregnancy (4–14 weeks), three studies used samples collected in the 

second trimester (Hamm et al. 2010; Maisonet et al. 2012; Whitworth et al. 2012), Washino et al. 

(2009) used samples collected in the third trimester, and two studies used cord blood samples 

(Apelberg et al. 2007; Chen et al 2012). These studies also differed in the percent of births that 

were preterm (ranging from approximately 3% to 13%), and presented results using different 

types of analyses (i.e., the form of the exposure and outcome variables, continuous, lntransformed, categorical, etc). Each of the analyses indicates a negative association between 

PFOA levels and birth weight (i.e., a decrease in birth weight with increase in PFOA), although 

CIs were wide. 

In a systematic review based on the Navigation Guide methods (Woodruff and Sutton 2014), 

Johnson et al. (2014) identified the general population studies shown in Table 3-10 and the highexposure C8 Health Project studies published through 2012. The results from the meta-analysis 

showed that a 0.001 μg/mL increase in serum or plasma PFOA was associated with a -18.9 g 

(95% CI -29.8, -7.9) difference in birth weight. 

Preeclampsia is a condition that causes the pregnant female to be hypertensive because of 

reduced renal excretion associated with a decrease in GFR. Preeclampsia is often accompanied 

by low birth weight (Whitney et al. 1987). Morken et al. (2014) used a subset of the Norwegian 

Mother and Child Cohort to evaluate the relationship between GFR and fetal size. Participants 

included 470 preeclamptic patients and 483 nonpreeclamptic females; plasma creatinine 

measured during the second trimester was used to estimate GFR. For the overall cohort, for each 

mL/min increase in GFR, infant weight at birth increased 0.73–0.83 g, depending on the method 

used to calculate GFR. The increases were greater and statistically significant in females with 

preeclampsia. Differences were not statistically significant for the nonpreeclamptic group. 

Morken et al. (2014) was not a study of perfluorochemicals and there were no serum 

measurements of any PFASs. However, because PFOA/PFOS serum levels are expected to be 

higher with a lower GFR, the finding stimulated examination of the GFR as it relates to serum 

PFAS levels and the low birth weight identified in the epidemiology studies (Verner et al. 2015; 

Vesterinen et al. 2014). 

The evidence for an inverse association between PFOA levels and birth weight raised the 

question of whether reverse causality linked to maternal GFR played a role in the association of 

low birth weight with serum PFOA. PFOA excretion by the kidney is dependent, in part, on the 

Perfluorooctanoic acid (PFOA) – May 2016 3-42
GFR. Conditions that result in impairment of GFR (and thus increased serum PFOA) also could 

be related to fetal growth restriction, confounding the association between serum PFOA and 

decreased birth weight. Vesterinen et al. (2014) examined evidence pertaining to the relationship 

between fetal growth and maternal GFR using Navigation Guide systematic review methods. 

They identified 35 relevant studies published between 1954 and 2012 that met the Navigation 

Guide criteria for inclusion in the analysis. All studies were rated as “low” or “very low” quality 

due to inconsistency of findings among studies, small sample sizes resulting in large CIs around 

a mean, and high risk of bias in conduct of the study. The quality rating led to the conclusion that 

data were “inadequate” to determine an association between fetal growth and GFR. However, a 

more recent publication described below, expanded the database on the relationship between 

GFR and fetal size.

Verner et al. (2015) modified the human pregnancy/lactation PK model of PFOA by 

Loccisano et al. (2013) described in section 2.6.1 to evaluate the association between GFR, 

serum PFOA levels, and birth weight. When GFR was accounted for in the model simulations, 

the reduction in birth weight associated with increasing serum PFOA was less than that found by 

the author’s meta-analysis of the same data. This finding suggests that a portion of the 

association between prenatal PFOA and birth weight is confounded by maternal GFR differences 

within the populations studied. The true association for each 1 ng/mL increase in PFOA could be 

closer to a 7-g reduction (95% CI -8, -6) compared to the 14.72-g reduction (95% CI: -

8.92, -1.09) predicted by meta-analysis of the epidemiology data without a correction for low 

GFR as observed in individuals with pregnancy-induced hypertension or evidence of 

preeclampsia. 

Other pregnancy outcomes. Gestational age and preterm birth and risk of miscarriage were not 

associated with PFOA in the studies examining pregnancy outcomes in the high-exposure 

community (Darrow et al. 2014; Nolan et al. 2009, 2010). In contrast, PFOS was positively 

associated with miscarriage (Darrow et al. 2014). 

Congenital anomalies were diagnosed in 1.8%, 1.9%, and 2.0% of the mothers with water 

provided completely, partially, or not at all by LHWA, respectively (Nolan et al. 2010). When 

adjusted for confounders, no statistically significant differences were found. Complications with 

labor and delivery were observed in 32.5%, 35.9%, and 41.9% of the mothers with water 

provided completely, partially, or not at all by LHWA, respectively. Mothers with water 

provided by LHWA did have in increased likelihood of having dysfunctional labor, but the 

number of reported cases was low. Mothers with one or more maternal risk factors were 37.5%, 

34.4%, and 39.3% of the populations with water provided completely, partially, or not at all by 

LHWA, respectively. Adjusted regression models showed no statistical differences across water 

service status. An increased likelihood of anemia (crude OR 11, 95% CI: 1.8–64) and 

dysfunctional labor (crude OR 5.3, 95% CI: 1.2–24) in mothers with water provided by LHWA 

was found, but the number of reported cases was low. No association was found between PFOA 

and increased incidence of congenital anomalies, other labor and delivery complications, or 

maternal risk factors.

The C8 Science Panel (2012) concluded that there is no probable link between PFOA and 

birth defects, miscarriage, preterm birth, or stillbirth. Their conclusion was based on findings in 

Nolan et al. (2010), Stein et al. (2009), and other data available to the panel. These other data 

included historical estimates of serum PFOA generated by the panel based on amounts released 

from the plant and an individual’s residential history.

Perfluorooctanoic acid (PFOA) – May 2016 3-43
Fei et al. (2009) examined the association between plasma PFOA concentration and longer 

time to pregnancy (TTP) as a measure of fecundity in 1,240 females. TTP was categorized as 

follows: immediate pregnancy (<1 month), 1–2, 3–5, 6–12, and >12 months. Having >12 months 

TTP or having used fertility treatment to get pregnant were used to define infertility. A total of 

620 females had a TTP within the first 2 months of trying to conceive and 379 had a TTP of 

≥6 months with 188 of those females having a TTP of >12 months. The mean plasma PFOA 

concentration was 0.0056 μg/mL for females who planned their pregnancies, and 0.0054, 0.0060, 

and 0.0063 μg/mL for TTPs <6 months, 6–12 months, and >12 months, respectively. Plasma 

PFOA concentration was significantly greater (p<0.001) in females who had TTPs >6 months 

than those with TTPs <6 months. The females with TTPs >6 months were more likely to be 

older, have middle socio-occupational status, and have a history of spontaneous miscarriage or 

irregular menstrual cycles. The adjusted odds for infertility increased 60–154% among females

with >0.00391 μg/mL plasma PFOA concentration compared to females with <0.00391 μg/mL 

plasma concentration. The fecundity OR was 0.72, 0.73, and 0.60 for the three highest PFOA 

concentration quartiles. In the likelihood ratio test, the trend was significant (p<0.001). Both TTP 

and infertility also were positively associated with serum PFOS levels in this study. Although the 

results of the study suggest that plasma PFOA concentration could reduce fecundity, the authors 

noted that selection bias, the unknown quality of the sperm, unknown frequency and timing of 

intercourse, and abnormal hormone levels might have an impact on the results and fecundity. 

Participants enrolled in the Maternal-Infant Research on Environmental Chemicals Study, a 

Canadian pregnancy and birth cohort, were evaluated for an association between serum PFOA 

levels (as well as PFOS and PFHxS) and TTP (Vélez et al. 2015). A total of 1,743 females, 

enrolled between 2008 and 2011 and having a blood sample collected during the first trimester 

were included. Infertility was defined as having a TTP of >12 months or requiring infertility 

treatment for the current pregnancy. The geometric mean plasma PFOA level was 0.00166 

μg/mL. The crude fecundity OR per one SD increase in log-transformed serum concentration 

was significantly lower for PFOA (OR=0.91, 95% CI 0.86, 0.96) (and for PFHxS). In fully 

adjusted models, PFOA (and PFHxS) was associated with an 11% reduction in fecundability per 

one SD increase in log-transformed serum concentration (OR=0.89; 95% CI 0.83, 0.94). The 

adjusted odds of infertility increased by 31% per one SD increase of PFOA (OR=1.31; 95% CI 

1.11–1.53) (and of PFHxS). No significant associations were observed for PFOS.

Fei et al. (2010b) reported on the effects of PFOA and PFOS on the length of breast-feeding. 

Self-reported data on the duration of breast-feeding were collected during the telephone 

interviews with each mother at 6 and 18 months after birth of the child. Higher levels of PFOA 

were significantly associated with a shorter duration of breast-feeding. In multiparous females, 

the adjusted OR for weaning before 6 months was 1.23 (95% CI, 1.13–1.33) for each 1-ng/mL 

increase in PFOA concentration in the maternal blood and the increase was dose-related. A 

similar association was observed with PFOS levels. No association was found between length of 

breast-feeding and PFOA levels in females having their first child. The authors speculate that the 

observed associations might be noncausally related to previous length of breast-feeding or to 

reduction of PFOA and PFOS through lactation. 

Pubertal development. Two studies examined development of puberty in females in relation to 

prenatal exposure to PFOA as measured through maternal or cord blood samples (Christensen et 

al. 2011; Kristensen et al. 2013), and another study examined PFOA exposure measured 

concurrently with the assessment of pubertal status in females and in males (Lopez-Espinosa et 

al. 2011) (Table 3-11).

Perfluorooctanoic acid (PFOA) – May 2016 3-44
Table 3-11. Summary of Epidemiology Studies of PFOA and Pubertal Development

Study PFOA Level Results

Prenatal Exposure: General Population

Christensen et al. 2011

United Kingdom

Pregnancy cohort, with casecontrol of early menarche in 

follow-up

n = 218 cases (menarche before 

age 11.5 yrs) and 230 controls

0.0036-0.0039 

μg/mL 

(maternal)

Median (75th percentile) in

 cases: 3.9 (5.0)

 controls: 3.6 (4.7) (p = 0.15)

OR (95% CI)

above versus below median 1.29 (0.86, 1.93)

per ln-unit increase in PFOA 1.01 (0.61, 1.68) 

Kristensen et al. 2013

Denmark

Pregnancy cohort, with followup of 343 (79% of eligible) 

daughters at age 20

Health questionnaire and 

exams/hormone measurements

(for n = 254)

0.0036 μg/mL 

(maternal)

Difference in age at menarche (months) by exposure group

low (0.001–0.003 μg/mL ) 0.0 (referent)

medium (0.003–0.0043 μg/mL) 0.9 (-3.0, 4.8)

high (0.0044–0.0198 μg/mL) 5.3 (1.3, 9.3)

continuous 1.01 (0.22, 1.89)

Peripubertal Exposure: High-Exposure Community

Lopez-Espinosa et al. 2011

United States (C8 Health Project)

n = 2,931 girl and 3,076 boy, 

aged 8-18 yrs

Self-reported menarche (girls) 

and free or total testosterone 

(boys) 

Median 0.058 

μg/mL

Prevalence of menarche in girls

OR 95% CI days delay

Q1: 1.0 (referent)

Q2: 1.01 (0.65-1.58) -4

Q3: 1.00 (0.64-1.58) -1

Q4: 0.75 (0.49-1.15) 69

Prevalence of delayed puberty in boys

OR 95% CI days delay

Q1: 1.0 (referent)

Q2: 0.54 (0.35-0.84) 142

Q3: 0.50 (0.32-0.87) 163

Q4: 0.57 (0.37-0.89) 130

Results were broadly similar when the analysis was based on estradiol 

levels to define menarche or when the models included PFOA and 

PFOS jointly, though significance was reduced in some comparisons.

Christensen et al. (2011) used data from a prospective cohort study in the United Kingdom to 

perform a nested case-control study examining the association between age at menarche and 

gestational exposure to perfluorinated chemicals, including PFOA and PFOS. The study 

population from the Avon Longitudinal Study of Parents and Children included single-birth 

female subjects who had completed at least two puberty staging questionnaires between the ages 

of 8 and 13 years and whose mothers provided at least one analyzable prenatal serum sample. If 

more than one serum sample was available, the earliest sample provided was used for analysis. 

The study does not provide information as to when samples were collected. The females were 

divided into two groups: those who experienced menarche prior to age 11.5 years (n = 218) and a 

random sample of those who experienced menarche after age 11.5 (n = 230). Confounders such 

as the mother’s prepregnancy BMI, age at delivery, age at menarche, educational level, and the 

child’s birth order and ethnic background were included in linear and logistic regression models 

used to analyze the data. The median maternal serum PFOA concentrations were 0.0039 and 

0.0036 μg/mL for the early menarche and nonearly menarche groups, respectively. The authors 

noted a modest nonsignificant association between the odds of earlier menarche and prenatal 

serum PFOA concentrations above the median. For all models, the CIs included the null value of 

1.0. Similar results were obtained for PFOS. 

Perfluorooctanoic acid (PFOA) – May 2016 3-45
Effects of prenatal exposure to PFOA (and PFOS) on female and male reproductive function 

was evaluated in 343 females and 169 males whose mothers participated in a cohort in 1988–

1989 (Kristensen et al. 2013; Vested et al. 2013). Maternal blood samples were collected during 

week 30 of gestation. Follow-up was initiated in 2008 when the offspring were ~20 years old. 

Median serum PFOA level was 0.0036 μg/mL for the mothers with daughters evaluated. In 

adjusted regression analysis, daughters from mothers in the highest PFOA tertile had a 

5.3-month later age at menarche (95% CI 1.3, 9.3) than those in the lowest tertile. No association 

was found between prenatal exposure to PFOS and age of menarche. No statistically significant 

relationships were found between PFOA (or PFOS) exposure and cycle length, reproductive 

hormone levels, or number of follicles assessed by ultrasound (Kristensen et al. 2013). 

Lopez-Espinosa et al. (2011) examined the association of serum PFOA concentration and the 

age of puberty in exposed children of the Mid-Ohio Valley. Data from the C8 Health project 

(e.g., sex steroid hormone levels, self-reported menarche status) along with detailed date of birth 

information were used to determine age of puberty in males (n = 3,076) and females (n = 2,931) 

aged 8–18 years. Serum PFOA concentrations were divided into quartiles: <0.0114, 0.0114–

0.023, >0.023–0.058, and >0.058 μg/mL. Confounders such as age at survey, BMI, BMI z-score, 

height, family income, ethnicity, smoking status, alcohol consumption, and date and time of 

sample collection were included in the logistic regression models used to analyze the data. The 

median PFOA concentrations were 0.026 and 0.020 μg/mL in males and females, respectively. 

No association between PFOA concentration and puberty was observed for males. Reduced odds 

of having reached puberty was associated with higher PFOA exposure in females (OR=0.57, 

95% CI 0.37–0.89). There were 130 days of delay between the highest and lowest quartile. 

Reduced odds of experiencing menarche at a younger age (10–15 years) also was observed (OR 

0.83, 95% CI 0.74–0.93). The results suggested that PFOA was associated with a later age of 

menarche. PFOS was associated with delayed puberty in both males and females. The authors 

expressed caution in interpretation of the data because of lack of serum PFOA concentration 

prior to puberty, PFOA concentration having been measured after the attainment of puberty, and 

lack of secondary sexual maturation data (i.e., physical, Tanner criteria, and biomarker 

measurements).

Male reproductive effects. Joensen et al. (2009) examined the association between PFASs, 

including PFOA, and testicular function in 105 Danish males who provided semen and blood 

samples as part of reporting for the military draft in 2003. The males chosen for the study had the 

highest testosterone concentrations (ranging from 30.1 to 34.8 nmol/L; n = 53; 18.2–24.6 years) 

and lowest testosterone concentrations (ranging from 10.5 to 15.5 nmol/L; n = 52; 18.2–25.2 

years). Regression models were used to analyze associations between PFOA and testicular 

function. Median serum PFOA concentration was 0.0044, 0.0050, and 0.0049 μg/mL in the high 

testosterone, low testosterone, and combined groups, respectively. A nonsignificant negative 

association was observed between serum PFOA concentration and semen volume, sperm 

concentration, sperm count, sperm motility, or sperm morphology. No association was observed 

between serum PFOA concentration and testosterone, estradiol, sex hormone-binding globulin 

(SHBG), luteinizing hormone (LH), follicle-stimulating hormone (FSH), and inhibin B. 

However, significantly fewer (p<0.05) morphologically normal sperm were seen in males with 

high combined levels of PFOA/PFOS (6.2 million spermatozoa) than in males with low 

PFOA/PFOS levels (15.5 million spermatozoa).

In a slightly expanded study, Joensen et al. (2013) investigated the associations between 

PFASs, including serum PFOA concentration, and reproductive hormones and semen quality in 

Perfluorooctanoic acid (PFOA) – May 2016 3-46
247 healthy young Danish males (mean age 19.6 years). Serum samples were analyzed for PFOA

as well as total testosterone (T), estradiol, SHBG, LH, FSH, and inhibin-B. The mean PFOA 

level was 0.0035 μg/mL. No associations were found between PFOA levels (or 12 other PFAS) 

and any hormone level or semen quality parameters. PFOS levels were negatively associated 

with testosterone.

An association between serum levels of seven PFASs and 35 semen quality parameters was 

evaluated in 462 males enrolled in the LIFE study cohort (Buck Louis et al. 2015). The males

were from Michigan and Texas with a mean age of 31.8 years and mean PFOA levels 0.00429–

0.00509 μg/mL. PFOA was significantly associated with a lower percentage of sperm with 

coiled tails, an increased curvilinear velocity, and a slightly larger acrosome area of the head. In

total, six PFASs (including PFOA) were associated with changes in 17 semen quality endpoints. 

Effects of prenatal exposure to PFOA (and PFOS on male reproductive function was 

evaluated in 169 males whose mothers participated in a cohort in 1988–1989 (Vested et al. 

2013). Maternal blood samples were collected during week 30 of gestation. Follow-up was 

initiated in 2008 when the offspring were ~20 years old. Median serum PFOA level was 0.0038 

μg/mL for mothers with sons evaluated. Multivariable regression models showed significant 

negative trends for sperm concentration and total sperm count in association with in utero

exposure to PFOA. A 34% reduction in sperm concentration (95% CI 58, 5%) and a 34% 

reduction in total count (95% CI 62, 12%) were estimated for the highest exposure tertile 

compared with the lowest tertile. Maternal PFOA level also was positively associated with 

higher FSH and LH levels in the sons. No associations were found between PFOA level and 

percentage of progressive sperm, sperm morphology, semen volume, or testicular volume. PFOS 

was not associated with any outcome (Vested et al. 2013). 

3.1.1.9 Steroid Hormones

Occupational exposure studies. Olsen et al. (1998) examined several hormones, including 

cortisol, estradiol, FSH, dehydroepiandrosterone sulfate, 17 gamma-hydroxyprogesterone (a 

testosterone precursor), free testosterone, T, LH, prolactin, and SHBG in male workers at the 

Cottage Grove, Minnesota, production plant for 1993 and 1995. This was the same population 

used for the thyroid hormone study described above for 111 workers in 1993 and 80 in 1995. 

Employees were placed into four exposure categories based on their serum PFOA levels: 

0–1 μg/mL, 1– < 10 μg/mL, 10– < 30 μg/mL, and >30 μg/mL. Statistical methods used to 

compare PFOA levels and hormone values included multivariable regression analysis, ANOVA, 

and Pearson correlation coefficients. No association between serum PFOA and any hormone was 

observed, but some trends were observed. When the mean measures of the various hormones 

were compared by exposure categories, there was a statistically significant elevation in prolactin

(p = 0.01) in 1993 only for the 10 workers whose serum PFOA levels were between 10 and 

30 μg/mL compared to the lower two exposure categories. 

Estradiol levels in the >30 μg/mL PFOA group in both years were 10% higher than in the 

other PFOA groups, but the difference was not statistically significant. These results were 

confounded by estradiol being correlated with BMI (r = 0.41, p<0.001 in 1993, and r = 0.30, 

p<0.01 in 1995). The authors postulated that the study might not have been sensitive enough to 

detect an association between PFOA and estradiol because measured serum PFOA levels were 

likely below the observable effect levels suggested in animal studies (e.g., 55 μg/mL PFOA in 

the CD rat). Only three employees in this study had PFOA serum levels that high. They also 

Perfluorooctanoic acid (PFOA) – May 2016 3-47
suggest that the higher estradiol levels in the highest exposure category could suggest a threshold 

relationship between PFOA and estradiol. 

In the Sakr et al. study (2007a) of 1,025 workers at the DuPont Washington Works facility in 

West Virginia, an association was observed between serum PFOA and serum estradiol 

(p = 0.017) and testosterone (p = 0.034) in male workers; however, circadian variations of 

hormones were not taken into consideration during analysis. The biological significance of the 

results is unknown. 

Costa et al. (2009) found no association between serum PFOA concentration and estradiol or 

testosterone in 53 male workers at a PFOA production plant in Italy based on medical 

surveillance data collected between 2000 and 2007. 

High-exposure community studies. Knox et al. (2011) examined the endocrine disrupting 

effects of perfluorocarbons in females from the C8 Health Project by analyzing the relationship 

between serum PFOA, serum estradiol concentration, and menopause onset. The population 

included females over age 18 years (n = 25,957). Serum PFOA and estradiol concentrations were 

determined from blood samples. Females who were pregnant; had had full hysterectomies; and 

were taking any prescription hormones, selective estrogen receptor modulators, and/or fertility 

agents were excluded from estradiol analysis. Serum PFOA concentrations were grouped into 

quintiles (natural log-transformation)—Q1 = 0.00025–0.0112; Q2 = 0.0113–0.0198; Q3 = 

0.0199–0.0367; Q4 = 0.0368–0.0849; and Q5 = 0.0850–22.412 μg/mL. Estradiol analysis was 

calculated by age group—18–42 years, >42 ≤ 51 years, and >51 ≤ 65 years. Menopause was 

determined by questionnaire. Menopause analysis was calculated by age group—30–42 years, 

>42 ≤ 51 years, and >51 ≤ 65 years—and excluded those who reported having had 

hysterectomies. Logistic regression models were adjusted for smoking, age, BMI, alcohol 

consumption, and regular exercise. PFOA concentration in females who had had hysterectomies

was significantly higher than in females who had not had hysterectomies. Serum PFOA and 

estradiol concentrations were not associated, while PFOS levels were negatively associated with 

estradiol. The odds of attaining menopause analysis in the oldest group of females, showed that 

all quintiles were significantly higher for all quintiles than the lowest, and in females between the 

ages of 42 and 51 years, Q3, Q4, and Q5 were significantly higher than the lowest. PFOS also 

was associated with increased odds of attaining menopause in women 42–51 years and >51 

years. Data interpretation was limited by the cross-sectional study design and survey-reported 

menopause without age or independent confirmation. 

3.1.1.10 Neurodevelopment

High-exposure community studies. A subset of 321 children enrolled in the C8 Health Project 

was assessed for neurobehavioral development 3–4 years after enrollment (Stein et al. 2013). 

The children had serum samples collected at enrollment in 2005–2006 with the current follow-up 

evaluation conducted in 2009–2010, when the children were 6–12 years old. Both the mother and 

teacher completed surveys to elicit information on each child’s executive function, attention 

deficit hyperactivity disorder- (ADHD-) like behavior, and behavioral problems. Information on 

family demographics and other health conditions of the child were included as confounders. 

Linear regression was used to determine the association between PFOA levels and mother and 

teacher reports.

Perfluorooctanoic acid (PFOA) – May 2016 3-48
The median PFOA level was 0.0351 μg/mL with an IQR of 0.0158–0.0941 μg/mL. When 

comparing the highest to the lowest PFOA quartile, survey results from the mother for both 

executive function and ADHD showed a favorable association for males, but an adverse 

association for females. These findings were not replicated when males and females were 

analyzed together or with results from the teacher surveys. No association was found between 

PFOA levels and either mother or teacher scores for behavioral problems in females and males.

In 2012, the C8 Science Panel concluded that there is no probable link between PFOA 

exposure and neurodevelopmental disorders in children, including attention deficit disorders and 

learning disabilities. Their conclusion was based on epidemiology studies conducted by the panel 

and other data available.

General population studies. Fei et al. (2008b) examined the association between plasma PFOA 

concentration in pregnant females and motor and mental developmental milestones of their 

children. The mothers self-reported the infant’s fine and gross motor skills and mental 

development at 6 and 18 months of age. There was no association between maternal plasma 

PFOA concentration and Apgar score or between maternal plasma PFOA concentration and fine 

motor skills, gross motor skills, or cognitive skills at 6 and 18 months of age. The children born 

to females having higher plasma PFOA concentrations reached developmental milestones at the 

same times as children born to females having lower plasma PFOA concentrations. The authors 

concluded that there was no association between maternal early pregnancy levels of PFOA and 

motor or mental developmental milestones in offspring. However, in children at 18 months, 

mothers with higher PFOS levels were slightly more likely to report that their babies started 

sitting without support at a later age.

A subset of the Danish National Birth Cohort was evaluated for an association between 

prenatal PFAS exposure and the risk of cerebral palsy (Liew et al. 2014). A total of 156 cases of 

cerebral palsy were identified and matched to 550 randomly selected controls. Stored maternal 

plasma samples were analyzed for 16 PFAS and six compounds were quantifiable in >90% of 

the samples. For the cerebral palsy cases and matched controls, median maternal PFOA levels 

were 0.00456 and 0.00400 μg/mL, respectively, for males and 0.00390 and 0.00404 μg/mL, 

respectively, for females. Per natural-log unit increase in maternal PFOA level, the risk of 

developing cerebral palsy in males was significantly increased (RR=2.1; 95% CI 1.2, 3.6). 

Positive associations were also found with PFOS and perfluoroheptane sulfonate. No association 

was found between any PFAS level and risk of cerebral palsy in females.

Fei and Olsen (2011) examined the association between prenatal PFOA (and PFOS) exposure 

and behavior or coordination problems in children at age 7. The children and their mothers were 

part of the Danish National Birth Cohort. Behavioral problems were assessed using the Strengths 

and Difficulties Questionnaire (SDQ), and coordination problems were assessed using the 

Developmental Coordination Disorder Questionnaire (DCDQ) completed by the mothers. A total 

of 787 mothers completed the SDQ and 537 completed the DCDQ for children aged 7.01–8.47 

years (mean age 7.15 years). The mean maternal PFOA concentration was 0.0057 μg/mL, and 

PFOA levels were divided into quartiles: <LLOQ-0.00395, 0.00396–0.00532, 0.00535–0.00711, 

and 0.00714–0.02190 μg/mL. A child having higher scores in total difficulties, emotional 

symptoms, and hyperactivity was negatively associated with the second or third PFOA quartiles 

(OR=0.56, 95% CI 0.27–1.19; p<0.05 and OR=0.36, 95% CI 0.15–0.82; p<0.05, respectively) 

when compared with females in the lowest quartile. ORs adjusted for parity, maternal age, 

prepregnancy BMI, pregnancy smoking and alcohol consumption, socio-occupational status, 

child gender, breast-feeding, birth year, home density, gestational age at blood draw, and 

Perfluorooctanoic acid (PFOA) – May 2016 3-49
parental behavior problem as children did not show a positive association between prenatal 

PFOA exposure and behavior or coordination problems. Overall, no significant association 

between behavioral or coordination problems in children 7 years of age and prenatal PFOA (and 

PFOS) exposure was found. 

Similar to the above study, the association between maternal PFOA (and PFOS) levels and 

offspring behavior and motor development was investigated in a subset of the Biopersistent 

Organochlorines in Diet and Human Fertility study (INUENDO) birth cohort (Høyer et al. 

2015a). Pregnant females were enrolled between May 2002 and February 2004 with a total of 

1,106 mother-child pairs at follow-up between January 2010 and May 2012, when the children 

were 7–9 years old. The study population consisted of 526 pairs from Greenland, 89 pairs from 

Poland, and 491 pairs from Ukraine. Maternal blood samples for measurement of plasma PFOA 

levels were taken any time during pregnancy. Behavior of children was assessed with SDQ

score, and logistic regression models were used in the analyses of PFOA tertile levels and 

behavioral problems. Motor development was assessed with DCDQ score, and linear regression 

was used for analyses. All analyses were performed on the entire cohort as well as by country, 

except that not all analyses could be performed on the Polish subset because of the small number 

of cases. The median maternal plasma PFOA level was 0.0014 μg/mL for the combined 

population and 0.0018, 0.001, and 0.0027 μg/mL for the pregnant females from Greenland, 

Ukraine, and Poland, respectively. 

No associations were found between PFOA (and PFOS) levels and motor development score. 

Total SDQ score was not associated with PFOA levels; however, the OR of having an abnormal 

total SDQ score was 2.7 (95% CI 1.2, 6.3) for all groups combined. PFOS levels were associated 

with higher total SDQ score only in Greenland. The highest PFOA tertile was associated with a 

0.5-point higher hyperactivity score in both the combined analysis and in Greenland, but no 

associations were found in Poland and Ukraine. The OR for hyperactive behavior in the 

combined analysis was 3.1 (95% CI 1.3, 7.2) for the highest tertile compared to the lowest PFOA 

tertile. In Greenland, the ORs for hyperactivity were increased for the middle (OR=5.4, 95% CI 

1.1, 25.6) and highest (OR=6.3, 95% CI 1.3, 30.1) tertiles (Høyer et al. 2015a). 

Hoffman et al. (2010) examined the associations between perfluorochemicals, including 

PFOA, and diagnosis of ADHD using the NHANES data from 1999–2000 and 2003–2004. The 

study population comprised 571 children aged 12–15 years, including those who had been 

diagnosed as having ADHD (n = 48) and/or were taking ADHD medications (n = 21). Age, 

gender, and race/ethnicity were included as covariates; and socioeconomic status, health 

insurance coverage and having a routine health care provider, living with someone who smokes, 

birth weight, admittance to a neonatal intensive care unit, maternal smoking, and preschool 

attendance were confounders. Regression models were used to analyze the data. The median 

serum PFOA level was 0.0044 μg/mL and ranged from 0.0004 to 0.0217 μg/mL. Serum PFOA 

was positively associated with parental report of ADHD (OR=1.12, 95% CI 1.01–1.23). The OR

for serum PFOA and parental report of ADHD and ADHD medication use was 1.19 (95% CI 

0.95–1.49). Both PFOS and perfluorohexane sulfonate also were positively associated with 

parentally reported ADHD. Data interpretation was limited by the cross-sectional study design, 

random misclassification error resulting from using current PFOA levels as proxy measures of 

etiologically relevant exposures, and other confounders not included in the available data. 

Perfluorooctanoic acid (PFOA) – May 2016 3-50
3.1.1.11 Postnatal Development 

General population studies. Andersen et al. (2010) examined the association between maternal 

plasma PFOA concentration and offspring weight, length, and BMI at 5 and 12 months of age. 

The mothers (n = 1,010) reported the information during an interviews, and weight and length 

measurements were used to calculate BMI. The median PFOA level was 0.0052 μg/mL with a 

range of 0.0005–0.0219 μg/mL. Maternal plasma PFOA concentration was inversely associated 

with weight at 5 months (β -30.2, 95% CI -59.3–-1.1), BMI at 5 months (β -0.067, 95% CI -

0.129–-0.004), weight at 12 months (β -43.1, 95% CI -82.9–-3.3 ), and BMI at 12 months in 

male children (β -0.078, 95% CI -0.144–-0.011) in models adjusted for maternal age, parity, 

prepregnancy BMI, smoking, gestational age at blood draw, socioeconomic status, and breastfeeding. Similar inverse associations were found with PFOS. No associations were observed 

between maternal plasma PFOA concentration and the endpoints for female children in the 

adjusted models.

The latest report on the Danish National Birth Cohort evaluated an association between 

maternal plasma PFOA levels and the children’s BMI, waist circumference, and risk of being 

overweight at 7 years of age (Andersen et al. 2013). From the subset of 1,400 females who 

provided blood samples during their first trimester, children were included if they had weight 

and height information (n = 811) or waist measurements (n = 804) at age 7 years. The median 

PFOA level was 0.0053 μg/mL with a range of 0.0005–0.0219 μg/mL. Maternal PFOA levels 

were inversely associated with all of the children’s anthropomorphic endpoints, but statistical 

significance was not attained and a dose response was not observed. Maternal PFOA (or PFOS) 

did not affect the risk of being overweight in either males or females. 

The association between maternal PFOA (and PFOS) levels and prevalence of offspring that 

are overweight plus waist-to-height ratio >0.5 was investigated in a subset of the INUENDO

birth cohort (Høyer et al. 2015b). Pregnant females were enrolled between May 2002 and 

February 2004 with a total of 1,022 mother-child pairs at follow-up between January 2010 and 

May 2012, when the children were 7–9 years old. The study population consisted of 531 pairs 

from Greenland and 491 pairs from Ukraine. Maternal blood samples for measurement of plasma 

PFOA levels were taken at a mean gestational age of 24 weeks. Each child’s weight and height 

were measured and BMI calculated. All analyses were performed on the entire cohort as well as 

by country. 

The median maternal plasma PFOA level was 0.0018 μg/mL in pregnant females from 

Greenland and 0.0010 μg/mL in pregnant females from Ukraine. No associations were found 

between PFOA (and PFOS) levels and risk of being overweight in the combined analysis or in 

Ukraine. In Greenland, the risk of being overweight was slightly increased only for females 

(RR=1.81, 95% CI 1.04, 3.17). PFOA association with risk of having waist-to-height ratio >0.5 

was slightly increased for the combined analysis (RR=1.30, 95% CI 0.97, 1.74), but statistical 

significance was not attained. PFOS levels were significantly associated with waist-to-height 

ratio >0.5 in the combined analysis (Høyer et al. 2015b). 

Halldorsson et al. (2012) examined prenatal exposure to PFASs, including PFOA, and the 

risk of being overweight at 20 years of age in a prospective study. A birth cohort consisting of 

665 mother-offspring pairs was recruited from a midwife center in Aarhus, Denmark. Maternal 

PFOA levels were measured in serum samples collected during week 30 of gestation for 

assessment of in utero PFOA exposure and offspring anthropometry at 20 years. The median 

PFOA concentration was 0.0037 ± 0.0020 μg/mL with quartiles of 0.0024 ± 0.0006, 0.0033 ± 

Perfluorooctanoic acid (PFOA) – May 2016 3-51
0.0004, 0.0042 ± 0.0005, and 0.0058 ± 0.0019 μg/mL. Three PFASs, including PFOS, 

perfluorooctane sulphonamide, and perfluorononanoate, increased across quartiles of PFOA 

concentration, but eight other PFASs did not. In covariate-adjusted analyses, female offspring 

whose mothers were in the highest quartile had 1.6 kg/m2 higher BMI (95% CI: 0.6, 2.6) and 

4.3 cm larger waist circumference (95% CI: 1.4, 7.3) than offspring whose mothers were in the 

lowest quartile. Female offspring of mothers in the highest versus lowest PFOA quartile were 

also more likely to be overweight [RR 3.1 (95% CI: 1.4, 6.9)] and to have a waist circumference 

>88 cm at 20 years of age [3.0 (95% CI: 1.3, 6.8)]. Among female participants who provided 

blood samples at clinical examination (n = 252), maternal PFOA concentration was positively 

associated with insulin, leptin, and the leptin-adiponectin ratio; and inversely associated with 

adiponectin levels. PFOA was not associated with being overweight or obesity in male offspring. 

The other PFASs were not significantly associated with any endpoint after adjustment for PFOA.

Geiger et al. (2014b) used data from the NHANES to determine whether there was a 

relationship between serum PFOA levels and hypertension in children. A total of 1,655 

participants (aged 12–18 years) from the 1999–2000 and 2003–2008 cycles of the survey who 

had PFOA measurements available were examined. Blood pressure was measured to determine 

the presence of hypertension, and linear regression modeling was used to study the association 

between increasing quartiles of serum PFOA and mean changes in systolic and diastolic blood 

pressures. Mean PFOA level was 0.0044 ± 0.0001 μg/mL. No association was found between 

serum PFOA (or PFOS) levels and hypertension in either unadjusted or multivariable-adjusted 

analyses. Compared with the lowest quartile, the multivariable-adjusted OR (95% CI) of 

hypertension in the highest quartile of exposure was 0.69 (0.41–1.17) (P-trend >0.30). 

3.1.1.12 Summary and Conclusions from the Human Epidemiology Studies

Numerous epidemiology studies have been conducted of workers, a large highly exposed 

community (the C8 Health Project), and the general population to evaluate the association of 

PFOA exposure to a variety of health endpoints. Health outcomes assessed include blood lipid 

and clinical chemistry profiles, thyroid effects, diabetes, immune function, birth and fetal and 

developmental growth measures, and cancer.

Serum lipids. The association between PFOA and serum lipids has been examined in several 

studies in different populations. Cross-sectional and longitudinal studies in occupational settings 

(Costa et al. 2009; Olsen et al. 2000, 2003; Olsen and Zobel 2007; Sakr et al. 2007a, 2007b; 

Steenland et al. 2015) and in the high-exposure community (the C8 Health Project study 

population) (Fitz-Simon et al. 2013; Frisbee et al. 2010; Steenland et al. 2009; Winquist and 

Steenland 2014a) generally observed positive associations between serum PFOA and TC in 

adults and children (aged 1–< 18 yrs); most of these effect estimates were statistically 

significant. Although exceptions to this pattern are present (e.g., some of the analyses examining 

incidence of self-reported high cholesterol based on medication use [Steenland et al. 2015; 

Winquist and Steenland 2014a]), the results are relatively consistent and robust. Similar 

associations were seen in analyses of LDL, but were not seen with HDL. The range of exposure 

in occupational studies is large (with means varying between 0.4 and > 12 μg/mL), and the mean 

serum levels in the C8 population studies were around 0.08 μg/mL. Positive associations 

between serum PFOA and TC (i.e., increasing lipid level with increasing PFOA) were observed 

in most of the general population studies at mean exposure levels of 0.002–0.007 μg/mL 

(Eriksen et al. 2013; Fisher et al. 2013; Geiger et al 2014a; Nelson et al. 2010; Starling et al. 

2014). The interpretation of results for these general population studies is limited, however, by 

the moderately strong correlations (Spearman r > 0.6) and similarity in results seen for PFOS and 

Perfluorooctanoic acid (PFOA) – May 2016 3-52
PFOA. Additionally, many of the C8 studies do not appear to have controlled for the impact of 

diet on serum lipids.

Liver disease and liver function. Few studies of the relationship between PFOA and liver 

disease are available, but the C8 Health Project did not observe associations with hepatitis, fatty 

liver disease, or other types of liver disease. In the studies of PFOA exposure and liver enzymes 

(measured in serum), positive associations were seen. The results of the occupational studies 

provide evidence of an association with increases in serum AST, ALT, and GGT, with the most 

consistent results seen for ALT. The associations were not large and might depend on the 

covariates in the models, including BMI, use of lipid lowering medications, and triglycerides 

(Costa et al. 2009; Olsen et al. 2000, 2003; Olsen and Zobel 2007; Sakr et al. 2007a, 2007b). 

Two population-based studies of highly exposed residents in contaminated regions near a 

fluorochemical industry in West Virginia have evaluated associations with liver enzymes, and 

the larger of the two studies reported associations of increasing serum ln ALT and ln GGT levels 

with increasing serum PFOA concentrations (Emmett et al. 2006; Gallo et al. 2012). A crosssectional analysis of data from the NHANES, representative of the U.S. national population, also 

found associations with ln PFOA concentration with increasing serum ALT and ln GGT levels. 

Serum bilirubin was inversely associated with serum PFOA in the occupational studies. A 

U-shaped exposure-response pattern for serum bilirubin was observed among the participants in 

the C8 Health Project, which might explain the inverse associations reported for occupational 

cohorts. Overall, an association of serum PFOA concentration with elevations in serum levels of 

ALT and GGT has been consistently observed in occupational, highly exposed residential 

communities, and the U.S. general population. The associations are not large in magnitude, but 

indicate the potential of PFOA to affect liver function.

Immune function. Associations between prenatal, childhood, or adult PFOA exposure and risk 

of infectious diseases (as a marker of immune suppression) have not been consistently seen, 

although there was some indication of effect modification by gender (i.e., associations seen in 

female children but not in male children) (Fei et al. 2010a; Granum et al. 2013; Looker et al. 

2014; Okada et al. 2012). Three studies have examined associations between maternal and/or 

child serum PFOA levels and vaccine response (measured by antibody levels) in children 

(Grandjean et al. 2012; Granum et al. 2013) and in adults (Looker et al. 2014). The study in 

adults was part of the high-exposure community C8 Health Project. A reduced antibody response 

to one of the three influenza strains tested after subjects received the flu vaccine was seen with 

increasing levels of serum PFOA; these results were not seen with PFOS. The studies in children 

were conducted in general populations in Norway and in the Faroe Islands. Decreased vaccine 

response in relation to PFOA levels was seen in these studies, but similar results also were seen 

with correlated PFASs (e.g., PFOS). 

Thyroid. Three large studies provide support for an association between PFOA exposure and 

incidence or prevalence of thyroid disease in women or children, but not in men (LopezEspinosa et al. 2012; Melzer et al. 2010; Winquist and Steenland 2014b). In addition, 

associations between PFOA and TSH were seen in pregnant females with anti-TPO antibodies 

(Webster et al 2014). In contrast, generally null associations were found between PFOA and 

TSH in people who had not been diagnosed with thyroid disease. 

Diabetes. No associations were observed between serum PFOA levels and type II diabetes 

incidence rate in general or worker populations with mean serum PFOA up to 0.0913–0.113 

μg/mL (MacNeil et al. 2009; Steenland et al. 2015). PFOA was not associated with measures of 

metabolic syndrome in adolescents or adults (Lin et al. 2009). However, one study found an 

Perfluorooctanoic acid (PFOA) – May 2016 3-53
increased risk for developing gestational diabetes in females with mean serum PFOA (measured 

preconception) of 0.00394 μg/mL (Zhang et al. 2015). 

Fertility, pregnancy, and birth outcomes. There are no occupational exposure or general 

population studies examining pregnancy-related hypertension and preeclampsia in relation to 

PFOA exposure. The only data available come from the high-exposure C8 Health Project study 

population. Several studies, using different designs and exposure measures, have examined that

outcome in this population (Darrow et al. 2013; Savitz et al. 2012a, 2012b; Stein et al. 2009). 

There is a progressively greater refinement and reduction in misclassification (or exposure and 

outcome) among this set of studies. Each of the studies provides some evidence of an association 

between PFOA exposure and risk of pregnancy-induced hypertension or preeclampsia, with the 

most robust findings from the methodologically strongest study (Darrow et al. 2013). 

The association between PFOA and birth weight was examined in numerous studies. Most 

studies measured PFOA using maternal blood samples taken in the second or third trimester or in 

cord blood samples. Studies on the high-exposure C8 community population did not observe 

associations between PFOA and either birth weight among term births or the risk of low birth 

weight among all (singleton) births (Darrow et al. 2013; Nolan et al. 2009; Savitz et al. 2012a, 

2012b; Stein et al. 2009). In contrast, several analyses of general populations indicate a negative 

association between PFOA levels and birth weight (Apelberg et al. 2007; Fei et al. 2007; 

Maisonet et al. 2012), while others did not attain statistical significance (Chen et al. 2012; Hamm 

et al. 2010; Monroy et al. 2008; Washino et al. 2009). A meta-analysis of many of these studies 

found a mean birth weight reduction of 19 g (95% CI: -30, -9) per each one unit (ng/mL) 

increase in maternal or cord serum PFOA levels (Johnson et al. 2014). It has been suggested that 

GFR can impact birth weight (Morken et al. 2014). Verner et al (2015) conducted a metaanalysis based on PBPK simulations and found that some of the association reported between 

PFOA and birth weight is attributable to GFR and that the actual association could be closer to a 

7-g reduction (95% CI: -8, -6). Verner et al. (2015) showed that, in individuals with low GFR, 

there are increased levels of serum PFOA and lower birth weights. While there is some 

uncertainty in the interpretation of the observed association between PFOA and birth weight 

given the potential impact of low GFR, the available information indicates that the association

between PFOA exposure and birth weight for the general population cannot be ruled out. In 

humans with low GFR (which includes females with pregnancy-induced hypertension or 

preeclampsia), the impact on body weight is likely due to a combination of the low GFR and the 

serum PFOA.

Two studies examined development of puberty in females in relation to prenatal exposure to 

PFOA as measured through maternal or cord blood samples in follow-up of pregnancy cohorts 

conducted in England (Christensen et al. 2011) and in Denmark (Kristensen et al. 2013). The 

results of these two studies are conflicting, with no association (or possible indication of an 

earlier menarche seen with higher PFOA) in Christensen et al. (2011), and a later menarche seen 

with higher PFOA in Kristensen et al. (2013). Another study examined PFOA exposure 

measured concurrently with the assessment of pubertal status (Lopez-Espinosa et al. 2011). An 

association between later age at menarche and higher PFOA levels was observed, but the 

interpretation of this finding is complicated by the potential effect of puberty on the exposure 

biomarker levels (i.e., reverse causality).

Perfluorooctanoic acid (PFOA) – May 2016 3-54
Studies found a positive association with ADHD in children in the highly exposed 

community (Stein et al. 2013) and the general population (Hoffman et al. 2010). No other 

behavior endpoints in children were associated with maternal PFOA levels in either population.

Limited data suggest a correlation between higher PFOA levels (>0.02 μg/mL) in females

and decreases in fecundity and fertility (Fei et al. 2009; Vélez et al. 2015), but there are no clear 

effects of PFOA on male fertility endpoints (0.0035–0.005 μg/mL) (Joensen et al. 2009, 2013). 

C8 Science Panel conclusions. As part of the C8 Health Project, the C8 Science Panel used 

epidemiological and other data available to them to assess probable links between PFOA 

exposure and disease (C8 Science Panel 2012). Analyses conducted by the C8 Science Panel 

used historical serum PFOA estimates over time, which were developed based on estimated 

intake of contaminated drinking water. The panel concluded that a probable link existed between 

PFOA exposure and ulcerative colitis, high cholesterol, pregnancy-induced hypertension, and 

thyroid disease. 

The C8 Science Panel found no probable link between PFOA exposure and multiple other 

conditions, including birth defects, other autoimmune diseases (e.g., rheumatoid arthritis, lupus, 

type 1 diabetes, Crohn’s disease, MS), type II diabetes, high blood pressure, coronary artery 

disease, infectious disease, liver disease, Parkinson’s disease, osteoarthritis, neurodevelopmental 

disorders in children (e.g., ADHD, learning disabilities), miscarriage or stillbirth, chronic kidney 

disease, stroke, asthma or COPD, and preterm birth or low birth weight (C8 Science Panel 2012).

3.1.2 Cancer

Occupational exposure studies. Several occupational studies examining cancer mortality have 

been conducted at 3M’s Cottage Grove facility in Minnesota and at the DuPont Washington 

Water Works plant in West Virginia. These studies have focused on kidney, bladder, liver, 

pancreatic, testicular, prostate, thyroid, and breast cancers. For cancers with a high survival rate 

(i.e., bladder, kidney, prostate, testicular, thyroid, and breast cancer), studies that use mortality 

data provide a more limited basis for drawing conclusions than studies that use incidence data. 

The discussion in this section summarizes the design and results of the available studies, 

focusing on the most recent update of occupational cohorts. Table 3-12 presents results for 

studies of kidney and testicular cancer. 

Raleigh et al. (2014) is the latest update of the analyses of mortality in the 3M Cottage Grove 

workers, previously analyzed in Lundin et al. (2009) and Gilliland and Mandel (1993). Raleigh 

et al. (2014) followed 4,668 Cottage Grove workers through 2008, using an improved exposure 

reconstruction method and adding a nonexposed worker referent group from a different 3M 

plant. In addition to the mortality data, incidence data based on state cancer registries also were 

included. Exposure estimates for inhalation exposures were calculated from work history records 

and industrial hygiene monitoring data; blood levels were not included. No associations were 

found between PFOA exposure and the risk of dying from any cancer type (see Table 3-12 for 

bladder, kidney, and testicular cancer results). The mean age of the workers was 29 years at the 

start of employment and 63 years at the end of follow-up. 

Perfluorooctanoic acid (PFOA) – May 2016 3-55
Table 3-12. Summary of PFOA Epidemiology Studies of Kidney and Testicular Cancer

Reference and Study Details Analysis Group Kidney Testicular

Occupational Settings

Raleigh et al. 2014

3M, Minnesota

n = 4,668, follow-up through 2008

Mean age: 29 yrs at start of 

employment

Mortality and incidence

Comparison based on another (nonPFOA) 3M plant in Minnesota (n = 

4,359)

Cumulative exposure level based on 

industrial hygiene data (air 

monitoring), and PFOA production 

levels

[update of Gilliland and Mandel 1993 

and Lundin et al. 2009]

All (Minnesota referent)

By quartile, mortality 

analysis

1 up to 0.000026 μg/m3-yr

2 up to 0.00014

3 up to 0.00073

4 maximum not reported

By quartile, incidence 

analysis

1 up to 0.000029 μg/m3-yr

2 up to 0.00015

3 up to 0.00079

4 maximum not reported

0.53 (0.20, 1.16) (n = 6)

Mortality

0.32 (0.01, 1.77) (n = 1)

0.74 (0.09, 2.69) (n = 2)

1.66 (0.08, 2.38) (n = 2)

0.42 (0.01, 2.34) (n = 1)

Incidence

1.07 (0.36, 3.16) (n = 4)

1.07 (0.36, 3.17) (n = 4

0.98 (0.33, 2.92) (n = 4)

0.73 (0.21, 2.48) (n = 4)

n = 5 incident cases 

reported; no further 

analysis done

Steenland and Woskie 2012

DuPont Washington Water Works, 

West Virginia

n = 5,791, follow-up through 2007

Mortality

[update of Leonard et al. 2008, which 

did not include analysis by cumulative 

exposure]

[Steenland and Woskie 2012 examined 

incidence of bladder, colorectal, and 

prostate cancers, and of melanoma]

DuPont referent (plants from 

8 surrounding states)

U.S. referent

Cumulative exposure (ppmyrs)

0 - < 904

904 - < 1,520

15,20 - < 2,720

≥ 2720

1.28 (0.66, 2.24) (n = 

12)

1.09 (0.56, 1.90) (n = 

12)

1.07 (0.02, 3.62) (n = 1)

1.37 (0.28, 3.99) (n = 3)

 -- (0.0, 1.42) (n = 0)

2.66 (1.15, 5.24) (n = 8)

1.80 (0.05, 10.03) (n 

= 1)

High-Exposure Community

Vieira et al. 2013

C8 Health Project population (Ohio and 

West Virginia)

Incidence

Modeled estimates for 1951–2008 

using residence at time of diagnosis 

and emissions data and environmental 

characteristics

Total for 6 water districts 

(median serum level ranged 

from 5 to 125 μg/l)

Annual serum levels (μg/l); 

assumed 10-year residence 

and 10-year latency (Ohio)

 Unexposed

 Low: 3.7 - 12.

 Medium: 12.9 - 30.7

 High: 30.8 – 109

 Very high: > 100

1.1 (0.9, 1.4) (n = 94)

1.0 (referent)

0.8 (0.4, 1.5) (n = 11)

1.2 (0.7, 2.0) (n = 17)

2.0 (1.3, 3.2) (n = 22)

2.0 (1.0, 3.9) (n = 9)

(0.6, 1.8) (n = 18)

(referent)

0.2 (0.0, 1.6) (n = 1)

0.6 (0.2, 2.2) (n = 3)

0.3 (0.0, 2.7) (n = 1)

2.8 (0.8, 9.2) (n = 6)

Barry et al. 2013

C8 Health Project population (Ohio and 

West Virginia)

Case-control (n varies by cancer)

Incidence

Modeled estimates for 1951–2008 

using individual-level data on 

residential history, drinking water 

source, tap water consumption, 

emissions data, environmental 

characteristics, water pipe installation, 

PK data, and workplace water 

consumption (and for workers, 

workplace exposure based on job 

exposure matrix and modeling using 

serum samples from 1979–2004 and 

job history data. 

Full sample

Cumulative exposure, 

quartiles (cutpoints based on 

cancer-specific case 

distribution; approximate 

midpoints)

1 (30-50 μg/mL-yr)

2 (90-200 μg/mL-yr)

3 (800-1400 μg/mL-yr)

4 (100,000 μg/mL-yr)

1.09 (0.97, 1.21) (n = 

105)

(referent)

0.99 (0.53, 1.85)

1.69 (0.3, 3.07)

1.43 (0.76, 2.69)

trend p = 0.34

community cohort; HR 

= 1.0, 0.94, 1.08,1.50, 

trend p = 0.02;

worker cohort HR = 1.0, 

1.22, 3.27, 0.99, trend p 

= 0.42

1.28 (0.95, 1.73) (n = 

17)

1.0 (referent)

0.87 (0.15, 4.88)

1.08 (0.20, 5.90)

2.36 (0.41, 13.7)

trend p = 0.02

15 of the cases from 

the community 

sample; HR = 1.0, 

0.98, 1.54, 4.66, 

trend p = 0.02

Perfluorooctanoic acid (PFOA) – May 2016 3-56
Steenland and Woskie (2012) updated the cohort study by Leonard et al. (2008) of 

employees at the DuPont Washington Works plant in West Virginia (see Table 3-12 for bladder, 

kidney, and testicular cancer results). This study included 5,791 individuals who had worked at 

the DuPont West Virginia plant for at least 1 year between 1948 and 2002. Mean duration of 

employment was 19 years. Deaths through 2008 were ascertained through either the National 

Death Index or death certificate data. Exposure quartiles were assessed by estimated cumulative 

annual serum levels based on blood samples from 1,308 workers taken during 1979–2004 and 

time spent in various job categories (ppm-years). Referent groups included both nonexposed 

DuPont workers in the same region and the U.S. population. Overall, the mean cumulative 

exposure was 7.8 ppm-years and the estimated average annual serum level was 0.35 μg/mL. A 

significant positive trend was found for kidney cancer with the SMR=2.66 (n = 8; 95% CI 1.15, 

5.24) for workers in the highest quartile. The most recent report on the same cohort included 

6,026 workers evaluated for disease incidence, based on self-report with validation from medical 

records (Steenland et al. 2015). Lifetime serum cumulative dose was estimated by combining 

occupational and nonoccupational exposures. Median measured serum level was 0.113 μg/mL 

based on samples collected in 2005. Bladder cancer incidence (n = 29 cases) decreased with 

increased PFOA levels (RR 1.0, 0.55, 0.47, and 0.31 across quartiles, trend p = 0.03). Prostate 

cancer risk increased in Q1 compared to Q2 (n = 1.92), and remained at this level in the 

remaining quartiles (RR 1.89 and 2.15 in Q3 and Q4, respectively, trend p = 0.10). 

Cholecystokinin (CCK) is a peptide hormone that stimulates the digestion of fat and protein, 

causes the increased production of hepatic bile, and stimulates contraction of the gall bladder. 

Research in rats suggests that pancreas acinar cell adenomas observed in rodents might be the 

result of increased CCK levels secondary to blocked bile flow (Obourn et al. 1997). CCK was 

measured in male workers (n = 74 males) at the 3M’s Cottage Grove plant in 1997 as part of the 

medical surveillance program (Olsen et al. 1998, 2000). Employees’ serum PFOA levels were 

stratified into three categories (<1, 1– <10, and ≥10 ppm). The mean CCK values for the three 

PFOA categories were 33.4, 28.0, and 17.4 pg/mL, respectively. The means in the two serum 

categories < 10 ppm were at least 50% higher than in the ≥ 10 ppm category. A statistically 

significant negative association between mean CCK levels and the three PFOA categories was 

observed (p = 0.03). A multiple regression model of the natural log of CCK and serum PFOA 

levels continued to display a negative association after adjusting for potential confounders. As 

stated previously (Olsen et al. 2000), no abnormal liver function, hypolipidemia, or cholestasis 

was observed in the workers. The authors suggested that the lack of a positive association 

between PFOA and CCK in workers could have resulted from serum PFOA levels too low to 

cause an increase in CCK provided that the same mechanism that increases CCK levels in 

rodents exists in humans. 

High-exposure community studies. Vieira et al. (2013) investigated the relationship between 

PFOA exposure and cancer among the residents living near the DuPont plant in Parkersburg, 

West Virginia. This analysis included incident cases of 18 cancers diagnosed from 1996–2005 in 

five Ohio counties and eight West Virginia counties that included public water districts 

contaminated with PFOA. The dataset included 7,869 cases from Ohio geocoded to residence 

and 17,238 cases from West Virginia linked to water district. Exposure levels and serum PFOA 

concentrations were estimated based on residence at time of diagnosis, using modeled data based 

on previous work in the C8 study population (Shin et al. 2011). Individual-level exposure was 

categorized as very high, high, medium, low, or unexposed based on serum concentrations of 

>0.110 μg/mL, 0.0308–0.109 μg/mL, 0.0129–0.0307 μg/mL, 0.0037–0.0129 μg/mL, and 

unexposed (background levels not given), respectively. Logistic regression was applied to 

Perfluorooctanoic acid (PFOA) – May 2016 3-57
individual-level data to calculate ORs and CIs for each cancer category. Data were first analyzed 

by water district. The adjusted ORs were increased for testicular cancer and for kidney cancer 

(OR: 5.1, 95% CI: 1.6, 15.6; n = 8 and OR: 1.7, 95% CI: 0.4, 3.3; n = 10, respectively) in the 

Little Hocking water district and for kidney cancer (OR: 2.0, 95% CI: 1.3, 3.1; n = 23) in the 

Tuppers Plains water district. Both districts are in Ohio. Residents of Little Hocking also had 

increased OR for non-Hodgkin lymphoma (OR: 1.6, 95% CI: 0.9, 2.8; n = 14) and prostate 

cancer (OR: 1.4, 95% CI: 0.9, 2.3; n = 36). The analysis by exposure level for kidney and 

testicular cancers is shown in Table 3-12. Kidney cancer was positively associated with very 

high and high exposure categories (OR: 2.0, 95% CI: 1.0, 3.9; n = 9 and OR: 2.0, 95% CI: 1.3, 

3.2; n = 22, respectively), while ORs for medium and low exposure categories were close to the 

null when compared to the unexposed category. The largest OR was for testicular cancer with the 

very high exposure category (OR: 2.8, 95% CI: 0.8, 9.2; n = 6), but the estimate was imprecise 

because of the small numbers. ORs for the other exposure categories were all <1.0. Ovarian 

cancer, non-Hodgkin’s lymphoma, and prostate cancer were positively associated with the very 

high exposure category, but showed weaker or negative associations for the other exposure 

categories (Vieira et al. 2013).

Barry et al. (2013) extended the study of cancer incidence in the C8 Health Project 

population in an analysis of data from 32,254 study participants; there is some overlap in the 

cases included in Vieira et al (2013) and in Barry et al. (2013). The cohort included 3,713 current 

and former DuPont Washington Works employees, but results for this subset were limited by the 

small sample size for cancers of interest. Median serum PFOA levels, measured in 2005–2006 at 

enrollment in C8, were 0.024 and 0.113 μg/mL for community and worker populations, 

respectively. A proportional hazard regression model was run for each cancer type with the 

cancer as the outcome, time-varying cumulative PFOA serum concentration as the independent 

variable, and age as the time scale. Cumulative PFOA serum concentrations were estimated 

based on historical regional monitoring data and individual residential histories. Self-reported 

cancers were validated through a cancer registry or medical record. Confounders included 

smoking, alcohol consumption, gender, education, and 5-year birth year period. Testicular cancer 

risk was significantly increased with an increase in the log of estimated cumulative PFOA serum 

level (HR: 1.34, 95% CI: 1.00, 1.79; n = 17). Using estimated cumulative PFOA serum 

concentration quartiles, a significant monotonic trend was found for testicular cancer. Slight 

nonsignificant increases were seen for kidney cancer (HR: 1.10, 95% CI: 0.98, 1.24; n = 105) 

and for thyroid cancer (HR: 1.10, 95% CI: 0.95, 1.26; n = 86) (Barry et al. 2013). 

Members of the C8 Health Project were evaluated for an association between serum PFOA 

levels and incidence of colon or rectal cancer (Innes et al. 2014). This cross-sectional study 

compared serum PFOA (and PFOS) levels at enrollment with diagnosis of primary colorectal 

cancer; 47,151 cancer-free adults and 203 cases were included. Serum PFOA levels ranged from 

<0.0005 to 22.4 μg/mL, with an average of 0.0866 μg/mL. An inverse relationship was found 

between PFOA level and diagnosis of colorectal cancer with OR = 0.64 (95% CI 0.44, 0.94; 

highest to lowest quartile, p for trend = 0.002). A concentration-related inverse association also 

was found between PFOS and colorectal cancer.

In 2012, the C8 Science Panel concluded that there is a probable link between exposure to 

PFOA and testicular and kidney cancer, but no other types of cancers. Their conclusion was 

based on the studies presented above, other epidemiology studies on cancer incidence in the midOhio population, worker cohorts, and published data. Panel studies addressed 21 different 

Perfluorooctanoic acid (PFOA) – May 2016 3-58
categories of cancer and looked for positive trends with increasing exposure as measured by 

cumulative serum levels.

General population studies. Eriksen et al. (2009) examined the association between plasma 

PFOA concentration and the risk of cancer in the general Danish population. The study 

population was chosen from individuals (aged 50–65 years) who had enrolled in the prospective 

Danish cohort Diet, Cancer and Health study between December 1, 1993, and May 31, 1997. The 

Danish Cancer Registry and Danish Pathology Data Bank were used to identify cancer patients 

diagnosed between December 1, 1993, and July 1, 2006. The cancer patients (n = 1,240) 

consisted of 1,111 males and 129 females whose median age was 59 years and who had prostate 

cancer (n = 713), bladder cancer (n = 332), pancreatic cancer (n = 128), or liver cancer (n = 67). 

The individuals (n = 772) in the subcohort comparison group were randomly chosen from the 

cohort study and consisted of 680 males and 92 females whose median age was 56 years. The 

participants each answered a questionnaire upon enrollment in the cohort study, and data on 

known confounders were obtained from the questionnaires. The plasma PFOA concentrations, 

based on blood samples provided by cancer patients at enrollment (1993–1997) were as follows: 

males 0.0068 μg/mL, females 0.0060 μg/mL, prostate cancer 0.0069 μg/mL, bladder cancer 

0.0065 μg/mL, pancreatic cancer 0.0067 μg/mL, and liver cancer 0.0054 μg/mL. The plasma 

PFOA concentrations for the subcohort comparison group were 0.0069, 0.0054, and 0.0066 

μg/mL for males, females, and combined, respectively. IRRs, crude and adjusted for 

confounders, did not indicate an association between plasma PFOA concentration and prostate, 

bladder, pancreatic, or liver cancer (see Table 3-12 for bladder cancer results). The plasma 

PFOA levels in the population were lower than those observed in occupational cohorts. This 

study is novel in that it is the first to examine PFOA levels and cancer in the general population. 

A subset of females enrolled in the Danish National Birth Cohort was evaluated for an 

association between plasma PFOA levels (as well as 15 other PFASs) measured during 

pregnancy and risk of breast cancer during a follow-up period of 10–15 years (BonefeldJørgensen et al. 2014). A total of 250 females diagnosed with breast cancer were matched for age 

and parity with 233 controls. The mean PFOA level in the controls was 0.0052 μg/mL while 

levels in the cases were divided into quintiles ranging from <0.0037 up to >0.0065 μg/mL. No 

association was found between PFOA levels and breast cancer risk. A weak positive association 

was found only with perfluorooctane sulfonamide. 

Hardell et al. (2014) investigated an association between prostate cancer and levels of 

perfluoroalkyl acids (PFAAs) in whole blood. Patients with newly diagnosed prostate cancer 

(n = 201) had median PFOA levels of 0.002 μg/mL while the case-control group (n = 186) had a 

median level of 0.0019 μg/mL. PFOA levels were not associated with higher risks of prostate 

cancer when compared to controls or when analyzed according to Gleason score (pathology 

grade) and prostate-specific antigen. A significantly higher risk for prostate cancer was found for 

PFOA levels above the median combined with a first-degree relative with prostate cancer, 

indicating a genetic risk factor.

Two studies found no differences in blood and tissue PFOA levels between cancer and 

noncancer patients; the types of cancer in the patients were not defined. Vassiliadou et al. (2010) 

found that median serum PFOA concentrations among 40 cancer patients (0.00227 μg/mL in 

males; 0.00185 μg/mL in females) were similar to two control groups (0.00314 and 0.00181 

μg/mL in males; 0.0017 and 0.00171 μg/mL in females). Yeung et al. (2013) found similar 

PFOS levels in serum and liver tissue between controls and those with hepatocellular carcinoma. 

Median serum levels in controls (n = 25) and patients with liver cancer (n = 24) were 0.00234 

Perfluorooctanoic acid (PFOA) – May 2016 3-59
and 0.0025 μg/mL, respectively, and liver tissue were 0.506 (n = 9) and 0.495 (n = 12) ng/g, 

respectively.

3.1.2.1 Summary and Conclusions from the Human Cancer Epidemiology Studies

Evidence of carcinogenic effects of PFOA in epidemiology studies is based on studies of 

kidney and testicular cancer. These cancers have relatively high 5-year survival rates of 73% for 

kidney cancer and 95% for testicular cancer (based on National Cancer Institute [NCI] 

Surveillance, Epidemiology, and End Results data for 2005–2011). Thus studies that examine 

cancer incidence are particularly useful for these types of cancer. The high-exposure community 

studies also have the advantage for testicular cancer of including the age period of greatest risk, 

as the median age at diagnosis is 33 years. The two occupational cohorts in Minnesota and West 

Virginia (most recently updated, respectively, in Raleigh et al. 2014 and Steenland and Woskie 

2012) do not support an increased risk of these cancers, but each of them is limited by a small 

number of observed deaths and incident cases. Two studies involving members of the C8 Health 

Project showed a positive association between PFOA levels (mean at enrollment of 0.024 

μg/mL) and kidney and testicular cancers (Barry et al. 2013; Vieira et al. 2013). There is some 

overlap in the cases included in these studies. None of the general population studies examined 

kidney or testicular cancer, but no associations were found in the general population between 

mean serum PFOA levels up to 0.0866 μg/mL and colorectal, breast, prostate, bladder, or liver 

cancer (Bonefeld-Jørgensen et al. 2014; Eriksen et al. 2009; Hardell et al. 2014; Innes et al. 

2014).

As part of the C8 Health Project, the C8 Science Panel (2012) concluded that a probable link 

existed between PFOA exposure and testicular and kidney cancer.

A group of independent toxicologists and epidemiologists critically reviewed the 

epidemiological evidence for cancer based on 18 studies of occupational exposure to PFOA and 

general population exposure with or without coexposure to PFOS. The project was funded by 

3M, but the company was not involved in the preparation or approval of the report. The authors 

evaluated the published studies based on the study design, subjects, exposure assessment, 

outcome assessment, control for confounding, and sources of bias. They followed the Bradford 

Hill guidelines on the strength of the association, consistency, plausibility, and biological 

gradient in reaching their conclusion. They found a lack of concordance between community 

exposures and occupational exposures one or two magnitudes higher than those for the general 

population. The discrepant findings across the study populations were described as likely due to 

chance, confounding, and/or bias (Chang et al. 2014). 

3.2 Animal Studies

Acute and short-term studies in monkeys, rats, and mice provide data on systemic toxicity 

and MoA. Subchronic studies in monkeys and rats found decreased body weight, increased liver 

weight accompanied by microscopic lesions, and decreased serum cholesterol. The most 

prominent microscopic lesion of the liver in both monkeys and rats was centrilobular 

hepatocellular hypertrophy. Data from studies of inhalation and dermal exposures are limited.

Chronic exposure studies were conducted in monkeys, rats, and mice providing information 

on tumor incidences for both rats and mice. Effects on development and reproduction were found 

in both rats (a 2-generation study) and mice (male fertility) and included developmental delays 

and increased neonatal mortality. Many developmental studies focused on the impact of 

Perfluorooctanoic acid (PFOA) – May 2016 3-60
gestational/lactational exposure on mammary gland development and effects observed in 

offspring at maturity.

3.2.1 Acute Toxicity

Oral Exposure 

Dean and Jessup (1978) reported an oral lethal dose for 50% of animals (LD50) of 680 mg/kg 

and 430 mg/kg PFOA for male and female CD rats, respectively. Glaza (1997) reported an oral 

LD50 of greater than 500 mg/kg in male Sprague-Dawley rats and between 250 and 500 mg/kg in 

females. Gabriel (1976a) reported an oral LD50 of less than 1,000 mg/kg for male and female 

Sherman-Wistar rats. According to the Hodge Sterner Scale, these LD50 values suggest that 

PFOA can be classified as moderately toxic after acute oral exposures.

Rigden et al. (2015) exposed groups of five male Sprague-Dawley rats to doses of 0, 10, 33, 

and 100 mg/kg/day for 3 days and maintained them for 4 additional days with daily body weight 

measurement and overnight collection of urine. Following the recovery period, the animals were 

sacrificed with collection of serum samples for analysis. Major organs were weighed, and the 

liver homogenized. The serum samples, liver homogenate, and supernatant were kept frozen at -

80°C until they were analyzed. Phase I and II drug metabolizing enzymes and palmitoylcoenzyme A (-CoA) oxidase were measured in the liver homogenate. Urine was analyzed for 

malondialdehyde (MDA) and 8-hydroxydeoxyguanine. The results for PFOA were compared 

with those for 100-mg/kg/day doses of di(2-ethylhexyl) phthalate (DEHP) and fenofibrate, 

known inducers of PPARα.

There was a dose-related statistically significant increase in palmitoyl-CoA oxidase and liver 

weight at all PFOA doses. The palmitoyl-CoA increase was not significant for DEHP and 

fenofibrate with 100-mg/kg doses; liver weight increased significantly for fenofibrate but not 

DEHP. The only serum parameter that showed a significant dose-related response with PFOA 

was a decrease in uric acid compared to controls. Serum was analyzed for several minerals, 

proteins, enzymes (e.g., ALP, AST, ALT), glucose, cholesterol, and triglycerides. Phase I drug 

metabolizing enzyme activities (ethoxyresorufin-O-deethylase [EROD] and pentoxyresorufin-Odepentilase [PROD]) were significantly increased at the 100-mg PFOA/kg/day dose, and 

glutathione-S-transferase (GST) activity was significantly decreased at the two highest doses, but 

not in a dose-related fashion. UDP-glucuronyltransferase (UDP-GT) was significantly lower than 

controls at all doses, but the changes did not demonstrate a dose-related response. There were no 

dose-related significant changes for the other analytes. The 10-mg PFOA/kg/day dose 

administered for 3 days was a LOAEL for effects on the liver associated with PPARα activation 

and for a decrease in serum uric acid. PFOA at 10 mg/kg/day for 3 days had a stronger impact on 

liver weight and palmitoyl-CoA activation than 100 mg/kg/day of DEHP and fenofibrate for the 

same exposure duration (Table 3-13). The 10-mg/kg/day dose was a LOAEL for liver effects 

usually associated with PPARα activation. The PFOA response was stronger than that for a 

100-g/kg/day dose for the two known activators of PPAR-α.

Perfluorooctanoic acid (PFOA) – May 2016 3-61
Table 3-13. Comparison of PPAR-α Related Effects in Rats for PFOA, DEHP, and 

Fenofibrate after a 3-day Exposure 

Chemical

Dose 

mg/kg/

day

Liver wt.

g

Palm. CoA

abs/min/g 

prot

EROD

nmol/min/

mg prot

PROD

nmol/min/

mg prot

UDP-GT

nmol/min/

mg prot

GST

nmol/min/

mg prot

Control 0 4.28 ± 0.20 1.02 ± 0.37 0.066 ± 0.022 0.045 ± 0.012 1.69 ± 0.12 1.21 ± 0.11

PFOA 10 5.73 ± 0.29* 3.17 ± 0.65* 0.096 ± 0.024 0.080 ± 0.024 0.88 ± 0.09* 1.11 ± 0.09

PFOA 33 6.40 ± 0.20* 4.89 ± 0.79* 0.080 ± 0.015 0.078 ± 0.031 1.00 ± 0.14* 0.88 ± 0.12*

PFOA 100 6.62 ± 0.47* 6.11 ± 1.51* 0.113 ± 0.025* 0.107 ± 0.029* 1.12 ± 0.20* 0.94 ± 0.19*

DEHP 100 4.14 ± 0.34 1.40 ± 1.09 0.060 ± 0.012 0.039 ± 0.031 1.45 ± 5031 1.27 ± 0.11

Fenofibrate100 5.73 ± 0.24* 1.71 ± 0.58 0.060 ± 0.016 0.046 ± 0.020 1.09 ± 0.14* 0.94 ± 0.13*

Notes: Mean ±SD

* Significant ANOVA followed by Dunnett post-hoc test p < 0.05. 

Inhalation Exposure 

Rusch (1979) reported no mortality in male or female Sprague-Dawley rats following 

inhalation exposure to 186,000 mg/m3 PFOA for 1 hour. Kennedy et al. (1986) reported a 4-hour 

lethal concentration for 50% of animals (LC50) of 980 mg/m3

 for groups of six male rats exposed 

to PFOA as a dust in air. As reported in a later publication (Kennedy et al. 2004), body weight 

loss, irregular breathing, and red discharge around the nose and eyes were observed. Corneal 

opacity and corrosion were seen at concentrations greater than or equal to 810 mg/m3. 

Dermal/Ocular Exposure 

The dermal LD50 in New Zealand White rabbits was determined to be greater than 

2,000 mg/kg (Glaza 1995). Kennedy (1985) determined a dermal LD50 of 4,300 mg/kg for 

rabbits, 7,000 mg/kg for male rats, and 7,500 mg/kg for female rats. The animals lost body 

weight and exhibited lethargy, labored breathing, diarrhea, and severe skin irritation (Kennedy et 

al. 2004). PFOA is an ocular irritant in rabbits when the compound is not washed from the eyes 

(Gabriel 1976b), but is not an irritant in rabbits when washed from the eye (Gabriel 1976c). 

Markoe (1983) found PFOA to be a skin irritant in rabbits, while Gabriel (1976d) did not 

conclude that PFOA is a skin irritant. 

3.2.2 Short-Term Studies

Oral Exposure 

Monkey. In a range-finding study, Thomford (2001) administered PFOA to male cynomolgus 

monkeys as an oral capsule containing 0, 2, and 20 mg/kg/day PFOA for 4 weeks. There were 

three monkeys in the 2- and 20-mg/kg/day groups and one monkey in the control group. Animals 

were observed twice daily for mortality and moribundity and were examined at least once daily 

for signs of poor health or abnormal behavior. Body weights were recorded weekly and food 

consumption was assessed qualitatively. The monkeys were fasted overnight and blood samples 

were collected 1 week prior to the start of the study and on day 30 for measurement of serum 

PFOA, clinical hematology, and clinical chemistry, plus analysis for hormones (estradiol, 

estrone, estriol, TSH, total and FT3, and total and FT4). Blood samples also were collected from 

each animal on day 2 (approximately 24 hours after the first dose) for clinical chemistry 

measurements.

Perfluorooctanoic acid (PFOA) – May 2016 3-62
At scheduled necropsy, liver samples were collected for palmitoyl-CoA oxidase activity (a 

biomarker for peroxisome proliferation) and serum PFOA. Liver, testes, and pancreas were 

collected and assayed for cell proliferation using antibodies to proliferating cell nuclear antigen

(PCNA). Bile was collected from each animal for measurement of bile acid. The adrenals, liver, 

pancreas, spleen, and testes from each animal were examined microscopically.

All animals survived to scheduled sacrifice. There were no clinical signs of toxicity in the 

treated groups and there was no effect on body weight. Low or no food consumption was 

observed for one animal given 20 mg/kg/day. There were no effects on the hormones measured 

with the exception of estrone, which was notably lower in the 2- and 20-mg/kg/day PFOA 

groups. There was no evidence of peroxisome proliferation or cell proliferation in the liver, 

testes, or pancreas of the treated monkeys. No adverse effects were noted in either the gross or 

clinical pathology evaluations. Under the conditions of this study, the NOAEL was 20 mg/kg and 

no LOAEL was established.

Rat. Pastoor et al. (1987) dosed male Crl:CD (SD) BR rats (n = 6 per group) for 1, 3, and 7 days 

with 0 and 50 mg PFOA/kg. Liver sections were collected at necropsy and stained with 

hematoxylin and eosin. Sections also were examined by electron microscopy. DNA content was 

also determined for the livers of rats dosed for 7 days. Treatment with 50 mg PFOA/kg for 7 

days caused a 17% decrease (p<0.05) in mean body weight. Pair-fed control rats had a 24% 

decrease in body weight. Body weight was no different in the rats treated for 1 and 3 days than in 

the control rats. Liver weight of rats treated for 1 day was no different than control liver weight.

The relative liver weight of rats treated for 3 days was significantly increased (p<0.05) compared 

to control relative liver weight. Absolute and relative liver weights were significantly increased 

(p<0.05) after the 7-day treatment with PFOA. A 57% decrease (p<0.05) was observed in 

relative hepatic DNA/g liver, but no difference was observed between total amount of hepatic 

DNA/liver and total amount of DNA/liver in control rats.

The hepatocytes of rats treated with PFOA for 3 days were enlarged with partially occluded 

sinusoids, and had numerous basophilic granules, eosinophilic granular material in the 

cytoplasm, and fewer perinuclear glycogen vacuoles compared to control hepatocytes. Enlarged 

hepatocytes with hyperplastic smooth endoplasmic reticulum (ER), increased mitochondria, 

increased peroxisomes, decreased rough ER, and increased autophagosomes with electron-dense 

material also were observed in the hepatocytes.

Loveless et al. (2008) administered 0, 0.3, 1, 10, and 30 mg linear PFOA/kg by oral gavage 

to groups of male CD rats (n = 10 per group) for 29 days. Body weight was recorded on days 0, 

3, and 6–28. At necropsy, blood was collected for hematology, clinical chemistry, and 

corticosterone (CORT) measurements. Tissues were collected for weight and microscopic 

examination. Body weight, weight gain, hematocrit, and hemoglobin were reduced at ≥10 mg 

PFOA/kg/day. Increased reticulocytes and hematopoieses were observed in the rats dosed with 

30 mg PFOA/kg/day. Total and non-HDL cholesterol were significantly reduced at 0.3 and 

1 mg/kg/day compared to control. HDL cholesterol was significantly decreased at 0.3, 1, and 

10 mg/kg/day. Triglyceride levels were significantly decreased at all doses except 1 mg/kg.

Absolute liver weight (≥1 mg/kg/day) and relative liver weight (≥10 mg/kg/day) were 

significantly increased. Hepatocellular hypertrophy was graded as minimum to mild 

(0.3–1 mg/kg/day) and moderate (≥10 mg/kg/day), and focal necrosis was present at doses 

≥ 10 mg/kg/day. Although not statistically significant, serum CORT was increased at 

≥ 10 mg/kg/day. The decrease in cholesterol and triglycerides at the lowest dose are not 

necessarily adverse. The 1-mg/kg/day dose is classified as the NOAEL and the 10-mg/kg/day 

Perfluorooctanoic acid (PFOA) – May 2016 3-63
dose as the LOAEL based on the observations of increased liver weight, hepatocellular 

hypertrophy, and hepatic necrosis at that dose. Data on several immunological endpoints were 

reported as part of the Loveless et al. (2008) publication. The immunological data from that

study are included in section 3.3.2 of this report. 

Cui et al. (2009) exposed male Sprague-Dawley rats (10 per group) to PFOA (96% active 

ingredient) at 0, 5, and 20 mg/kg/day for 28 days by gavage once daily. The activity of the rats 

was observed over the course of the study. All rats were sacrificed after the final exposure. The 

rats dosed with 5 mg/kg/day exhibited hypoactivity, decreased food consumption, cachexia, and 

lethargy during the third week of the study. Rats dosed with 20 mg/kg/day also exhibited 

sensitivity to external stimuli. The visceral index (i.e., hepatic, renal, gonad weight/animal’s 

body weight) used to evaluate hyperplasia, swelling, or atrophy was significantly increased in the 

treated animals compared to control animals. In the liver, treatment with 5 or 20 mg PFOA/kg 

caused hepatic hypertrophy, fatty degeneration, and acidophilic lesions as well as angiectasis

(gross dilation) and congestion in the hepatic sinusoid or central vein. In the lung, treatment with 

5 or 20 mg PFOA/kg caused pulmonary congestion and focal or diffuse thickened epithelial 

walls. No histopathologic lesions were observed in the kidneys of the low-dose animals, but 

turbidness and swelling in the epithelium of the proximal convoluted tubule were observed at 

20 mg PFOA/kg. Under the conditions of this study, the LOAEL was 5 mg/kg/day based on 

increased visceral indices, and liver and pulmonary lesions; no NOAEL was established.

Male Sprague-Dawley rats (n = 10 per group) were fed diets containing 0 and 300 ppm 

PFOA for 1, 7, and 28 days in two studies (Elcombe et al. 2010). The mean daily intake for study 

1 and study 2 were 19 and 23 mg/kg/d, respectively. A group of rats was fed diets containing 

50 ppm Wyeth 14,643, a PPARα agonist, as a positive control. The animals were observed daily 

and body weights and food consumption were recorded. At necropsy, day 2, day 8, or day 29, the 

organs were weighed, examined for gross pathology and preserved for histopathology. In study 

1, liver DNA content and concentration were determined, and plasma was collected for analysis 

of liver enzymes, cholesterol, triglycerides, and glucose. Hepatic cell proliferation and apoptosis 

also were determined.

In both studies, body weight significantly decreased (p<0.05) after 7 and 28 days on the 

PFOA diet. Body weight was not affected by Wyeth 14,643. Absolute liver weight was 

significantly increased (p<0.05) in rats fed PFOA diets for 7 days in the first study and in rats 

treated for 7 and 28 days in the second study (Table 3-14). The liver-to-body-weight ratio was 

significantly higher in rats fed PFOA diets for 7 and 28 days in both studies. Absolute liver 

weight and liver-to-body-weight ratios were significantly increased in rats fed the Wyeth 14,643 

diet in both studies.

After 1 day of eating the PFOA diet, subjects’ plasma AST was significantly decreased and 

triglycerides were significantly increased. After 7 and 28 days on the PFOA diet, TC, 

triglycerides, and glucose levels were significantly decreased. The AST response did not show a 

duration-related response because there was a significant decrease at 1 and 28 days, but not at 

7 days on the PFOA diet. Liver DNA concentration was significantly decreased (p<0.05) in all 

PFOA-exposed rats except those treated for 1 day in the second study, but liver DNA content 

was not altered by PFOA, suggesting that the increase in volume was responsible for the change 

in concentration. 

Perfluorooctanoic acid (PFOA) – May 2016 3-64
Table 3-14. Hepatic Effects of Rats Exposed to PFOA 

Day

Study 1 Study 2

Control

300 ppm 

PFOA 

(19 

mg/kg/day)

50 ppm

Wyeth 

14,643 Control

300 ppm 

PFOA 

(32 

mg/kg/day)

50 ppm

Wyeth 14,643

Liver 

weight (g)

1 13.6 ± 1.3 14.1 ± 2.4 15.7 ± 1.2 15.2 ± 1.9 14.4 ± 0.9 15.8 ± 1.4

7 15.3 ± 1.3 19.2 ± 3.1* 23.1 ± 3.1* 16.6 ± 1.7 22.8 ± 2.6* 23.4 ± 2.5*

28 18.3 ± 2.5 20.8 ± 3.2 30.6 ± 3.2* 17.2 ± 2.0 24.6 ± 2.2* 29.2 ± 4.0*

Liver-tobw (g/kg)1 4.25 ± .34 4.39 ± 0.44 4.64± 0.17* 4.39 ± 0.36 4.27 ± 0.14 4.49 ± 0.23

7 4.10 ± 0.26 5.83 ± 0.55* 6.26 ± 0.48* 4.28 ± 0.24 6.56 ± 0.38* 6.34 ± 0.33*

28 3.96 ± 0.36 5.83 ± 0.56* 7.09 ± 0.42* 3.70 ± 0.21 6.13 ± 0.53* 6.65 ± 0.59*

Labeling 

index (%)

1 0.22 ± 0.17 0.74 ± 0.55* 2.10 ± 1.10* 1.02 ± 0.37 2.18 ± 0.73* 4.54 ± 1.03*

7 1.42 ± 0.65 5.94 ± 2.12* 12.56 ± 6.42* 2.57 ± 1.31 13.18 ± 3.18* 23.85 ± 7.02*

28 ND 2.08 ± 1.03 10.15 ± 2.69 0.66 ± 0.45 1.74 ± 0.96* 5.34 ± 2.79*

Source: Elcombe et al. 2010

Notes: *Significantly different from control (p < 0.05); ND = No Data. 

After 1 day on the Wyeth 14,463 diet, subjects’AST and TC were significantly decreased. 

After 7 and 28 days on the Wyeth 14,643 diet, subjects’ ALT, TC, triglycerides, and glucose 

levels were significantly decreased. AST was not significantly decreased in rats fed Wyeth 

14,643 diets for 28 days, but it was after 7 days. In the Wyeth 14,643 rats, liver DNA 

concentration was significantly decreased after 1 and 7 days in the first study and 7 and 28 days 

in the second study. Total liver DNA content in the Wyeth 14,643-treated rats was significantly 

increased after 7 and 28 days in both studies. 

Labeling indices for hepatic cell proliferation, as measured by bromodeoxyuridine (5-bromo2-deoxyuridine, BrdU) incorporation, was significantly increased after day 1 and 7 in study 1 in 

both PFOA (p<0.05) and Wyeth 14,643 (p<0.01) diet-fed rats. Samples from control livers at 

day 29 were not available for comparison. In study 2, labeling was significantly increased 

(p<0.05) at all time points in both groups of rats compared to labeling in control rats (Table 

3-14). Apoptosis of hepatic cells was not altered by treatment with PFOA at any time point. In 

rats fed diets containing Wyeth 14,643 for 28 days, hepatic apoptosis was significantly decreased 

(p<0.01) compared to apoptosis observed in control livers. 

Histological examination of the livers of PFOA and Wyeth 14,643 diet-fed rats showed 

decreased glycogen after 1, 7, and 28 days. An increase in hepatocellular hypertrophy was 

observed after 7 and 28 days on the diets, fatty vacuolation was observed after 7 days on the 

diets, and increased hepatocellular hyperplasia was observed after 28 days on the diets. The 

hepatic observations were similar in both studies, and findings in Wyeth 14,643 diet-fed rats 

were generally more pronounced or severe than those in PFOA diet-fed rats. Although there 

were many similarities in response to the PFOA and Wyeth 14,463 diets, the body weight and 

apoptosis responses differed.

Mouse. Kennedy (1987) fed male and female Crl:CD-1 mice diets containing 0, 30, 300, and 

3,000 ppm PFOA for 14 days. At necropsy body weight and liver weight were recorded and 

analyzed. No histological evaluations were conducted. All mice died at 3,000 ppm. At 300 ppm, 

body weight was decreased and one female died. Both male and female mice had significantly 

increased absolute and relative liver weights at all doses (p<0.05) compared to the control. The 

Perfluorooctanoic acid (PFOA) – May 2016 3-65
LOAEL was 30 ppm based on increased liver weight, and no NOAEL was established. Kennedy 

(1987) used lower doses in a follow-up study lasting 21 days. Male and female mice were fed 

diets containing 0, 0.01, 0.03, 0.1, 0.3, 1, 3, 10, and 30 ppm PFOA. Absolute and relative liver 

weights for male and female mice were significantly increased (p<0.05) at ≥ 3 ppm PFOA. The 

LOAEL was 3 ppm based on increased liver weight, and the NOAEL was 1 ppm. 

Loveless et al. (2008) administered 0, 0.3, 1, 10, and 30 mg linear PFOA/kg by oral gavage 

to groups of male CD-1 mice (n = 20 per group) for 29 days. Body weight was recorded on days 

0, 3, and 6–28. At necropsy, blood was collected for hematology, clinical chemistry, and CORT

measurements. Tissues were collected for weight and microscopic examination. Body weight 

was significantly reduced at 10 and 30 mg/kg/day. An increase in neutrophils and monocytes 

was observed at ≥10 mg/kg/day along with a decrease in eosinophils. Serum CORT levels were 

significantly increased in mice dosed with 10 mg/kg and elevated in those dosed with 

30 mg/kg/day. Total serum cholesterol and triglycerides were significantly decreased at 

≥10 mg/kg/day. HDL was significantly reduced at ≥1 mg/kg/day. In mice treated with 30 mg/kg 

and water, triglycerides and HDL levels were significantly decreased compared to control levels.

Absolute and relative liver weights were significantly increased at ≥1 mg PFOA/kg/day.

Increased incidences of microscopic lesions in the liver included mild hepatocellular hypertrophy 

at 0.3 mg/kg/day, moderate-to-severe hypertrophy and individual cell necrosis at ≥1 mg/kg/day, 

and increased hepatocellular mitotic figures, fatty changes, and bile duct hyperplasia at 

≥10 mg/kg/day. The LOAEL for this study was 1 mg/kg/day based on increased liver weight,

hepatocellular hypertrophy, and cell necrosis; the NOAEL was 0.3 mg/kg/day. Data on several 

immunological endpoints were reported as part of the Loveless et al. (2008) publication. The 

immunological data are included in section 3.3.2 of this report. 

Some of the epidemiology studies report that evaluated serum lipids demonstrate a positive 

correlation between total serum cholesterol and triglycerides and serum PFOA. Tan et al. (2013) 

used male C57BL/6N mice to determine if dietary fat content could be an important variable 

influencing the impact of PFOA on serum lipids. Groups of seven or eight 4-month-old male 

mice were given either a liquid regular fat diet (RFD) or a high-fat diet (HFD), with or without 

PFOA, for 3 weeks. The RFD provided 12% and the HFD provided 35% of their calories from 

fat. Calories from protein (18%) were equivalent in both diets. The RFD provided 60% and the 

HFD provided 40% of their calories from carbohydrate. The fats were primarily 

monounsaturated (olive oil) or polyunsaturated (safflower and corn oil). PFOA was added to 

both diets for 3 weeks at a level that maintained a dose of 5 mg/kg/day to the mice. The PFOAtreated groups were fed ad libitum, and the control groups were given the amount consumed by 

the PFOA-treated groups the previous day. Body weight; liver weight; plasma ALT, AST, and 

ALP; total and direct bilirubin; free fatty acids and liver triglycerides; as well as subcutaneous 

and epididymal white adipose tissue were monitored. Statistical differences between groups 

(p<0.05) were determined using one-way ANOVA. Liver and epididymal white fat tissue

samples were examined histologically. 

The fat content of the diets alone resulted in significant differences in body weight and 

subcutaneous white adipose tissue, but not in liver weight. The addition of PFOA to the RFD 

resulted in significant increases in body weight, liver weight, ALT, ALP, and plasma free fatty 

acids, but not in AST or bilirubin. The addition of PFOA to both the RFD and HFD resulted in 

decreases in the mass of both epididymal and subcutaneous white fat deposits.

The HFD alone did not result in definitive alterations in liver histopathology. When PFOA 

was added to the RFD, indications of hepatocyte hypertrophy, necrosis, and inflammatory cell 

Perfluorooctanoic acid (PFOA) – May 2016 3-66
infiltration were observed. The liver damage in the animals being fed the HFD with PFOA was 

increased more than in the RFD-PFOA animals, as indicated by higher levels of necrosis and 

inflammation accompanied, in this case, by lipid droplet accumulation and significantly 

increased liver triglycerides, but not liver cholesterol or free fatty acids. In the epididymal 

adipose tissues, adipocyte size was increased in the HFD control compared to the RFD control 

but decreased with the addition of PFOA compared to both the RFD and HFD controls. 

Inflammatory cell infiltration was observed in the epididymal adipose tissues when PFOA was 

added to the HFD but not the RFD. No data for the subcutaneous white fat tissues was provided.

The authors evaluated the hepatic expression of 84 genes involved in the regulation of fatty 

acid metabolism using RT2

 Profiler PCR Arrays. HFD and/or PFOA altered the expression of 

33 genes (> 1.5 fold). PFOA alone upregulated 13 genes (>1.5) and downregulated 4 (>1.5) 

genes with fatty acid and triglyceride catabolism. Eight fatty acid transport-related genes were 

upregulated by PFOA and one was downregulated. The study demonstrates the importance of the 

fat content of the diet as a modulator of the effects of PFOA on the liver in animals. Damage to 

the liver tissues was intensified in the presence of the HFD.

Son et al. (2008) administered 0, 2, 10, 50, and 250 mg/L PFOA (0, 0.49, 2.64, 17.63, and 

47.21 mg/kg PFOA, respectively) in the drinking water to 4-week-old male imprinting control 

region (ICR) mice for 21 days. Food and water consumption, and body weight were recorded 

daily. At sacrifice, blood was collected and the liver and kidneys were removed and weighed. 

Plasma from the blood was used to determine levels of ALT, AST, BUN, and creatinine.

Sections of the liver and kidney were processed and stained with hematoxylin and eosin or 

stained for caspase 3 (a biomarker for apoptosis). Expression of mRNA for tumor necrosis 

factor-α, interleukin-1β, and transforming growth factor-β were determined using reverse 

transcription polymerase chain reaction (RT-PCR). 

The mice exposed to 250 mg/L PFOA (47.21 mg/kg/day) had significantly reduced food and 

water consumption (p<0.05), and body weight gain (p<0.05) compared to the control mice. Body 

weight gain also was significantly reduced (p<0.05) in mice receiving 50 mg/L PFOA in the 

drinking water. In all PFOA-exposed mice, relative liver weight was significantly increased in a 

dose-dependent manner (p<0.05) compared to liver weight of the control mice. Relative kidney 

weight was not affected by PFOA exposure. At ≥ 10 mg/L PFOA (2.64 mg/kg/day), plasma ALT 

activity was significantly increased, and at ≥ 50 mg/L PFOA (17.63 mg/kg/day), plasma AST 

activity was significantly elevated compared to the activity level in the control mice. Exposure to 

PFOA did not affect BUN or creatinine.

The livers of mice exposed to ≥50 mg/L PFOA were characterized by enlarged hepatocytes 

with acidophilic cytoplasm and the presence of eosinophils. No apoptotic bodies were observed 

in the liver with staining for caspase 3. Exposure to PFOA did not affect kidney morphology and 

did not cause toxic damage or necrosis in the kidney. In the liver, tumor necrosis factor-α 

expression was significantly reduced at ≥ 50 mg/L PFOA, interleukin-1β expression was 

significantly reduced at 250 mg/L PFOA, and transforming growth factor-β expression was 

significantly elevated at ≥ 50 mg/L PFOA. Under the conditions of this study, the LOAEL was 

2 mg/L (0.49 mg/kg/day) based on increased liver weight, and no NOAEL was established. The 

LOAEL for increased plasma ALT was 2.64 mg/kg/day.

Wolf et al. (2008a) gavage-dosed wild-type 129S1/SvlmJ mice (n = 7–8 per group) and 

PPARα-null mice (129S4/SvJae-PPARαtm1Gonz/J, n = 6–8 per group) with 0, 1, 3, or 10 mg 

PFOA/kg or 50 mg Wyeth 14,643, a PPARα agonist, and wild-type CD-1 (n = 7–8 per group) 

Perfluorooctanoic acid (PFOA) – May 2016 3-67
with 0, 1, and 10 mg PFOA/kg for 7 days to characterize hepatic effects resulting from exposure.

The mice were sacrificed 24 hours following the last dosing. Blood was collected for serum, and 

the livers were removed and weighed. Liver sections were stained with hematoxylin and eosin 

for examination by light microscopy and with uranyl acetate for transmission electron 

microscopy. Liver sections were also processed for immunohistochemistry of PCNA. Hepatocyte 

hypertrophy and vacuolation, observed in both strains of wild-type mice, were assigned a score 

from 0 to 4 based on severity, with 0 being no lesions observed and 4 being panlobular 

hypertrophy with cytoplasmic vacuolation. Hepatic lesions in PPARα-null were assigned a score 

(0–4) based on cytoplasmic vacuolation as no hypertrophy was observed. The percentage 

labeling index was obtained by counting the number of positive PCNA cells in 900–1,000 

hepatocyte nuclei per animal. Slides were read blind to treatment but with knowledge of genetic 

status.

Compared to control values, the absolute and relative liver weights, lesion score, and labeling 

index were significantly increased (p<0.05) in a dose-dependent manner in both strains of wildtype mice exposed to PFOA and also were significantly increased (p<0.05) in the wild-type 

129S1/SvlmJ mice exposed to Wyeth 14,643 (see Table 3-15). The absolute and relative liver 

weights and lesion score were significantly increased (p≤0.05) in a dose-dependent manner in all 

PFOA-exposed PPARα-null mice. The labeling index was significantly increased (p<0.05) in 

PPARα-null mice exposed to 10 mg PFOA/kg. Absolute and relative liver weights, lesion score, 

and labeling index of PPARα-null mice exposed to Wyeth 14,643 were no different from control 

values.

Table 3-15. Hepatic Effects in PFOA-Treated Mice

Group Liver Weight (g)

Relative Liver 

Weight (%) Lesion Score Labeling Index

Wild-type CD-1 Mice

Control 1.53 ± 0.14 4.5 ± 0.4 0.3 ± 0.5 0.6 ± 0.4

1 mg/kg/day PFOA 2.26 ± 0.24* 6.5 ± 0.5* 2.1 ± 0.9 0.7 ± 0.5

10 mg/kg/day PFOA 3.48 ± 0.54* 10.5 ± 0.8* 3.0 ± 0* 7.7 ± 3.0*

Wild-type 129S1/SvlmJ Mice

Control 0.87 ± 0.08 3.3 ± 0.4 0.3 ± 0.5 0.3 ± 0.2

1 mg/kg/day PFOA 1.22 ± 0.22* 1.6 ± 0.2* 2.0 ± 0.0* 0.7 ± 0.6

3 mg/kg/day PFOA 1.70 ± 0.12* 6.4 ± 0.4* 2.0 ± 0.0* 1.0 ± 0.4

10 mg/kg/day PFOA 2.20 ± 0.23* 8.3 ± 0.2* 4.0 ± 0.0* 2.4 ± 0.9*

50 mg/kg/day Wyeth

14,643

1.5 ± 0.13* 5.6 ± 0.1* 3.3 ± 0.5* 2.1 ± 1.2*

PPARα-null Mice

Control 0.92 ± 0.08 3.4 ± 0.4 1.1 ± 0.4 0.2 ± 0.2

1 mg/kg/day PFOA 1.2 ± 0.14* 4.5 ± 0.2* 1.9 ± 0.6* 0.6 ± 0.4

3 mg/kg/day PFOA 1.46 ± 0.21* 5. 8 ± 0.3* 3.0 ± 0.0* 0.6 ± 0.3

10 mg/kg/day PFOA 2.8 ± 0.18* 9.4 ± 0.6* 4.0 ± 0.0* 7.7 ± 3.0*

50 mg/kg/day

Wyeth 14,643

1.07 ± 0.24 3.9 ± 0.5 1.4 ± 0.5 0.6 ± 0.5

Source: Wolf et al. 2008a 

Note: * Statistically different from control, p < 0.05. 

Perfluorooctanoic acid (PFOA) – May 2016 3-68
Ultrastructure evaluations were done on liver sections from wild-type 129S1/SvlmJ mice and 

PPARα-null mice, but not from CD-1 mice. There were the expected differences in the 

characteristics of hepatocytes from the control wild-type mice when compared to both the 

PFOA-treated and Wyeth 14,643 wild-type mice. In the PPARα-null mice, the responses of the 

control and Wyeth 14,643-dosed animals were similar, but the response of the PFOA-dosed 

animals differed. Table 3-16 summarizes the cellular characteristics of the hepatocytes for the 

control, POFA-treated, and Wyeth 14,643-treated wild-type and PPARα-null mice on the basis 

of their glycogen content, Golgi bodies and associated rough ER, mitochondria, peroxisomes, 

and lipid-like cytoplasmic vacuoles.

Table 3-16. Mouse Hepatocyte Ultrastructure After PFOA or Wythe 14,643 Treatment

Mouse/Treatment

Characteristics

Glycogen

Golgi/ Rough 

ER Mitochondria Peroxisomes

Lipid-like 

Vacuoles

Wild-type/Control Prominent Prominent Numerous Few Rare

Wild-type/PFOA

(10 mg/kg)

Negative Nominal/ scarce 

ER

Numerous Numerous Scattered

Wild-type/Wyeth Negative Nominal/ scarce 

ER

Numerous Numerous Scattered

PPARα-null/Control Prominent Prominent Numerous Absent Scattered

PPARα-null/PFOA

(10 mg/kg)

Limited Limited Not reported Not reported Numerousa

PPARα-null/Wyeth Prominent Prominent Numerous Absent Scattered

Source: Wolf et al. 2008a 

Note: a Described as electron-dense, nonmembrane-bound spaces morphologically consistent with lipids ranging from the size of 

mitochondria to the size of nuclei. The vacuoles were believed to be an accumulation of PFOA.

It is apparent from the data in Table 3-16 that PFOA and Wyeth 14,643 behaved similarly in 

the wild-type strains but differently in the PPARα-null mice. The hepatocytes of PFOA-dosed 

PPARα-null mice exhibited lower glycogen content, Golgi bodies, and associated rough ER than 

both the control and Wyeth 14,643 PPARα-null mice. In addition, the PFOA-dosed PPARα-null 

mice had numerous large nonmembrane-bound lipid-like vacuoles throughout the cytoplasm. At 

the high dose (10 mg/kg/day), there was an increase in the labeling index that was not observed 

with Wyeth 14,643. The authors concluded that the large lipid-like vacuoles in the hepatocytes of 

PFOA-dosed PPARα-null mice were likely accumulations of PFOA. Under the conditions of this 

study, the LOAEL was 1 mg/kg/day based on increased absolute and relative liver weight and 

hepatic morphology changes; no NOAEL was established.

Nakamura et al. (2009) investigated the functional difference in PFOA response between 

mice and humans using a humanized PPARα transgenic mouse strain (hPPARα). Humanized 

PPARα mice express a high level of human PPARα protein in the liver. Male 8-week-old wildtype (mPPARα) mice, PPARα-null mice, and hPPARα mice were gavage-dosed with 0, 0.1, and 

0.3 mg/kg/day PFOA (n = 4–6 per group) for 2 weeks and sacrificed 18–20 hours following the 

last dose. Blood was collected and analyzed for triglyceride and cholesterol concentrations, and 

ALT measurements. Livers were collected and analyzed for triglyceride and cholesterol 

concentrations, plus histopathological changes. The differences in the observations for the three 

strains of mice are summarized in Table 3-17. 

Perfluorooctanoic acid (PFOA) – May 2016 3-69
Table 3-17. Relative Response of hPPARα, mPPARα, and PPARα-null Mice to PFOA

Parameter hPPARα mPPARα PPARα-null

Liver weight ND ↑ compared to control (0.3 

mg/kg/day)

↓ compared to control (0.1 

mg/kg/day)

Liver/body weight ratio ND ↑ compared to control (0.3 

mg/kg/day)

ND

Hepatocyte hypertrophy Mild (0.3 mg/kg/day) Mild (0.3 mg/kg/day) ND

ALT ND ND ND

Plasma cholesterol ↑ compared to mPPARα 

(all doses)

ND ND

Liver cholesterol ↓ compared to PPARα-null 

(0.1, 0.3 mg/kg/day), 

mPPARα (0.3 mg/kg/day)

↑ compared to control (0.3 

mg/kg/day)

ND

Plasma triglyceride ND ND ND

Liver triglyceride ↓ compared to PPARα-null 

(0.3 mg/kg/day)

↓ compared to PPARα-null 

(0.1, 0.3 mg/kg/day; ↑ 

compared to control (0.3 

mg/kg/day)

↑ compared to mPPARα 

(all doses)

Source: Nakamura et al. 2009

Notes:

hPPARα: transgenic mice (that express a high level of human PPARα protein in the liver); mPPARα: wild-type mice. 

↑ = significant increase (p < 0.05). 

↓ = significant decrease (p < 0.05) 

ND = no differences. 

Body weight of the hPPARα mice was slightly lower than the mPPARα and PPARα-null 

mice prior to PFOA treatment and remained lower throughout the dosing regimen. Treatment 

with PFOA did not affect plasma ALT or triglyceride concentrations in any group. The hPPARα 

mice differed from the wild-type mice in that their plasma cholesterol was significantly increased 

and their liver cholesterol and triglycerides significantly decreased at the highest dose (Table 

3-17). In addition, the increases in absolute and relative liver weights were less than those 

observed in the wild-type mice. The PPARα-null mice differed from the wild-type in that liver 

triglycerides were significantly increased. Comparable to the Wolf et al. (2008a) report, the 

cytoplasmic vacuoles were larger in the PPARα-null mice than in the wild-type and hPPARα 

mice. There were no other significant differences between PPARα-null mice and wild-type mice.

Under the conditions of the study, the LOAEL for mPPARα mice was 0.3 mg/kg/day of 

PFOA based on increased liver weight and increased liver triglyceride and cholesterol 

concentrations. The NOAEL for mPPARα mice was 0.1 mg/kg/day of PFOA. The NOAEL for 

PPARα-null mice was 0.3 mg/kg/day because the changes in absolute liver weight were not 

dose-related and the increase in relative liver weight was not significantly different from the 

control. The NOAEL for hPPARα mice was 0.3 mg/kg/day of PFOA, the highest dose tested. 

However, a nonsignificant but dose-related increase was observed in plasma cholesterol.

Minata et al. (2010) examined hepatobiliary injury in mice treated with PFOA. Male wildtype 129S4/SvlmJ mice (n = 39) and PPARα-null (129S4/SvJae-Pparαtm1Gonz/J mice (n = 40) 

were orally dosed with 0, 12.5, 25, and 50 μmol/kg/day of PFOA (equivalent to ~0, 5.4, 10.8, 

and 21.6 mg/kg/day of PFOA) for 4 weeks. At the end of 4 weeks, animals were sacrificed and 

blood, liver, and bile were collected for clinical chemistry analysis and determination of PFOA 

concentration. Sections of the liver were processed for histological examination, oxidative DNA 

damage, and multidrug resistance protein 2 (Mdr2) and tumor necrosis factor α (TNF-α) mRNA 

Perfluorooctanoic acid (PFOA) – May 2016 3-70
expression. Bile acid and phospholipid contents in bile were determined as well as the protein 

expression of canalicular bile salt export pump (BSEP) and canalicular MRP2. 

Absolute and relative liver weights in all PFOA treated wild-type and PPARα-null mice were 

significantly increased (p<0.05) at sacrifice compared to control liver weight. Plasma AST was 

significantly increased in wild-type mice at 25 and 50 μmol/kg/day (equivalent to 10.8 and 

21.6 mg/kg/day) and in PPARα-null mice at 50 μmol/kg/day compared to the concentrations of 

their respective controls. Plasma ALT was no different from control in the treated mice. In wildtype mice, total bilirubin was significantly decreased at 12.5 μmol/kg/day and significantly 

increased at 50 μmol/kg/day. In PPARα-null mice, total bilirubin was significantly increased at 

50 μmol/kg/day. Total bile acid was significantly increased at 50 μmol/kg/day in PPARα-null 

mice. TC was significantly decreased in wild-type mice at 25 and 50 μmol/kg/day, and total 

triglyceride was significantly increased at 12.5 and 25 μmol/kg/day. TC was significantly 

decreased at 12.5 and 25 μmol/kg/day and significantly increased at 50 μmol/kg/day in PPARαnull mice. In PPARα-null mice, total triglycerides were significantly increased at all doses.

Hepatocellular hypertrophy was observed in wild-type mice treated with 12.5, 25, and 

50 μmol/kg/day (equivalent to 5.4, 10.8 and 21.6 mg/kg/day). A dose-dependent increase in 

eosinophilic cytoplasmic changes consistent with peroxisome proliferation was observed in liver 

parenchyma, but no fat droplets or focal necrosis were observed in wild-type mice. An increase 

in bile duct epithelium thickness suggested slight cholangiopathy in wild-type mice at 25 and 

50 μmol/kg/day. Increased apoptosis in hepatic cells, hepatic arterial walls, and bile duct 

epithelium was observed at 25 and 50 μmol/kg/day in wild-type mice. Ultrastructure examination 

of livers from PFOA-treated wild-type mice showed decreased glycogen granules, degranulated 

or disrupted rough ER, nuclear vacuoles, extensive peroxisome proliferation, and slight 

mitochondria proliferation. 

In PPARα-null mice treated with 12.5, 25, and 50 μmol/kg/day of PFOA (equivalent to 5.4, 

10.8 and 21.6 mg/kg/day), hepatocellular hypertrophy, cytoplasmic vacuolation, and increased 

microvesicular steatosis were observed. These observations are consistent with Wolf et al. 

(2008a). At 50 μmol/kg/day, focal necrosis was observed. Areas of bile fibrosis and bile plaque 

and few inflammatory cells were observed in the bile ducts of PPARα-null mice at 25 and 

50 μmol/kg/day. Increased apoptosis was observed in bile duct epithelium at 25 and 

50 μmol/kg/day in PPARα-null mice. Ultrastructure examination of livers from PFOA-treated 

PPARα-null mice showed decreased glycogen granules, degranulated or disrupted rough ER, 

increased cytoplasmic lipid accumulation, mitochondria proliferation, and mitochondrial changes 

(e.g., swelling and decreased matrix density). Peroxisome proliferation was not observed.

Ultrastructure of bile duct showed degradation of cytoplasmic structure, vacuolization, 

disintegration of nuclei and organelles, periductal infiltration of fibroblasts and macrophages, 

and fibrosis. 

The marker for oxidative damage, 8-hydroxydeoxyguanosine (8-OH-dG), and TNF-α were 

not elevated or upregulated in wild-type mice. In PPARα-null mice, 8-OH-dG was elevated in 

the liver at 21.6 mg/kg/day and TNF-α mRNA was significantly increased at 10.8 and 21.6 

mg/kg/day. The transporter Mdr2 moves biliary phospholipids from hepatocytes to bile and was 

significantly upregulated in wild-type mice at all doses, but only at 5.4 mg/kg/day in PPARα-null 

mice. The BSEP transports bile acid from hepatocytes to bile and was significantly decreased in 

wild-type mice at 21.6 mg/kg/day, significantly increased in PPARα-null mice at 5.4 mg/kg/day, 

and significantly decreased at 21.6 mg/kg/day. The transporter MRP2 also transports bile acid 

and was significantly decreased at 21.6 mg/kg/day in both groups of mice. 

Perfluorooctanoic acid (PFOA) – May 2016 3-71
Under the conditions of the study, the LOAEL for male wild-type and PPARα-null mice was 

5.4 mg/kg/day of PFOA based on increased liver weight. A NOAEL was not established. At the 

LOAEL, the difference between the PPARα-null mice and the wild-type mice was the presence 

of cytoplasmic vacuoles and microvesicular steatosis in addition to hypertrophy in the PPARαnull mice.

The effects of gavage exposure on groups of six male Klunming mice (8 weeks old) to doses 

of 0, 2.5, 5, and 10 mg PFOA/kg/day for 14 days on the testes and epididymis was examined by 

Liu et al. (2015). The lowest dose tested was a LOAEL for dose-related effects on decreased 

sperm count, testicular superoxide dismutase (SOD), catalase, nuclear respiratory factor 2 

(NRF2), and BAX expression (0<0.05) plus increases in MDA, hydrogen peroxide, BAX and 

BCL expression (p<0.05). There was no effect on relative testes weight at any dose. Some effects 

were observed on testicular morphology at the lowest dose, including atrophy of the 

seminiferous tubules, depletion of spermatogonial cells, detachment of germ cells from the 

seminiferous epithelium, and decreased sperm production. The severity of the testicular 

morphological changes increased with dose. Six animals per dose group were used for the 

evaluation of testicular weight and 4 animals per dose group were used for the other assays. The 

increase in MDA and hydrogen peroxide accompanied by the decrease in SOD and carnitine 

acyltransferase (CAT) activity and NRF2 expression indicate that oxidative stress played a major 

role in the observed toxicity. NRF2 plays an important role as a messenger that upregulates 

genes involved in response to oxidative stress.

Lu et al. (2015) reported on the testicular effects of PFOA on the blood testes barrier after a 

28-day exposure of BALBL/c male mice (14 days old) to gavage doses of 0. 1.25, 5, and

20 mg/kg/day (3–5 animals per dose group). The blood testes barrier divides the seminiferous 

epithelium into apical and basolateral compartments and plays an important role in germinal cell 

development and male fertility. The barrier prevents the passage of large molecules from one 

compartment to the other. At termination of the exposure, the animals were sacrificed and the 

testes recovered for analysis. A second component of this study examined the impact of the 

PFOA treatment on male fertility and is reported in section 3.2.6. 

The blood testes barrier integrity was weakened at the lowest dose tested and in a dosedependent manner as indicated by the passage of a red fluorescent dye injected into the 

interstitium and concentrations of IgG measure in gel electrophoresis columns visualized by 

chemiluminescence (three per dose group). Membrane integrity is dependent on coexisting tight 

junctions, basal ectoplasmic specializations, and gap junctions. Accompanying in vitro assays of 

cultured sertoli cells demonstrated downregulation of key proteins associated with the tight 

junction and gap junction intercellular communication (GJIC) and regulation of the ectoplasmic 

specialization protein N-cadhedran. Tumor necrosis factor actin protein in the testes increased in 

a dose-related fashion at 5 and 20 mg/kg/day on observation of three per dose group. The authors 

identified the 5-mg/kg/day dose as a LOAEL for PFOA effects on the blood testes barrier and the 

1.25-mg/kg/day dose as a NOAEL, apparently based on the results for the key protein 

biomarkers for cellular intercommunication rather than the IgE and fluorescence results where 

1.25 mg/kg/day was a LOAEL. 

Li et al. (2011) investigated the involvement of mouse and human PPARα in PFOA-induced 

testicular toxicity. Wild-type, PPARα-null, and humanized PPARα male 129/Sv mice were given 

PFOA daily by gavage at doses of 0, 1, and 5 mg/kg/day for 6 weeks. Body weight and testis 

weight were not affected by treatment in any group. Absolute and relative-to-body weights of the 

epididymis and seminal vesicle plus prostate gland were decreased only in high-dose wild-type 

Perfluorooctanoic acid (PFOA) – May 2016 3-72
mice compared to the wild-type controls. No effects on sperm count and motility were seen in 

any group. Sperm abnormalities were significantly increased in both treated groups of wild-type 

and humanized PPARα mice, but not in the PPARα-null mice. Plasma testosterone levels were 

slightly decreased in low-dose wild-type mice, and significantly decreased in high-dose wildtype and low- and high-dose humanized PPARα mice compared to the control groups.

Testosterone levels were slightly reduced in a dose-related manner in the PPARα-null mice, but 

statistical significance was not attained. 

Using real-time quantitative PCR, the mRNA levels for several genes associated with 

testicular cholesterol synthesis, transport, and testosterone biosynthesis were examined. Levels 

of 3-hydroxy-3-methylglutaryl coenzyme A (HMG-CoA) synthase, HMG-CoA reductase, and 

aromatase were not changed after treatment in any group. Expression of steriodogenic acute 

regulatory protein (which transports cholesterol into mitochondria) was inhibited in wild-type 

mice at the high dose and in humanized PPARα mice at both doses; peripheral benzodiazepine 

receptor level was decreased only in high-dose humanized PPARα mice; cytochrome P450 sidechain cleavage enzyme was decreased in both groups of wild-type mice; cytochrome P450 17αhydroxylase/C17-20 lyase was inhibited at the high dose in both wild-type and humanized 

PPARα mice; and 3β-hydroxysteroid dehydrogenase was decreased in both treated groups of 

humanized PPARα mice. Decreased expression of 17β-hydroxysteroid dehydrogenase was the 

only change found in treated PPARα-null mice.

In the mitochondria, carnitine palmitoyltransferase (CPT) was decreased in both groups of 

wild-type and high-dose humanized PPARα mice, and SOD levels were reduced in all treated 

wild-type and humanized PPARα mice. Histopathological lesions of the testes, including 

abnormal seminiferous tubules, lack of germ cells, or necrotic cells, were observed in high-dose 

wild-type and humanized PPARα mice. No morphological changes were observed in the testes 

from PFOA treatment in PPARα-null mice. The 1-mg/kg/day dose was the LOAEL for 

significant (p<0.05) sperm abnormalities, decreased testosterone, and several biochemical 

alterations in the PPARα and hPPARα mice, but not in the PPARα-null mice. There were doserelated decreases in testosterone in the PPARα-null mice, but they did not achieve statistical 

significance.

Inhalation Exposure 

No data on the effects of short-term inhalation exposures to PFOA were identified in the 

literature.

Dermal Exposure

Fairley et al. (2007) investigated the role of dermal exposure to PFOA in an experiment to 

evaluate toxicity in BALB/c mice. The mice were exposed to 0, 0.01%, 0.1%, 0.25%, 0.5%, 

1.0%, and 1.5% PFOA (equivalent to 0, 0.25, 2.5, 6.25, 12.5, 25, and 50 mg/kg PFOA). It was 

applied to the dorsal surface of both ears daily for 4 days. The mice were sacrificed 6 days later.

Dermal PFOA exposure did not cause reductions in body weight or signs of inflammation at the 

application site. A significant increase in liver weight was observed in mice dosed with 

≥ 6.25 mg/kg PFOA (p<0.01) compared to control liver weight. Under the conditions of the 

study, the LOAEL was 6.25 mg/kg PFOA based on increased liver weight, and the NOAEL was 

2.25 mg/kg PFOA. 

Perfluorooctanoic acid (PFOA) – May 2016 3-73
3.2.3 Subchronic Studies

Oral Exposure 

Monkey. Goldenthal (1978) administered rhesus monkeys (two per gender per group) doses of 

0, 3, 10, 30, and 100 mg/kg/day PFOA by gavage for 90 days. Animals were observed twice 

daily and body weights were recorded weekly. Blood and urine samples were collected once 

during a control period, and at 1 and 3 months for hematology, clinical chemistry, and urinalysis. 

Organs and tissues from animals that were sacrificed at the end of the study and from animals 

that died during the treatment period were weighed, examined for gross pathology, and 

processed for histopathology. 

All monkeys in the 100-mg/kg/day group died between weeks 2–5 of the study. Signs and 

symptoms that first appeared during week 1 included anorexia, frothy emesis, swollen face and 

eyes, decreased activity, prostration, and body trembling. Three monkeys from the 30-mg/kg/day 

group died during the study. Beginning in week 4, all four animals showed slight to moderate, 

and sometimes severe, decreased activity. One monkey had emesis and ataxia, swollen face, 

eyes, and vulva. Beginning in week 6, two monkeys had black stools and one monkey had slightto-moderate dehydration. No monkeys in the 3- or 10-mg/kg/day groups died during the study. 

One monkey in the 10-mg/kg/day group was anorexic during week 4, had a pale and swollen 

face in week 7, and had black stools for several days in week 12. Animals in the 3-mg/kg/day 

group occasionally had soft stools or moderate-to-marked diarrhea and frothy emesis.

Changes in body weight were similar to the controls for animals from the 3- and 10-

mg/kg/day groups. Monkeys from the 30- and 100-mg/kg/day groups lost body weight after 

week 1. At the end of the study, this loss was statistically significant for the one surviving male 

in the 30-mg/kg/day group and reflected in body weight (2.30 kg versus 3.78 kg for the control). 

The results of the urinalysis, and hematological and clinical chemistry analyses were comparable 

for the control and the 3- and 10-mg/kg/day groups at 1 and 3 months. 

At necropsy, there were significant decreases in the absolute heart and brain weight and 

relative liver weight in 10-mg/kg/day females. At 3 mg/kg/day, the relative pituitary weight in 

males was significantly increased. The biological significance of these weight changes is 

difficult to assess, as they were not accompanied by morphologic changes. 

In animals that died, one male and two females from the 30-mg/kg/day group and all animals 

from the 100-mg/kg/day group had marked diffuse lipid depletion in the adrenal glands. All 

males and females from the 30- and 100-mg/kg/day groups also had slight to moderate 

hypocellularity of the bone marrow and moderate atrophy of lymphoid follicles in the spleen. 

One female from the 30-mg/kg/day group and all animals in the 100-mg/kg/day group had 

moderate atrophy of the lymphoid follicles in the lymph nodes. 

The one male in the 30-mg/kg/day group that survived until terminal sacrifice had slight-tomoderate hypocellularity of the bone marrow and moderate atrophy of lymphoid follicles in the 

spleen. Under the conditions of this study, the male LOAEL was 3 mg/kg/day based on increased 

relative pituitary weight, and no NOAEL was established. The female LOAEL was 

10 mg/kg/day based on decreased heart and brain weight, and the NOAEL was 3 mg/kg/day. 

Perfluorooctanoic acid (PFOA) – May 2016 3-74
Rat. In a dietary study reported by Perkins et al. (2004), male ChR-CD rats (45–55 per group) 

were administered concentrations of 1, 10, 30, and 100 ppm PFOA for 13 weeks. These doses 

are equivalent to 0.06, 0.64, 1.94, and 6.50 mg/kg/day. There were two control groups—a 

nonpair-fed control group and a pair-fed control group for the 100-ppm dose group); both were 

fed the basal diet. Following the 13-week exposure period, 10 animals per group were fed basal 

diet for an 8-week recovery period. The animals were observed twice daily for clinical signs of 

toxicity, and body weights and food consumption were recorded weekly. Food consumption was 

recorded daily for the pair-fed animals.

A total of 15 animals per group were sacrificed following 4, 7, and 13 weeks of treatment; 

10 animals per group were sacrificed after 13 weeks of treatment and an 8-week recovery period. 

Serum samples collected from 10 animals per group at each scheduled sacrifice during treatment 

and from five animals per group during recovery were analyzed for estradiol, T, LH, and PFOA. 

The level of palmitoyl-CoA oxidase was analyzed from a section of liver that was obtained from 

five animals per group at each scheduled sacrifice. Weights of the brain, liver, lungs, testis, 

seminal vesicle, prostate, coagulating gland, and urethra were recorded, and these tissues also 

were examined histologically. In addition, the brain, liver, lungs, testis, seminal vesicle, and 

prostate were preserved in glutaraldehyde for electron microscopic examination. 

In the analysis of the data, animals exposed to 1, 10, 30, and 100 ppm PFOA were compared 

to the control animals in the nonpair-fed group, while the data from the pair-fed control animals 

were compared to animals exposed to 100 ppm PFOA. At 100 ppm, significant reductions in 

body weight and body weight gain were seen compared to the pair-fed control group during 

week 1 and the nonpair-fed control group during weeks 1–13. Body weight data in the other 

dosed groups were comparable to controls. At 10 and 30 ppm, mean body weight gains were 

significantly lower than for the nonpair-fed control group at week 2. These differences in body 

weight and body weight gains were not observed during the recovery period. Animals fed 

100 ppm consumed significantly less food during weeks 1 and 2 than the nonpair-fed control 

group. Overall, there was no significant difference in food consumption. There were no 

significant differences among the groups for any of the hormones evaluated, although there was

some indication of elevated estradiol for the 100 ppm group at week 4. The elevated estradiol for 

the high-dose group at week 4 should be interpreted with caution because most of the 

measurements for control and treated groups were below the level of detection at all other 

timepoints (Perkins et al. 2004). 

Significant dose-related increases in absolute and relative liver weights and hepatocellular 

hypertrophy were observed at weeks 4, 7, and 13 in the 10, 30, and 100 ppm groups 

(Table 3-18). There was no significant evidence of any dose-related degenerative changes. 

Hepatic palmitoyl-CoA oxidase activity was significantly increased at weeks 4, 7, and 13 in the 

30 and 100 ppm groups. At 10 ppm, hepatic palmitoyl-CoA oxidase activity was significantly 

increased at week 4 only. At 13 weeks, the palmitoyl-CoA oxidase activity was lower than it was 

at weeks 4 and 7 for the 10, 30, and 100 ppm dose groups, possibly suggesting attenuation of the 

peroxisomal response. Histologically, liver effects were limited to minimal or slight coagulative 

necrosis observed in 0/45, 1/45, 0/45, 2/45, and 3/44 in the control, 1, 10, 30, and 100 ppm 

groups, respectively. 

Perfluorooctanoic acid (PFOA) – May 2016 3-75
Table 3-18. Liver Effects in Male Rats

Parameter Week

Dose (mg/kg/day)

0 a 0 b 0.06 (1 ppm) 0.64 (10 ppm) 1.94 (30 ppm) 6.5 (100 ppm)

Palmitoyl-CoA

Oxidase (IU/g)

4

7 

13

8 ± 0.5

7 ± 1.5

8 ± 0.9

8 ± 0.4

7 ± 1.5

5 ± 1.1

9 ± 1.7

7 ± 0.8

8 ± 1.9

14 ± 3.6c

16 ± 5.5

10 ± 2.1

24 ± 11.4c

32 ± 12.2c 

14 ± 3.4c 

37 ± 14.8c, d

54 ± 35.3c, d 

17 ± 4.5cd

Hepatocellular

Hypertrophy

4

7 

13

0/15

0/15

0/15

0/15

0/15

0/15

0/15

0/15

0/15

12/15

12/15

13/15

15/15

15/15

14/15

14/15

15/15

15/15

Hepatocellular

Necrosis,

Coagulative

4

7 

13

0/15

0/15

0/15

1/15

0/15

0/15

0/15

0/15

1/15

0/15

0/15

0/15

1/15

0/15

1/15

2/14

1/15

0/15

Absolute Liver

Weight (g)

4

7 

13

16.34 ± 2.14

17.78 ± 2.12

19.73 ± 2.01

15.83 ± 1.13

16.91 ± 2.22

16.30 ± 1.62

15.45 ± 1.71

17.68 ± NA

18.03 ± 2.81

17.89 ± 2.13

19.42 ± 2.10

20.44 ± 2.87

23.23 ± 2.83c

27.76 ± 3.51c 

22.74 ± 4.21

25.44 ± 1.89c

27.76 ± 3.51c 

26.78 ± 5.47c 

Mean Body 

Weight (g) 

4

7 

13

388 ± 21

457 ± 29

541 ± 41

365 ± 11

434 ± 19

508 ± 22

388 ± 23

461 ± 30

548 ± 37

383 ± 25

458 ± 30

551 ± 42

380 ± 27

448 ± 31

531 ± 46

356 ± 27c

432 ± 39c

494 ± 64c

Liver/Body

Weight (%)

4

7 

13

3.97 ± 0.37

3.75 ± 0.29

3.53 ± 0.28

4.07 ± 0.27

3.76 ± 0.37

3.24 ± 0.23

3.73 ± 0.23

3.64 ± 0.33

3.24 ± 0.30c 

4.49 ± 0.32c

4.12 ± 0.37

3.69 ± 0.32

5.77 ± 0.60c

5.14 ± 0.53c 

4.21 ± 0.56c 

6.73 ± 0.49c

6.06 ± 0.72c 

5.50 ± 0.84c 

Source: Perkins et al. 2004

Notes: Mean ± SD; NA= not available.

a

 Nonpair-fed controls. b Pair-fed controls. c Statistically significant (p < 0.05) with nonpair-fed control. d Statistically significant (p < 0.05) with pair-fed control.

Under the conditions of this study, the authors identified the LOAEL as 10 ppm 

(0.64 mg/kg/day) based on increases in absolute and relative liver weight and hepatocellular 

hypertrophy (Perkins et al. 2004). The NOAEL identified was 1.0 ppm (0.06 mg/kg/day). 

However, the liver weight and palmitoyl-CoA responses were associated with the activation of 

PPARα and were not accompanied by significant dose-related changes that would classify them 

as adverse for humans (e.g., fibrosis, macrovesicular steatosis, inflammation) as enumerated by 

Hall et al. (2012). Therefore, for the current assessment, the LOAEL is identified as 

1.94 mg/kg/day based on a slight increased incidence of coagulative necrosis in the liver. The

NOAEL is 0.64 mg/kg/day. 

Serum samples were collected from 8 to 10 animals prior to each sacrifice. PFOA 

concentrations in serum increased with the dose, but all dose levels appeared to have reached

steady state by the first sacrifice at week 4. Following the 8-week recovery period, serum levels 

were below detection for many animals and consistent with a half-life of about seven days in 

male ChR-CD rats.

Inhalation and Dermal Exposure 

No data on the effects of subchronic inhalation or dermal exposures to PFOA were identified 

in the literature.

Perfluorooctanoic acid (PFOA) – May 2016 3-76
3.2.4 Neurotoxicity 

Johansson et al. (2008) gave male neonatal Naval Medical Research Institute (NMRI) mice 

(3–4 litters, ~5–6 male pups per litter) a single gavage dose of 0, 0.58, and 8.7 mg PFOA/kg in a 

lecithin/peanut oil emulsion on PND 10, the approximate peak time of rapid brain growth in 

mice. Spontaneous behavior (e.g., locomotion, rearing, and total activity) and habituation in 

response to a placement in unfamiliar environment were tested in 10 mice in each group at ages 

2 and 4 months. Each test period was divided into three 20-min periods. The habituation ratio 

was determined by dividing the activity for the third 20-min period by the activity for the first 

period. A high habituation ratio indicated that movement patterns of the exposed animals when 

placed in an unfamiliar test chamber differed from control by displaying comparatively low 

activity for the first 20 mins and comparatively higher activity for the last 20 mins.

Exposure to PFOA did not affect body weight or body weight gain in male NMRI mice 

following treatment. Compared to controls, the habituation ratio for rearing and locomotion in 

the high-dose animals was elevated compared to controls at 2 and 4 months, with a significantly 

higher ratio (p<0.01) at 4 months than at 2 months. At 4 months, the changes in activity patterns 

for the high dose were significant (p<0.01) compared to controls for locomotion, rearing, and 

total activity. The results at the low dose were less pronounced, with a significant impact on 

locomotion and slight changes in rearing behavior. 

At 4 months of age, mice were tested for nicotine-induced behavior and behavior in the 

elevated plus maze. Increased activity is the expected response to nicotine injection (80 μg) as a 

result of stimulation of the cholinergic receptors in the brain. The activity responses of the 

PFOA-exposed animals to nicotine stimulation were significantly less than the response of the 

controls, but the differences were most pronounced in the high-dose animals.

The mice also were tested in an elevated plus maze, which determined whether they would 

select an enclosed environment (the expected response) over an open environment. No 

significant differences were observed in the PFOA-exposed mice in this test. Under the 

conditions of this study, the clear LOAEL was 8.7 mg/kg based on locomotion, rearing, and total 

activity; habituation ratio; and response to nicotine at 2 and 4 months after receiving a single 

gavage dose on PND 10. There were significant differences in locomotion and total activity at 4 

months in the low-dose animals, which supports identifying the 0.58-mg/kg dose as a marginal 

LOAEL. However, the data at the low dose are less compelling than those at the high dose. 

In a follow-up to their original study, Johansson et al. (2009) gave male neonatal NMRI mice 

(3–4 litters, ~5–6 male pups per litter) a single gavage dose of 0 and 8.7 mg PFOA/kg on PND

10. Protein levels of calcium/calmodulin-dependent protein kinase II (CaMKII), growthassociated protein-43 (GAP-43), synaptophysin, and tau protein were determined in the cerebral 

cortex and hippocampus. CaMKII regulates synapotogenesis and synaptic plasticity, GAP-43 

modulates axon sprouting and growth, synaptophysin is a membrane glycoprotein in presynaptic 

vesicles, and tau protein is responsible for outgrowth of neuronal processes and microtubule 

assembly and maintenance.

Levels of CaMKII protein in the hippocampus were significantly higher (58%, p≤0.05) in 

mice exposed to PFOA than levels in control mice, but unchanged in the cerebral cortex. Levels 

of GAP-43 protein in the hippocampus were significantly higher (17%, p≤0.05) in PFOAexposed mice than levels in control mice, but unchanged in the cerebral cortex. Synaptophysin 

levels in mice exposed to PFOA were significantly increased in the hippocampus (52%) and 

Perfluorooctanoic acid (PFOA) – May 2016 3-77
cerebral cortex (82%). Tau protein levels in PFOA-exposed mice were increased 92% and 142% 

(p≤0.001) in the hippocampus and cerebral cortex, respectively, above levels in the control mice.

The authors concluded that alterations of these proteins could be a factor in the altered behavior 

of adult mice that were exposed to PFOA as neonates because they are required for normal brain 

development. 

Onishchenko et al. (2011) exposed pregnant C57BL/6/Bkl mice (n = 6 per group) to 0 and 

0.3 mg PFOA/kg/day in the diet from GD 1 to the end of pregnancy. The behavior of the weaned 

offspring was analyzed in locomotor, circadian activity, elevated plus maze, and forced swim 

tests at 5–8 weeks of age. Muscle strength and motor coordination tests were given at 3–4 

months of age. The distance traveled over 30 mins was registered in 5-min intervals in the 

locomotor test. For the circadian activity test, the activity of the mice in social groups was 

monitored for 48 hours after placement in new cages. Anxiety-like behavior was determined 

using the elevated plus maze. Depression-like behavior was determined in the forced swim test 

by tracking the time spent floating passively for 2 seconds or longer. Muscle strength (three 

trials) was measured by how long within 60 seconds it took the mouse to fall off an upside-down 

lid onto the cage floor. Motor coordination (four trials) was measured by how long the mice 

remained on a rotating drum as a rotarod accelerated from 4 to 40 rpm over 5 mins.

Prenatal exposure to PFOA did not alter offspring locomotor activity, anxiety-related 

behavior, depression-like behavior, or muscle strength. In the circadian activity tests, male 

offspring exposed to PFOA were significantly more active (p = 0.013) and the female offspring 

were significantly less active (p = 0.036) than control offspring during the first hour of the test. 

PFOA-exposed male offspring were significantly more active (p<0.05) than control males from 

the dark phase of day 1 through the dark phase day 2. Both male and female offspring exposed to 

PFOA had significantly less inactive periods (p<0.05) during the light phase compared to their 

respective controls. In the accelerating rotarod test, female offspring exposed to PFOA exhibited 

decreased fall latency over the four trials compared to control females, but no effect of treatment 

was observed in male offspring. The authors concluded that prenatal exposure to 0.3 mg/kg/day 

of PFOA resulted in gender-related postnatal alterations in offspring behavior and motor function 

at 3–4 months of age. 

In vitro. Slotkin et al. (2008) characterized the neurotoxicity of PFOA using PC12 cells. The 

cells were derived from a neuroendocrine tumor of the rat adrenal medulla and serve as a model 

for neuronal development and differentiation. Exposure to nerve growth factors causes PC12 

cells to differentiate into cells expressing either dopamine or acetylcholine phenotypes. The cells 

were incubated with 10, 50, 100, and 250 μmol PFOA. Synthesis of DNA, cell viability, cell 

growth, and lipid peroxidation were measured to determine if PFOA targets specific events in 

neural cell differentiation. Differentiation shifts towards or away from the dopamine and 

acetylcholine phenotypes were measured by assessing the activities of tyrosine hydroxylase (TH, 

dopamine) and choline acetyltransferase (ChAT, acetylcholine). The undifferentiated cells were 

evaluated after a 24-hour exposure to PFOA, and differentiating cells were evaluated after 4–6 

days of exposure to PFOA. 

Significant inhibition of DNA synthesis (p<0.0001) occurred in the undifferentiated cells 

after exposure to 250 μmol PFOA with no change in DNA content. Lipid peroxidation was 

significantly increased (p<0.02) after exposure to 10 μmol PFOA, and cell viability was 

significantly decreased (p<0.03) after a 24-hour exposure to 100 and 250 μmol PFOA.

Perfluorooctanoic acid (PFOA) – May 2016 3-78
In differentiating PC12 cells, exposure to 250 μmol PFOA caused decreased DNA content 

with no change in total protein/DNA content ratio or the membrane/total protein ratio. The 

lowest and highest PFOA concentrations caused a significant increase in lipid peroxidation 

(p<0.007), but no effect was observed in cell viability. TH activity was decreased (p<0.05) after 

exposure to 10 and 250 μmol PFOA, and the TH/ChAT ratio was decreased (p<0.05) at 10 μmol

PFOA. The results suggest that PFOA exposure caused the differentiating cells to shift slightly to 

favor the acetylcholine phenotype. 

3.2.5 Developmental/Reproductive Toxicity

Reproductive Effects

A comprehensive two-generation reproductive toxicity study was conducted in SpragueDawley Rats with publication of the results by Butenhoff et al. (2004a). A subsequent study by 

York et al. (2010) provided details of male reproductive organ histopathology. One study in mice 

examined the impact of mating exposed males with unexposed females on fertility and neonatal 

body weight (Lu et al. 2015). 

Rat. A standard oral two-generation reproductive toxicity study of PFOA in Sprague-Dawley

rats was conducted (Butenhoff et al. 2004a). Five groups of male and female SD rats (30 per 

gender per group) were administered PFOA by gavage at doses of 0, 1, 3, 10, and 30 mg/kg/day. 

The parental generation (F0) rats (n = 30 per gender per group) were dosed for 10 weeks prior to 

mating and until sacrificed (after mating for males; after weaning for females). F1 generation rats 

(n = 60 per gender per group) were dosed similarly, beginning at weaning. The F2 generation 

rats were maintained through LD 22. Reproductive parameters evaluated in the F0 and F1 

generations included estrus cyclicity, sperm number and quality, mating, fertility, natural 

delivery, and litter viability and growth. Age at sexual maturation in F1 pups, anogenital distance 

in F2 pups, and presence of nipples (males) in F2 pups also were determined. Food consumption, 

body-weight gain, selected organ weights, gross pathology, and appropriate histopathology of 

reproductive organs were evaluated.

F0 Male Rats 

One F0 male rat in the 30 mg/kg/day dose group was sacrificed on day 45 of the study 

because of adverse clinical signs. Statistically significant increases in clinical signs also were 

observed in male rats in the high-dose group, including dehydration, urine-stained abdominal fur, 

and ungroomed coat. Significant reductions in body weight were reported beginning on postweaning day 50 at 3 mg/kg/day and for most of the study until termination in 10 and 

30 mg/kg/day dose groups (6%, 11%, and 25% decrease from controls, respectively, at the end 

of premating; p≤0.05). Absolute food consumption was significantly reduced to approximately 

91% of the control level during the study in the 30-mg/kg/day dose group but not for the lower

dose groups. Mean food consumption relative to body weight was increased in a dose-related 

manner for all treated males with statistical significance at ≥3 mg/kg/day; overall relative food 

consumption was 101, 105, 110, and 118% of the controls in the 1, 3, 10, and 30 mg/kg/day 

groups, respectively. The body weight and food consumption effects were not observed in 

female rats at any dose. 

Perfluorooctanoic acid (PFOA) – May 2016 3-79
Organ weight data for the F0 male rats is shown in Table 3-19. The absolute and relative-tobody and -brain weights of the liver were statistically significantly increased in all dose groups. 

Absolute kidney weights were statistically significantly increased in the 1-, 3-, and 10-mg/kg/day 

dose groups, but significantly decreased in the 30-mg/kg/day group. Organ weight-to-terminal 

body weight ratios for the left and right kidney were statistically significantly increased in all 

treated groups. Kidney weight-to-brain weight ratios were significantly increased at 1, 3, and 

10 mg/kg/day, but decreased at 30 mg/kg/day, following the trends in absolute weights. In the 

high-dose group, absolute and relative kidney weight changes occurred in a pattern typically 

associated with decrements in body weight. However, in the lower dose groups, consistent

significant increases in absolute kidney weight and relative-to-body and -brain weights are a 

response to the challenge of providing transporters for renal removal of the foreign molecule. 

Increased kidney weight can be regarded as an adaptive response to the transport challenge. It is 

beneficial for the individual but adverse in the sense that it signifies the need to upregulate 

tubular transporters in the kidney to excrete the foreign material and a reflection of PFOA 

bioaccumulation in serum and tissues.

Table 3-19. Organ Weight Data from F0 Male Rats 

0 mg/kg/day 1 mg/kg/day 3 mg/kg/day 10 mg/kg/day 30 mg/kg/day

Body weight (g) 581 ± 40 575 ± 48 542 ± 47** 513 ± 54** 432 ± 64**

Brain weight (g) 2.26 ± 0.17 2.28 ± 0.10 2.26 ± 0.12 2.24 ± 0.12 2.20 ± 0.14

Liver weight (g) 20.3 ± 2.5 24.3 ± 3.2** 27.7 ± 2.7** 28.7 ± 3.9** 27.5 ± 3.7**

Liver/body (%) 3.49 ± 0.29 4.22 ± 0.50** 5.13 ± 0.47** 5.61 ± 0.51** 6.42 ± 0.73**

Liver/brain (%) 903 ± 119 1066 ± 154** 1230 ± 120** 1285 ± 183** 1248 ± 144**

Rt. kidney (g) 2.19 ± 0.18 2.54 ± 0.30** 2.50 ± 0.18** 2.36 ± 0.25** 2.06 ± 0.20*

Rt. kidney/body (%) 0.379 ± 0.030 0.443 ± 0.048** 0.463 ± 0.039** 0.462 ± 0.034** 0.481 ± 0.051**

Rt. kidney/brain (%) 97.5 ± 9.9 111.6 ± 13.5** 111.0 ± 9.5** 105.6 ± 12.4** 93.5 ± 8.7

Lt. kidney (g) 2.19 ± 0.20 2.51 ± 0.28** 2.51 ± 0.21** 2.34 ± 0.24* 1.99 ± 0.19**

Lt. kidney/body (%) 0.378 ± 0.036 0.437 ± 0.047** 0.465 ± 0.043** 0.457 ± 0.040** 0.466 ± 0.054**

Lt. kidney/brain (%) 97.5 ± 10.7 110.1 ± 12.6** 111.7 ± 10.5** 104.6 ± 11.7* 90.4 ± 8.7*

Source: Butenhoff et al. 2004a

Notes: Mean±SD; n = 29–30; significantly different from control: *p < 0.05, **p < 0.01. 

The only histologic finding was increased thickness and prominence of the zona glomerulosa 

and vacuolation in the cells of the adrenal cortex observed in 2/10 males in the 10-mg/kg/day 

dose group and 7/10 males in the 30-mg/kg/day dose group. 

No treatment-related effects were reported at any dose level for any of the male reproductive 

parameters assessed. There was no evidence of altered testicular and sperm structure and 

function in PFOA-treated F0 rats with mean group serum PFOA concentrations of up to 

approximately 45 μg/mL (York et al. 2010). There was a significant dose-related increase in 

seminal vesicle weight (p<0.05) with and without fluid in the F1 males, but fertility of the 

exposed males in all generations was comparable to the controls.

No treatment-related effects were seen at necropsy or upon microscopic examination of the 

reproductive organs. 

Under the conditions of the study, the LOAEL for F0 parental male rats is 1 mg/kg/day, the 

lowest dose tested, based on significant increases in absolute and relative liver and kidney 

weights. A NOAEL for the F0 parental males could not be determined.

Perfluorooctanoic acid (PFOA) – May 2016 3-80
F0 Female Rats 

There were no treatment-related effects on clinical signs, body weight, food consumption, 

organ weights, or histology of the organs. There were no treatment-related effects on any of the 

reproductive parameters assessed, and no treatment-related effects were seen at necropsy other 

than slightly decreased liver weights (p<0.05) at doses of 3 and 10 mg/kg/day, but not 

30 mg/kg/day. No abnormalities were seen with microscopic examination of the reproductive 

organs. The NOAEL for F0 parental females is 30 mg/kg/day, the highest dose tested.

F1 Generation 

Pup body weight on a per-litter basis (genders combined) was significantly reduced (p≤0.01) 

by 8–10% throughout the first 2 weeks of lactation in the 30-mg/kg/day group; at weaning, the 

mean body weight was reduced 4.5%, but the difference was not statistically significant. 

Although there were no effects on the viability and lactation indices, the total number of dead 

pups during lactation was increased in the 30-mg/kg/day groups; the difference was statistically 

significant on LDs 6–8. No other effects were noted, and there were no treatment-related 

findings for the pups necropsied at weaning. The offspring toxicity LOAEL is 30 mg/kg/day 

based on decreased body weight and an increase in the number of dead pups; the NOAEL is 

10 mg/kg/day. 

F1 Male Rats 

Significant increases in treatment-related deaths (5/60 animals) were reported in F1 males in 

the high-dose group between days 2–4 postweaning. One rat was moribund sacrificed on day 39 

postweaning and another was found dead on day 107 postweaning. Clinical signs included a 

significant increase in emaciation at 10 and 30 mg/kg/day, and in urine-stained abdominal fur, 

decreased motor activity, and abdominal distention at 30 mg/kg/day. 

Mean body weight was significantly reduced in the 30-mg/kg/day group beginning on 

postweaning day 8, in the 10-mg/kg/day group beginning on postweaning day 36, and towards 

the end of the study in the 1- and 3-mg/kg/day groups. Terminal mean body weight was reduced 

in all treated groups at the time of sacrifice. For all groups, there was a significant, dose-related 

reduction in mean body weight gain for the entire dosing period (days 1–113). Absolute food 

consumption values were significantly reduced at 10 and 30 mg/kg/day during the entire 

precohabitation period (days 1–70 postweaning), while relative food consumption values were 

significantly increased.

Statistically significant delays in the average day of preputial separation (p≤ 0.01) were 

observed in high-dose animals versus concurrent controls (52.2 days of age versus 48.5 days of 

age, respectively). There were no other effects on any of the reproductive parameters assessed, 

and at necropsy no effects on reproductive organs or fertility were noted (York et al. 2010). 

The absolute and relative weights of the liver were statistically significantly increased in all 

treated groups (p ≤ 0.01). Treatment-related microscopic changes were described as diffuse 

hepatocellular hypertrophy in rats receiving doses of ≥ 3 mg/kg/day. At the same dose levels, 

there were scattered incidences of focal-to-multifocal necrosis and inflammation in the livers of 

the F1 male rats. As in the F0 males, the relative weight of the left and/or right kidneys was

statistically significantly increased compared to controls for all dose groups, except for the right 

kidney at the high dose in which it was lower than for the controls (Table 3-20). Organ weightto-terminal body weight and brain weight ratios for the kidney were statistically significantly 

Perfluorooctanoic acid (PFOA) – May 2016 3-81
increased in all treated groups. All other organ weight changes observed (i.e., thymus, spleen, 

left adrenal, brain, prostate, seminal vesicles, testes, and epididymis) were probably due to 

decrements in body weight and not a reflection of target organ toxicity. Treatment-related 

microscopic changes were observed in the adrenal glands of high-dose animals (i.e., cytoplasmic 

hypertrophy and vacuolation of the cells of the adrenal cortex) and in the liver of rats treated 

with 3, 10, and 30 mg/kg/day (hepatocellular hypertrophy).

Table 3-20. Organ Weight Data from F1 Male Rats 

0 mg/kg/day 1 mg/kg/day 3 mg/kg/day 10 mg/kg/day 30 mg/kg/day

Body weight (g) 560 ± 60 527 ± 55* 524 ± 48* 499 ± 64** 438 ± 42**

Brain weight (g) 2.34 ± 0.13 2.28 ± 0.16 2.31 ± 0.12 2.28 ± 0.10 2.18 ± 0.14**

Liver weight (g) 21.7 ± 3.2 24.6 ± 4.0** 28.2 ± 4.2** 29.3 ± 4.1** 29.7 ± 4.0**

Liver/body (%) 3.86 ± 0.32 4.65 ± 0.51** 5.41 ± 0.75** 5.90 ± 0.70** 6.79 ± 0.55**

Liver/brain (%) 927 ± 136 1075 ± 150** 1224 ± 179** 1285 ± 159** 1364 ± 166**

Rt. kidney (g) 2.24 ± 0.21 2.34 ± 0.28 2.48 ± 0.24** 2.33 ± 0.25 2.04 ± 0.21**

Rt. kidney/body (%) 0.402 ± 0.034 0.446 ± 0.041** 0.474 ± 0.041** 0.469 ± 0.050** 0.467 ± 0.036**

Rt. kidney/brain (%) 95.9 ± 9.1 102.6 ± 7.7** 107.4 ± 10.2** 102.3 ± 9.8* 93.6 ± 7.9

Lt. kidney (g) 2.21 ± 0.20 2.35 ± 0.26* 2.46 ± 0.20** 2.30 ± 0.22 2.03 ± 0.22**

Lt. kidney/body (%) 0.396 ± 0.031 0.446 ± 0.042** 0.472 ± 0.045** 0.464 ± 0.046** 0.465 ± 0.038**

Lt. kidney/brain (%) 94.8 ± 7.9 102.8 ± 7.6** 106.6 ± 9.1** 101.0 ± 7.9* 93.3 ± 10.0

Source: Butenhoff et al. 2004a 

Notes: Mean±SD; n = 29-30; significantly different from control: *p<0.05, **p<0.01.

The LOAEL for adult systemic toxicity in the F1 males is 1 mg/kg/day based on significant, 

dose-related decreases in body weights and body weight gains, and in terminal body weights; 

and significant increases in absolute and relative kidney weights. A NOAEL for adult systemic 

toxicity in the F1 males could not be determined. Liver weights were significantly increased at 

all doses, but only accompanied by microscopic lesions at doses ≥ 3 mg/kg/day. 

F1 Female Rats 

A statistically significant increase in treatment-related mortality (6/60 animals) was observed 

in F1 females on postweaning days 2–8 at the highest dose of 30 mg/kg/day. No adverse clinical 

signs of treatment-related toxicity were reported. Statistically significant decreases in body 

weight were observed in high-dose rats on days 8, 15, 22, 29, 50, and 57 postweaning, during 

precohabitation (recorded on the day cohabitation began, when F1 generation rats were 92–106 

days of age), and during gestation and lactation. Body weight gain was significantly reduced 

during days 1–8 and 8–15 postweaning. Statistically significant decreases in absolute food 

consumption were observed during days 1–8, 8–15, and 15–22 postweaning, during 

precohabitation, and during gestation and lactation in animals treated with 30 mg/kg/day. 

Relative food consumption values were comparable across all treated groups.

Statistically significant delays (p≤0.01) in sexual maturation (the average day of vaginal 

patency) were observed in high-dose animals versus concurrent controls (36.6 days of age versus 

34.9 days of age, respectively). Prior to the rats mating, the study authors noted a statistically 

significant increase in the average numbers of estrous stages per 21 days in high-dose animals 

(5.4 versus 4.7 in controls). For this calculation, the number of independent occurrences of estrus 

in the 21 days of observation was determined. This calculation can be used as a screen for effects 

on the estrous cycle, but should be followed by a more detailed analysis. 

Perfluorooctanoic acid (PFOA) – May 2016 3-82
Both 3M (2002, cited in USEPA [2005c]), and EPA (USEPA 2002b) conducted a more 

detailed analysis of the estrous cycle data. The 3M analysis of the data concluded that there were 

no differences in the number of females with ≥3 days of estrus or with ≥4 days of diestrus in the 

control and high-dose groups. This conclusion is consistent with that of EPA (USEPA 2002b). 

The cycles were evaluated as having either regular 4–5-day cycles, uneven cycling (defined as 

brief periods with irregular pattern) or periods of prolonged diestrus (defined as 4–6-day diestrus 

periods), extended estrus (defined as 3–4 days of cornified smears), possible pseudopregnancy

(defined as 6 or more days of leukocytes), or persistent estrus (defined as 5 or more days of 

cornified smears). The two groups were not different in any of the parameters measured. Thus, 

the increase in the number of estrus stages per 21 days noted by the study authors was an 

outcome of the approach used for the calculations and is not biologically meaningful. There were 

no effects on the other reproductive parameters assessed, and at necropsy no effects on 

reproductive organs were noted. 

No treatment-related effects were observed in the terminal body weights of the F1 female

rats. The absolute weight of the pituitary, the pituitary weight-to-terminal body weight ratio, and 

the pituitary weight-to-brain weight ratio were statistically significantly decreased at 3 

mg/kg/day and higher. Since there is not a clear dose-response relationship and histologic 

examination reveal no lesions, the biological significance of the pituitary weight data is 

problematic. No other differences were reported for the absolute weights or ratios for other 

organs evaluated. No treatment-related effects were reported following macroscopic and 

histopathologic examinations. 

For F1 females, the LOAEL for developmental/reproductive toxicity was considered to be 

30 mg/kg/day based on significantly reduced body weight and body weight gain during lactation, 

a delay in sexual maturation, and increased mortality during postweaning days 2–8; the NOAEL 

was 10 mg/kg/day. The NOAEL and LOAEL for adult systemic toxicity in F1 females are 10 

and 30 mg/kg/day, respectively, based on statistically significant decreases in body weight and 

body weight gains. 

F2 Generation Rats

There was a statistically significant increase (p ≤ 0.01) in the number of pups found dead on 

LD 1 in the 3- and 10-mg/kg/day groups. An independent statistical analysis was conducted by 

EPA (USEPA 2002c), and no significant differences were observed between dose groups and the 

response did not have any trend in dose. There were no treatment-related effects on any of the 

developmental parameters assessed, and at necropsy, no treatment-related effects were noted. 

The NOAEL for developmental/reproductive toxicity in the F2 offspring was 30 mg/kg/day. 

Mouse. In a follow-up to the 28-day component of the Lu et al. study (2015) of the impact on 

PFOA on the blood testes barrier (section 3.2.2), the authors examined the impact exposure to 

0 and 5 mg/kg/day PFOA had on the fertility of the treated male mice (6–8 weeks old; 15 per 

dose group). Each treated male was mated with three virgin ICR females (8–10 weeks old). 

Successful mating was determined by the presence of a vaginal plug in the morning. The 

pregnant females were separated from the males and caged alone throughout the pregnancy; they 

were not dosed with PFOA. At parturition, the pups were counted and the litter body weight 

recorded. There was a statistically significant decrease in the number of mated females per male 

mouse and pregnant females per male mouse (p<0.001) for the exposed males compared to the 

controls. The average number of pups per litter was smaller for the exposed group (10.20 ± 0.72 

versus 11.89 ± 0.54), but the difference was not significant. The average litter weight was 

Perfluorooctanoic acid (PFOA) – May 2016 3-83
significantly lower for the offspring of the paternally exposed mice than of the paternal controls 

(16.17 ± 1.63 g versus 19.95 ± 0.80 g; p<0.05). The 5-mg/kg/day dose was a LOAEL for effects 

on male fertility and the significantly lower body weight of their progeny. 

Developmental Effects

Standard developmental studies in rats and mice found impacts on pup body weight and 

developmental delays. Most studies used the oral route of exposure; one study used inhalation 

exposure to PFOA particulate matter. Some examined the developmental impact of PFOA 

associated with exposures that occurred only during gestation and lactation or during the 

prepubertal period and the association of PPARα with the developmental effect spectrum. A 

substantial number of studies in mice focused on PFOA’s impact on mammary gland 

development. 

3.2.6 Prenatal Development

Rat. Pregnant Sprague-Dawley rats were gavage-dosed with 0, 3, 10, and 30 mg/kg/day PFOA

during days 4–10, 4–15, and 4–21 of gestation, or from GD 4 to LD 21 (Hinderliter et al. 2005; 

Mylchreest 2003). Clinical observations and body weights were recorded daily. On GDs 10, 15, 

and 21, five rats per group per time-point were sacrificed and the number, location, and type of 

implantation sites were recorded. Embryos were collected on day 10, and placentas, amniotic 

fluid, and embryos/fetuses were collected on days 15 and 21. Maternal blood samples were 

collected at 2 hours ± 30 mins post-dose. The remaining five rats per group were allowed to 

deliver. On LDs 0, 3, 7, 14, and 21, the pups were counted, weighed (genders separate), and 

examined for abnormal appearance and behavior. Randomly selected pups were sacrificed and 

blood samples were collected. On LDs 3, 7, 14, and 21, the dams were anesthetized and milk and 

blood samples were collected; dams were removed from their litters 1–2 hours prior to 

collection. Plasma, milk, amniotic fluid extract, and tissue homogenate (i.e., placenta, embryo, 

and fetus) supernatants were analyzed for PFOA concentrations by HPLC/MS.

All dams survived and there were no clinical signs of toxicity. In the 30-mg/kg/day group, 

mean body weight gain was approximately 10% lower than in the control group during gestation, 

and mean body weight was approximately 4% lower than for controls throughout gestation and 

lactation. The number of implantation sites, resorptions, and live fetuses were comparable among 

groups on days 10, 15, and 21 of gestation. One dam in the 3-mg/kg/day group and two dams in 

the 30-mg/kg/day group delivered small litters (3–6 pups per litter compared to 12–19 pups per 

litter in the control group); however, given the small sample size, the biological significance of 

this finding is unclear. There were no clinical signs of toxicity in the pups, and pup survival and 

body weights were comparable among groups. Under the conditions of this study, the maternal 

LOAEL was 30 mg/kg/day for decreased body weight gain during gestation, and the NOAEL 

was 10 mg/kg/day. The developmental NOAEL was 30 mg/kg/day. 

Mouse. Lau et al. (2006) conducted a developmental toxicity study of PFOA in mice to ascertain 

whether there was a gender-related difference in the bioaccumulation of PFOA in the mouse and 

to evaluate the effects of PFOA on prenatal and postnatal development in offspring exposed 

during pregnancy. In that study, groups averaging 9–45 timed-pregnant CD-1 mice were given 0, 

1, 3, 5, 10, 20, and 40 mg/kg PFOA daily by oral gavage on GDs 1–17. Maternal weight was 

monitored during gestation. Dams were divided into two groups. In the first group, dams were 

sacrificed on GD 18 and underwent maternal and fetal examinations that included measure of 

maternal liver weight and examination of the gravid uterus for numbers of live and dead fetuses 

Perfluorooctanoic acid (PFOA) – May 2016 3-84
and resorptions. Maternal blood was collected and analyzed for PFOA serum concentration. 

PFOA levels in the fetuses were not examined. Live fetuses were weighed and subjected to 

external gross necropsy and skeletal and visceral examinations. In the second group of dams, an 

additional dose of PFOA was administered on GD 18. Dams were allowed to give birth on 

GD 19. 

The day following parturition was designated as PND 1. Time of parturition, condition of 

newborns, and number of live offspring were recorded. The number of live pups in each litter 

and pup body weight were noted for the first 4 days after birth and then at corresponding 

intervals thereafter. Eye opening was recorded beginning at PND 12. Pups were weaned on PND 

23 and separated by gender. The time to sexual maturity was determined by monitoring vaginal 

opening and preputial separation beginning on PND 24. Two to four pups per gender per litter 

were randomly selected for observation of postnatal survival, growth, and development. Estrous 

cyclicity was determined daily by vaginal cytology. After weaning, dams were sacrificed and the 

contents of the uteri examined for implantation sites. Postnatal survival was calculated based on 

the number of implantations for each dam. 

Signs of maternal toxicity were observed following exposure to PFOA during pregnancy. 

Statistically significant dose-related increases (p ≤ 0.05) in maternal liver weight also were 

observed, beginning at 1 mg/kg/day. Dose-related decreases in maternal weight gain during 

pregnancy were observed beginning at 5 mg/kg/day, with statistical significance (p ≤ 0.05) seen 

in the 20- and 40-mg/kg/day dose groups. Under the conditions of the study, a maternal LOAEL 

of 1 mg/kg was indicated based on increased liver weight, and a NOAEL was not established. 

Signs of developmental toxicity were observed following in utero exposure to PFOA. The 

number of implantations in treated mice was comparable to control mice. Statistically significant 

increases (p ≤ 0.05) in full-litter resorptions were reported at doses of ≥ 5 mg/kg/day, with 

complete loss of pregnancies at the highest dose group of 40 mg/kg/day. A 20% reduction 

(p ≤ 0.05) in live fetal body weight at term was reported at 20 mg/kg/day. A statistically 

significant increase in prenatal loss was observed in the 20-mg/kg/day dose group. Ossification 

(number of sites) of the forelimb proximal phalanges was significantly reduced at all doses 

except 5 mg/kg. Ossification of hindlimb proximal phalanges was significantly reduced at all 

doses except 3 and 5 mg/kg. Reduced ossification (p ≤ 0.05) of the calvaria and enlarged 

fontanel was observed at 1, 3, and 20 mg/kg and at ≥ 10 mg/kg in the supraoccipital bone. 

Statistically significant increases (p ≤ 0.05) in minor limb and tail defects were observed in the 

fetuses at doses ≥ 5 mg/kg/day. Under the conditions of the study, a prenatal developmental 

LOAEL of 1 mg/kg was indicated based on increased skeletal defects, and the NOAEL was not 

established. 

Slight, but statistically significant, increases (p ≤ 0.05) in the average time to parturition were 

observed at 10 and 20 mg/kg/day. Increases (p ≤ 0.05) in stillbirths and neonatal mortality (or 

decreases in postnatal survival) were observed at doses ≥ 5 mg/kg/day, with as much as a 30% 

increase in these effects seen in the 10- and 20-mg/kg/day dose groups. Postnatal survival and 

viability in the 1- and 3-mg/kg/day dose groups were comparable to controls. At doses 

≥ 3 mg/kg/day, a trend in growth retardation (body weight reductions of 25–30%; p ≤ 0.05), was 

observed in the neonates at weaning. Body weights were at control levels by 6 weeks of age for 

females and by 13 weeks of age for males. A trend for increasing body weight (~6–10% greater 

than controls) was observed in animals dosed with 5 mg/kg at 13 weeks and in animals dosed 

with 1 and 3 mg/kg at 48 weeks. Deficits in early postnatal growth and development also were 

manifested by significant delays (p ≤ 0.05) in eye opening at doses ≥ 5 mg/kg/day. Slight delays 

Perfluorooctanoic acid (PFOA) – May 2016 3-85
(p ≤ 0.05) in vaginal opening and in time to estrous were observed at 20 mg/kg/day in females; 

in contrast, significant accelerations (p ≤ 0.05) in sexual maturation were observed in males, with 

preputial separation occurring 4 days earlier than controls at the 1-mg/kg/day dose and 2–3 days 

earlier in the 3–10-mg/kg/day dose groups, but the 20-mg/kg/day dose group was only slightly 

delayed compared to controls. Under the conditions of the study, a LOAEL for developmental 

toxicity of 1 mg/kg/day for males was indicated based on accelerated pubertal development, and 

a NOAEL was not established. For females, the developmental LOAEL was 3 mg/kg/day based 

on growth retardation, and the NOAEL was 1 mg/kg/day. 

Values for the benchmark dose (BMD) and the lower 95th percentile confidence bound on the 

BMD (BMDL) for the maternal and developmental endpoints (BMD5 and BMDL5) were 

calculated by the study authors and reported in Lau et al. (2006). For maternal toxicity, BMD5

and BMDL5 estimates for decreases in maternal weight gain during pregnancy were 6.76 and 

3.58 mg/kg/day, respectively. For increases in maternal liver weight at term, BMD5 and BMDL5 

estimates were 0.20 mg/kg/day and 0.17 mg/kg/day, respectively. BMD5 and BMDL5 estimates 

for the incidence of neonatal mortality (determined by survival to weaning) at 5 mg/kg/day were 

2.84 and 1.09 mg/kg/day, respectively. Significant alterations in postnatal growth and 

development were observed at 1 and 3 mg/kg/day, with BMD5 and BMDL5 estimates of 1.07 and 

0.86 mg/kg/day, respectively, for decreased pup weight at weaning; and 2.64 and 2.10 

mg/kg/day, respectively, for delays in eye opening. The BMD5 and BMDL5 estimates for reduced 

phalangeal ossification were < 1 mg/kg/day. BMD5 and BMDL5 estimates for reduced fetal 

weight at term were estimated to be 10.3 and 4.3 mg/kg/day, respectively.

Male and female 129S1/SvImJ and PPARα-null mice were used in studies to determine if 

PFOA-induced developmental toxicity was mediated by PPARα (Abbott et al. 2007). Pregnant 

129S1/SvlmJ wild-type and PPARα-null mice were orally dosed from GD 1–17 with 0, 0.1, 0.3, 

0.6, 1, 3, 5, 10, and 20 mg PFOA/kg/day. Heterozygous (HET) litters also were produced by 

mating wild-type and PPARα-null males with wild-type and PPARα-null dams to determine if 

genetic background affected survival. The HET litters were sacrificed on PND 15. Survival at 

birth was recorded and live offspring counted and weighed by gender. Litters were counted and 

offspring weighed on PND 1–10, 14, 17, and 22. Weaning occurred on PND 22, and all dams 

and one pup per litter were sacrificed. Blood was collected and the uteri were stained for 

implantation counts. 

There was no effect of treatment on maternal weight or maternal weight gain (excluding 

nonpregnant females and those with full-litter resorptions), number of implants, or pup weight at 

birth. Wild-type dams exposed to ≥0.6 mg/kg/day and PPARα-null dams exposed to ≥5 

mg/kg/day had a significantly greater percentage of litter loss compared to their respective 

controls. At ≥5 mg/kg/day in wild-type dams and 20 mg/kg/day in PPARα-null dams, 100% 

litter loss occurred. Relative liver weight was significantly increased in wild-type adult females 

dosed with ≥1 mg/kg/day and in PPARα-null adult females dosed with ≥3 mg/kg/day. 

Body weight in wild-type offspring born of dams dosed with 1.0 mg/kg/day was significantly 

reduced (p<0.05) compared to control offspring body weight gain on PND 9, 10, and 22 (males) 

and PND 7–10 and PND 22 (females). No differences were observed between PPARα-null 

offspring body weight and control offspring body weight. Survival of pups from birth to weaning 

was significantly reduced (p<0.05) in wild-type litters exposed to ≥0.6 mg/kg/day, but was not 

affected in PPARα-null litters. Survival was significantly decreased (p<0.05) for wild-type and 

HET pups born to wild-type dams dosed with 1 mg/kg/day and for HET pups born to PPARαnull dams dosed with 3 mg/kg. Offspring born of wild-type dams showed a dose-related trend for 

Perfluorooctanoic acid (PFOA) – May 2016 3-86
delayed eye opening compared to control offspring (significantly delayed at 1 mg/kg/day, 

p<0.05), but no difference in day of eye opening was observed in the offspring born of PPARαnull dams. At weaning, relative liver weight was significantly increased (p<0.05) in wild-type

offspring gestationally exposed to ≥0.1 mg/kg/day and in PPARα-null offspring gestationally 

exposed to 3 mg/kg/day. 

The authors concluded that survival of PPARα-null pups and deaths of HET pups born to 

PPARα-null dams indicates that expression of PPARα is required for PFOA-induced postnatal 

lethality; however, early prenatal lethality was independent of PPARα. Delayed eye opening and 

reduced postnatal weight gain appeared to be mediated by PPARα, but other mechanisms might 

also contribute. Under the conditions of the study, the maternal/reproductive LOAEL for wildtype mice was 0.6 mg/kg/day based on increased percentage of litter loss, and the NOAEL was 

0.3 mg/kg/day. The developmental LOAEL for wild-type offspring was 0.1 mg/kg/day based on 

increased liver weight, and the NOAEL was not established. The maternal LOAEL for PPARαnull mice was 3 mg/kg/day based on increased liver weight, and the NOAEL was 1 mg/kg/day. 

The developmental LOAEL for PPARα-null offspring was 3 mg/kg/day based on increased liver 

weight, and the NOAEL was 1 mg/kg/day. 

To further evaluate the developmental effects potentially mediated by PPARα, groups of 

female wild-type, PPARα-null, and PPARα-humanized mice were given 0 and 3 mg PFOA/kg 

on GDs 1–17 by oral gavage (Albrecht et al. 2013). Controls received the water vehicle. Females 

were either sacrificed on GD 18 (n = 5–8 per group) or allowed to give birth and then sacrificed, 

along with their litters (n = 8–14), on PND 20. Livers from dams, fetuses, and pups were 

weighed and collected for histopathological evaluation and RNA analysis. Gene expression 

results are given in section 3.3.4. Mammary gland whole mounts were prepared from female 

pups on PND 20 for quantification of ductal length and number of terminal end buds; these 

results are described below with other studies evaluating the effects of PFOA on mammary gland 

development. 

Evaluation on GD 18 showed no effects of PFOA administration on maternal body weight, 

body weight gain, gravid uterine weight, number of implantations per dam, or number of 

resorptions per litter in dams of any genotype. For animals allowed to litter, the average day of 

parturition was slightly later in PFOA-treated humanized mice than in the controls. Body weight 

of dams during lactation, the number of pups born per litter, pup body weight during lactation, 

and the onset of pup eye opening were similar between treated and control groups for all 

genotypes. Offspring survival during PNDs 1–5 was significantly reduced in the wild-type 

PFOA-treated group, but not in the other genotypes. 

Maternal liver weight was significantly increased in the treated groups of all genotypes on 

GD 18 and in wild-type animals on PND 20. Maternal liver weight was not affected on PND 20 

in the PPARα-null or PPARα-humanized mice. Relative fetal liver weight on GD 18 was 

significantly increased in fetuses from treated wild-type and humanized dams. On PND 20, 

relative liver weight was increased only in pups from treated wild-type dams. Microscopic 

evaluation of the maternal liver showed centrilobular hepatocellular hypertrophy in all PFOAtreated groups on GD 18 and PND 20, with decreased incidence and severity by PND 20. On GD

18, the liver lesions were graded as mild in the wild-type mice, minimal-to-mild in the 

humanized mice, and minimal in the null mice. The morphological features of the liver lesions 

differed slightly between genotypes and are described in more detail in section 3.4.1. Only wildtype fetuses and pups from treated dams showed similar liver lesions.

Perfluorooctanoic acid (PFOA) – May 2016 3-87
Yahia et al. (2010) gavage-dosed pregnant ICR mice (n = 5 per group) with 0, 1, 5, and 

10 mg PFOA/kg/day from GDs 0–17 or 0–18. The dams dosed from GDs 0–17 were sacrificed 

on GD 18, and the fetal skeletal morphology was evaluated. Dams dosed from GDs 0–18 were 

allowed to give birth and their offspring were either processed for pathological examination or 

observed for 4 days for neonatal mortality. Maternal liver, kidney, brain, and lungs were 

histologically examined after necropsy. Serum was collected for clinical chemistry and lipid 

analysis. Body weight was significantly decreased in dams receiving 10 mg/kg/day. Maternal 

absolute liver weight was significantly increased (p<0.05) at doses ≥ 5 mg/kg/day and relative 

liver and kidney weights were significantly increased at all doses. Hepatic hypertrophy was 

localized to the centrilobular region at the two lower doses and was diffuse at the highest dose.

Renal cells in the outer medullar and proximal tubule were slightly hypertrophic at all doses.

Treatment with 10 mg/kg/day caused a significant increase in AST, ALT, GGT, and ALP and a 

significant decrease in total serum protein, albumin, globulin, triglycerides, phospholipids, TC, 

and free fatty acids. At 5 mg/kg/day, total serum protein and globulin were significantly 

decreased, and phosphorus was increased. At 1 mg/kg/day, BUN and phosphorus were 

significantly increased. The maternal LOAEL was 1 mg/kg/day based on significantly increased 

relative liver and kidney weight, and no NOAEL was established. 

Live fetal birth weight was significantly decreased at the two highest doses. There was no 

difference in the percentage of live fetuses between treated and control groups. At 10 mg/kg/day, 

increased incidence of cleft sternum, reduced phalanges ossification, and delayed eruption of 

incisors was observed. Delayed parturition was observed in dams treated with 10 mg/kg/day, and 

~58% of all pups born to those dams were stillborn. Death occurred within 6 hours of delivery in 

the remaining pups, and whole body edema was observed in some of the pups. The body weight 

of the live pups born to dams treated with 5 or 10 mg/kg/day was significantly reduced compared 

to control pup body weight. By PND 4, 16% of offspring born to dams dosed with 5 mg/kg/day 

had died. No pathological differences were observed in the lungs or brains of treated and control 

offspring. The developmental LOAEL was 5 mg/kg/day based on decreased body weight and 

decreased survival rate, and the NOAEL was 1 mg/kg/day. 

Suh et al. (2011) examined placental prolactin-family hormone and fetal growth retardation 

in mice treated with PFOA. Pregnant CD-1 mice (n = 10 per group) were treated with 0, 2, 10, 

and 25 mg/kg/day PFOA from GDs 11–16. Dams were sacrificed on GD 16 and uteri were 

removed and examined. Three placentas per group were analyzed histochemically and the 

numbers of glycogen trophoblast cell (GlyT) in the junctional zone plus sinusoidal trophoblast 

giant cells (S-TGC) in the labyrinth zone were counted and compared. Trophoblast cells express 

prolactin-family genes. mRNA from three placentas per group were analyzed using situ 

hybridization, northern blot hybridization, and RT-PCR for mouse placental lactogen- (mPL-) II,

mouse prolactin-like protein- (mPLP-) E and F, Pit-1α, and β isotype (transacting factors of mPL 

and mPLP genes).

A significant difference in maternal body weight was observed from GD 13–16 in dams 

treated with 25 mg/kg/day of PFOA compared to controls. At ≥2 mg/kg/day of PFOA, placental 

weight was significantly decreased and the number of resorptions and dead fetuses was 

significantly increased. At ≥10 mg/kg/day of PFOA, fetal weight and the number of live fetuses 

were significantly decreased. There were no differences in the number of implantation sites

among the groups, and postimplantation loss was 3.87, 8.83, 30.98, and 55.41% for the 0, 2, 10, 

and 25 mg/kg/day PFOA groups, respectively. 

Perfluorooctanoic acid (PFOA) – May 2016 3-88
The placentas of dams dosed with ≥10 mg/kg/day of PFOA displayed necrotic changes.

Parietal and S-TGC and GlyT cell frequency in the placental junctional and labyrinth zones was 

significantly decreased (p<0.05) in a dose-dependent manner in treated dams. At 25 mg/kg/day 

of PFOA, S-TGCs showed signs of atrophy with crushed cell nucleus. A significant dosedependent decrease in mPL-II, mPLP-E, mPLP-F, and Pit-1α and β isotype mRNA and 

expression was observed. Correlation coefficients between fetal weight and maternal mPL-II, 

mPLP-E, and mPLP-F mRNA levels were positive (p<0.001). Based on the results, the authors 

suggested that inhibited prolactin-family gene expression could be secondary to insufficient 

trophoblast cell differentiation and increased cell necrosis. These effects reduced placental 

efficiency and contributed in part to fetal growth retardation. The 2-mg/kg/day dose was a 

LOAEL for increases in resorption and dead fetuses plus decreased placental weight. There was 

no NOAEL.

A meta-analysis was conducted to determine whether developmental exposure to PFOA was 

associated with fetal growth effects in animals (Koustas et al. 2014). Eight studies identified in 

the published literature met the criteria of the Navigation Guide systematic review methodology 

for inclusion in the analysis (Woodruff and Sutton 2014). The data sets included mouse gavage 

studies with offspring body weight at birth. Maternal PFOA doses ranged from 0.01 to 

20 mg/kg/day. The results from the meta-analysis showed that a 1-mg/kg/day increase in PFOA 

dose was associated with a -0.023 g (95% CI -0.029, -0.016) difference in pup birth weight. All

of the studies included in the analysis are described either above with the developmental toxicity 

studies or with the specialized developmental studies that follow.

3.2.7 Mammary Gland Development and Other Specialized Developmental Studies

The following studies used experimental study designs and/or examined endpoints not 

typically included in standard developmental toxicity studies. The studies were conducted to 

determine critical periods of exposure for outcomes that occurred later in life. A number of the 

studies focused on mammary gland development in dams and female offspring. Researchers 

focused on effects resulting from indirect exposure of offspring via treatment of pregnant 

animals and/or direct exposure of peripubertal animals starting about the time of weaning. 

Indirect gestational and/or lactational exposures 

Many studies evaluating indirect gestational and/or lactational exposure to PFOA are 

available and Table 3-21 provides an overview of experiments designed to assess the 

developmental effects of PFOA following exposures during gestation. Most of the studies focus 

on mammary gland effects as a consequence of gestational and lactational or prepubertal 

exposures in CD-1 mice. Some have included postnatal assessment of the liver. Additional 

details of the studies are described in the section following the table. 

Timed pregnant CD-1 mice were given PFOA by gavage at doses of 0, 0.01, 0.1, 0.03, and 

1 mg/kg/day from GD 1 through the end of lactation (PND 21) in a study by Quist et al. (2015). 

The litters were equalized on PND 4 to four females and six males. Only the females continued 

in the study after weaning. At the end of lactation, litters with less than four females were 

removed from the study. On PND 21, seven to 10 female pups per dose group were sacrificed by 

decapitation. The livers were removed for analysis.

Perfluorooctanoic acid (PFOA) – May 2016 3-89
Table 3-21. Studies of Pregnant CD-1 Mice Following Administration of PFOA

Dose

(mg PFOA 

/kg/day) Timing Endpoints Reference

0, 0.01. 0.1, 

0.3, 1 

GD 1 to PND 21 Liver histopathology; periportal inflammation; 

clinical chemistry; impact of postweaning HFD

Quist et al. 2015

0, 5 GD 1-17, 8-17, or 12-17 Body weight; mammary gland morphology GD 

18 (dams) and PNDs 10 and 20 (dams, female 

pups)

White et al. 2007

0, 5

20

GD 7-17, 10-17, 13-17, 

15-17

GD 15-17

Body weight; developmental landmarks and 

growth to PND 189; mammary gland 

morphology of female pups up to 18 months

White et al. 2009; 

Wolf et al. 2007

0, 3, 5 GD 1-17

Cross-fostered at birth

Body weight; developmental landmarks and 

growth to PND 245; mammary gland 

morphology of female pups up to 18 months

White et al. 2009; 

Wolf et al. 2007;

0, 5 GD 8-17

Cross-fostered at birth

Mammary gland morphology of dams and female 

pups on PNDs 1, 3, 5, and 10

White et al. 2009

0, 0.3, 1.0, 3 GD 1-17 Liver weight; mammary gland morphology of 

female pups on PNDs 7, 14, 21, 28, 42, 63, and 

84

Macon et al. 2011

0, 0.01, 0.1, 1 GD 10-17

0, 1, 5 GD 1-17 Body weight; reproductive parameters; 

mammary gland morphology of F0, F1, and F2 

females

White et al. 2011

0, 1

+ 5 ppb in 

drinking water 

to both groups

GD 1-17

Drinking water started on 

GD 7 and continued to F2 

generation

0, 3 GD 1-17 Wild-type, PPARα-null, and hPPARα vv/129 

mice; pup body weight at PNDs 14 and 20 plus 

mammary gland structure

Albrecht et al. 

2013

0, 0.01, 0.1, 

0.3, 1

GD 1-17

Study included both CD-1 

and C57BL/6 mice

Body weight; net body weight; absolute and 

relative liver weight on PNDs 21, 35, and 56; 

serum estradiol and progesterone; mammary 

gland morphology

Tucker et al. 

2015

During the lactation period all pups received a Purina control diet with a normal fat content. 

At sacrifice on PND 21, there was a dose-related increase in relative liver weight at doses 

≥0.3 mg/kg/day. When the animals from the same dose group were sacrificed on PND 91, there 

was no observed impact on relative liver weight. 

Starting on PND 35, one pup from each of 20 dams was placed on a HFD (with 60% of the 

calories from fat) until sacrificed on PND 91. Half were fasted for 4 hours before blood 

collection and sacrifice. Another seven to 10 pups per dose group received the Purina control diet 

with lower fat content until their sacrifice on PND 91. Blood samples were collected at sacrifice 

for determination of ALP, AST, ALT, SDH, LDL, HDL, cholesterol, triglycerides, total bile acid 

(TBA), glucose, leptin and insulin. Liver sections were collected for histological analysis and 

graded 1 to 4 for lesion severity (1 = minimal; 4 = severe). Selected samples of the liver tissues 

(four from the HFD groups and four from the PFOA-exposed control diet group) were fixed in 

osmium trioxide and prepared for evaluation using transmission electron microscopy. 

At PND 91, the animals on the PFOA + HFD weighed more than the Purina controls. The 

body weight for the group that received PFOA and the control diet did not differ from the 

untreated controls on the same diet. Serum samples from the PFOA-treated Purina controls, and 

the fasted high-fat and nonfasted high-fat groups were analyzed for LDL, HDL, TC, and leptin. 

Perfluorooctanoic acid (PFOA) – May 2016 3-90
At PFOA doses < 0.3 mg/kg/day, the LDL and TC levels in the fasted and nonfasted HFD

animals were greater than in the untreated Purina controls. There was no impact of PFOS on 

either parameter in animals on the PFOA plus Purina control diet, but both LDL and TC were 

statistically lower at the high PFOA dose than they were at the low doses for both parameters. A 

similar pattern was seen for the HDL levels.

The impact of PFOA dose on leptin was variable and not significant for the PFOA plus 

Purina control animals and the high-fat, fasted animals. For the high-fat, nonfasted animals, there 

was a trend towards decreasing leptin as the PFOA dose increased, which was significant at the 

high dose of 1 mg/kg/day (P< 0.01). In those animals, the liver showed chronic periportal 

inflammation and microvescicular intracytoplasmic lipid droplets. The transmission electron 

microscopy slides showed that the hypertrophic liver cells presented evidence of cellular damage 

and changes in both mitochondrial morphology and numbers. The observed mitochondrial 

abnormalities were not those generally associated with PPARα activation. The 0.01 mg/kg/day 

dose was a NOAEL. The LOAEL was 0.3 mg/kg/day for the effects on TC for animals receiving 

a HFD, but not for those receiving the PFOA plus Purina control diet. 

Effects of PFOA exposure on mammary gland morphology of CD-1 mice were evaluated in a 

series of studies by the same researchers (Macon et al. 2011; Tucker et al. 2015; White et al. 

2007, 2009, 2011). The focus was on mammary gland development of female pups, although 

limited evaluations were conducted on the dams. Mammary gland whole mounts were scored on 

a 1 to 4 subjective, age-adjusted, developmental scale (1 = poor development/structure; 

4 = normal development/structure). Tissue was assessed qualitatively for the gross presence of 

several histological criteria by two independent scorers and a mean score calculated. Neither 

standardization of these subjective measures nor training of the scorers was described in the 

publications. Quantitative measures of longitudinal growth, lateral growth, and number of 

terminal end buds also were made in the Macon et al. (2011) and Albrecht et al. (2013) studies. 

White et al. (2007) orally dosed pregnant CD-1 mice with 0 and 5 mg PFOA/kg/day on GD 

1–17 (n = 14), 8–17 (n = 16), and 12–17 (n = 16) to determine if decreased neonatal body 

weights and survival were linked to gestational exposure or lactational changes in milk quantity 

or quality. The control mice (n = 14) were dosed with vehicle on GD1–17. A subset of dams was 

sacrificed on GD 18. The remaining dams were allowed to give birth, and pups were pooled and 

randomly redistributed among the dams of the respective treatment groups. Litters were 

equalized to 10 pups per litter. Half of the dams and litters were sacrificed on PND 10, and the 

remaining dams and litters were sacrificed on PND 20. Mammary glands were collected from 

dams and female pups at time of sacrifice.

Treatment with PFOA did not affect maternal weight gain, number of implants, or the 

number of live fetuses. There was a significant increase (p<0.05) in prenatal loss in dams 

exposed during GD1–17. Body weight of pups exposed gestationally to PFOA was significantly 

decreased (p<0.05) at all time points measured and for all dosing regimens. On GD 18, stunted 

alveolar development was observed in the mammary gland of dams treated with PFOA on GD 

1–17 compared to the mammary glands of the control dams, which were saturated with milkfilled alveoli. Dams treated with PFOA on GD 1–17 or 8–17 exhibited significant mammary 

gland epithelial differentiation delays on PND 10 as evidenced by an excessive amount of 

adipose tissue. In comparison, mammary glands from control dams on PND 10 were well 

differentiated, full of alveoli filled with milk, and contained few apoptotic bodies and little 

adipose tissue. The mammary gland developmental score in dams treated on GD 12–17 was not 

Perfluorooctanoic acid (PFOA) – May 2016 3-91
statistically different from control dams on PND 10. At PND 20, the mammary gland scores 

from all PFOA-treated dams were not significantly different from those of the control group. 

The pups were impacted by their in utero PFOA exposure over all dosing intervals. Their 

mammary glands exhibited significantly stunted epithelial branching and longitudinal growth at 

PNDs 10 and 20; the resulting developmental scores were significantly less than those of 

controls. Very little mammary gland development occurred between PND 10–20 in the offspring 

of dams exposed to PFOA, even though postnatal growth and body weight gain increased in 

parallel to that of the controls. Thus, at the only dose tested, 5 mg/kg/day, effects were observed 

on the dam and pup mammary gland, accompanied by decreased pup body weight and decreased 

survival for the pups exposed during GD 1–17. 

In the study by Wolf et al. (2007), CD-1 mice were orally dosed with 0 and 5 mg 

PFOA/kg/day on GD 7–17 (n = 14), 10–17 (n = 14), 13–17 (n = 12), and 15–17 (n = 12) or with 

20 mg/kg on GD 15–17 (n = 6) to determine if there was a specific window during which PFOA 

exposure produced developmental effects. The developmental results from this study were 

published by Wolf et al. (2007) and the mammary gland effects were published by White et al. 

(2009). On PND 22, all dams and one male and female pup from each litter were necropsied.

Blood samples were collected and livers from dams and offspring were removed and weighed. 

Uterine implantation sites were counted. The fourth and fifth inguinal mammary glands were 

removed from female offspring and analyzed at various intervals up to 18 months of age (White 

et al. 2009). Mammary gland whole mounts from female offspring between PNDs 22 and 32 

were scored as described above; whole mounts from lactating dams and female offspring at 18 

months were qualitatively examined with respect to concurrent controls. 

Maternal weight gain was increased in dams exposed to PFOA beginning on GDs 7 and 10, 

but there was no effect on number of uterine implantation sites, litter loss, or number of pups per 

litter at birth. Male pup weight at birth was significantly decreased (p<0.05) in dams dosed with 

5 mg/kg/day on GD 7–17 or 10–17 or with 20 mg/kg/day on GD 15–17. By PND 78, all male 

offspring had recovered to control body weight levels. On PND 161, the offspring of dams dosed 

during GD 13–17 weighed significantly more than control. Litters exposed to 20 mg/kg/day on 

GD 15–17 experienced decreased survival (p<0.01) during PND 1–22. Maternal relative liver 

weight was significantly increased in all PFOA-treated dams except those treated during GD 

15–17. Relative liver weight in all male and female pups was significantly increased (p<0.01).

Eye opening and growth of body hair were delayed in pups exposed GD 7–17 and 10–17 (Wolf 

et al. 2007). 

Mammary gland developmental scores for all offspring of dams exposed to PFOA were 

significantly lower compared to scores from offspring of control dams at PND 29 and 32. 

Delayed ductal elongation and branching and delayed appearance of terminal end buds were 

characteristic of delayed mammary gland development at PND 32. At 18 months of age, 

mammary tissues were not scored due to a lack of a protocol applicable to mature animals. 

However, there were dark foci (composition unknown) in the mammary tissue that occurred at a 

higher frequency in the exposed animals compared to controls, but did not display a consistent 

response with dosing interval. Qualitatively, mammary glands from treated dams on LD 1 

appeared immature compared with control dams (White et al. 2009). The 5-mg/kg/day dose was 

associated with increased maternal and pup liver weight, altered pup mammary gland 

development, and delayed pup eye opening and growth of body hair. The 20-mg/kg/day dose 

was associated with decreased postnatal pup survival. 

Perfluorooctanoic acid (PFOA) – May 2016 3-92
The objective of a second component of the study by Wolf et al. (2007) and White et al. 

(2009) was to determine if postnatal body weight deficits, neonatal lethality, and developmental 

delays caused by PFOA exposure were the result of gestational exposure, lactational exposure, or 

a combination of gestational and lactational exposure. Pregnant CD-1 mice were orally dosed 

with 0 (n = 48), 3 (n = 28), and 5 (n = 36) mg PFOA/kg/day on GD 1–17 and their offspring 

cross-fostered at birth to create seven treatment groups: control, in utero exposure only (3U and 

5U), lactation exposure only (PFOA stored in milk during gestation and released during 

lactation; 3L and 5L), and in utero and lactation exposure (3U+L and 5U+L). On PND 22, all 

dams and one male and female pup from each litter were necropsied. Blood samples were 

collected and the liver was removed from dams and offspring and weighed. Implantation sites 

were counted from the uteri of dams. The fourth and fifth inguinal mammary glands were 

removed from female offspring and analyzed at various intervals up to 18 months of age (White 

et al. 2009). Mammary gland whole mounts from female offspring between PND 22 and 63 were 

scored as described above; whole mounts from female offspring at 18 months were qualitatively 

examined with respect to concurrent controls due to lack of an applicable protocol for mature 

animals. 

Maternal weight and weight gain were higher in PFOA-treated dams compared to control 

dams. Whole litter loss was significantly increased (p<0.05) at 5 mg/kg/day, but no differences 

in the number of implantation sites were observed between the treated and control mice.

Absolute and relative liver weights of PFOA-treated dams from both dose groups were 

significantly increased (p<0.001) compared to absolute and relative liver weights of control dams 

23 days after the last dose (PND 22). No difference in the number of live pups born per litter was 

found between treated and control mice, but male and female pup birth weight was reduced 

(p<0.01) in dams receiving 5 mg/kg/day (Wolf et al. 2007). The 3-mg/kg/day dose was a 

LOAEL for increases in liver weight in the dams while 5 mg/kg/day was a LOAEL for the pups, 

based on whole litter loss and significantly reduced male and female birth weight. 

A dose-dependent increase of PFOA was observed in the serum of dams treated with PFOA, 

providing a reservoir for lactational transfer. The control dams that nursed offspring exposed in 

utero (3U and 5U) had low concentrations of PFOA in their serum that originated from maternal 

grooming behavior of the pups and allowed for low-level lactational transfer.

Body weight of male and female pups (3U+L, 5U, and 5U+L) was significantly reduced as 

early as PND 2 and 1, respectively, and remained reduced throughout the lactation period. Body 

weight recovery to control levels was reached by male offspring within 2 weeks of weaning, but 

recovery in female offspring in the 5U and 5U+L groups did not occur until after PND 85. 

Postnatal survival in 5U+L pups was significantly decreased compared to control survival 

beginning at PND 4 and continuing throughout lactation. Survival in the other groups was no 

different than control survival. Eye opening and body hair growth were significantly delayed in 

the 3U+L, 5U, and 5U+L offspring. The relative liver weight was significantly increased in all 

offspring regardless of exposure scenario (Wolf et al. 2007). 

All female offspring of PFOA-exposed dams had reduced mammary gland developmental 

scores at PND 22, except for females in the 3L group. At PND 42, mammary gland scores from 

females in the 3U+L group were the only ones not statistically different from control scores. This 

might have been due to interindividual variance and multiple criteria used to calculate mammary 

gland development scores. All offspring of dams exposed to PFOA exhibited delayed mammary 

gland development at PND 63, including those exposed only through lactation (3L and 5L). A 

Perfluorooctanoic acid (PFOA) – May 2016 3-93
higher density of dark staining foci was observed in the mammary glands of these animals at 

18 months of age (White et al. 2009). 

White et al. (2009) also reported the results from pregnant CD-1 mice orally dosed with 0 

(n = 56) and 5 (n = 56) mg PFOA/kg/day from GD 8–17 to determine the timing of the 

mammary gland development deficits observed following gestational or lactational exposure to 

PFOA. The groups were cross-fostered at birth to create four treatment groups: control, in utero

exposure only (5U), lactation exposure only (5L), and in utero and lactation exposure (5U+L). 

Dams and litters were sacrificed on PNDs 1, 3, 5, and 10. Blood and liver samples were collected 

for PFOA analysis. The fourth and fifth inguinal mammary glands were collected from dams and 

female offspring and analyzed as described above; whole mounts from lactating dams were 

qualitatively examined with respect to concurrent controls. 

Maternal weight gain in treated dams was significantly higher than control weight gain, but 

there were no effects of treatment on litter size or pup birth weight at PND 1. Significantly 

decreased body weight occurred in the pups of the 5U+L group on PND 3 and in all PFOAexposed pups on PNDs 5 and 10. Relative liver weight of the treated dams was significantly 

increased (p<0.05) compared to relative liver weight of control dams. On PND 1, liver-to-body 

weight ratios were significantly increased (p<0.05) in pups exposed in utero (5U, 5U+L); serum 

PFOA levels were 65,000–70,000 ng/mL. The liver-to-body weight ratio was increased in pups 

exposed lactationally by PND 5; serum PFOA levels were approximately 15,000 ng/mL (White 

et al. 2009). 

On PND 1, the mammary glands of PFOA-exposed dams were qualitatively similar to glands 

seen in late pregnancy, prior to parturition. In control dams nursing offspring from PFOAexposed dams, reduced alveolar filling was noted as early as PND 3, presumably a result of 

exposure of the dam from maternal grooming behavior. The delayed lactational morphology in 

dams treated with PFOA and control dams nursing offspring from PFOA-treated dams was 

persistent up to PND 10 (terminal sacrifice). Reduced mammary gland developmental scores

were observed as early as PND 1 in all female offspring from PFOA-exposed dams, including 

those exposed through lactation only (5L). Delayed mammary gland development persisted 

throughout the study duration (White et al. 2009). 

Macon et al. (2011) gavage-dosed CD-1 mice with 0, 0.3, 1.0, and 3.0 mg PFOA/kg/day 

from GD 1–17 (n = 13 dams per group). Six offspring per group were sacrificed on PNDs 7, 14, 

21, 28, 42, 63, and 84, and blood, liver, brain, and the fourth and fifth mammary glands were 

collected from female pups. Mammary gland developmental scores were not included in the 

published article, but were available in supplemental materials.

Body weight in male and female offspring was not affected through PND 84. Absolute liver 

weight was significantly increased at ≥0.3 mg/kg/day in females and at ≥1.0 mg/kg/day in males 

on PND 7, and at 3.0 mg/kg in females at PND 14. Relative liver weight was significantly 

increased at ≥0.3 mg/kg/day in males and females on PND 7, at ≥1.0 mg/kg/day in females on 

PND 14, and at 3.0 mg/kg/day in males and females on PNDs 14, 21, and 28. No dose-related 

differences were observed in absolute and relative brain weights.

Delayed mammary gland development of female pups was observed as early as PND 7 at 

≥1.0 mg/kg/day and PND 14 at ≥0.3 mg/kg/day and persisted until the end of the study. 

However, the developmental scores did not show dose-related trends at each interval. The 

delayed development was characterized by reduced epithelial growth and the presence of 

Perfluorooctanoic acid (PFOA) – May 2016 3-94
numerous terminal end buds. Photographs of the mammary gland whole mounts at PNDs 21 and 

84 show differences in the duct development and branching pattern of offspring from dams given 

0.3 and 1 mg/kg/day (offspring from high-dose dams not pictured). The LOAEL was 

0.30 mg/kg/day based on significantly increased liver weight and delayed mammary gland 

development. No NOAEL was established. The lowest dose tested was a NOAEL at PND day 7 

and is a LOAEL at day PND 14. 

Macon et al. (2011) also gavage-dosed CD-1 mice with 0, 0.01, 0.1, and 1.0 mg 

PFOA/kg/day from GD 10–17 (n = 5–8 dams per group) to examine the effects of low doses of 

PFOA on mammary gland development. Female offspring (one from at least three litters per 

group) were sacrificed on PNDs 1, 4, 7, 14, and 21, and blood, liver, and the fourth and fifth 

mammary glands were collected. In addition to the qualitative mammary gland developmental 

scores, quantitative measurements of longitudinal growth, lateral growth, and numbers of 

terminal end buds and terminal ends were recorded. These data were presented only for animals 

sacrificed on PND 21. 

No differences in body weight or brain weight were observed for male or female offspring. 

At 1 mg/kg, absolute and relative liver weights were significantly increased at PNDs 4 and 7.

Relative liver weight also was significantly increased at PND 14. Mammary gland development 

was delayed by exposure to PFOA, especially longitudinal epithelial growth. At PND 21, all 

treatment groups had significantly lower developmental scores. At the highest dose, poor 

longitudinal epithelial growth and decreased number of terminal end buds were observed. As 

seen in Table 3-22, the quantitative measures were statistically significant only for the high dose 

compared to the controls, while the qualitative scores were significantly different from controls 

at all doses. The LOAEL was 0.01 mg PFOA/kg/day based on the qualitative / quantitative 

developmental score for mammary gland development and 1 mg/kg/day based on the 

quantitative score in the absence of the qualitative component. No NOAEL was established.

Table 3-22. Mammary Gland Measurements at PND 21 from Female Offspring of

Dams Treated GD 10–17 

Dose

mg/kg/d Score

Longitude

μm

Lateral

μm

∆ Longitude

μm

∆ Lateral

μm

TEBs

#/gland

TEs

#/gland

0 n = 5 3.3 ± 0.3 4321 ± 306 5941 ± 280 3394 ± 306 4358 ± 280 40 ± 4 81 ± 12

0.01 n = 4 2.2 ± 0,2* 3803 ± 386 5420 ± 326 3087 ± 386 3899 ± 326 33 ± 4 61 ± 8

0.1 n = 3 1.8 ± 0.3** 3615 ± 320 4822 ± 672 2370 ± 320 3035 ± 672 24 ± 4 58 ± 4

1.0 n = 5 1.6 ± 0.1*** 2775 ± 285** 4822 ± 313** 1553 ± 301 3380 ± 313 15 ± 2*** 47 ± 11

Longitude = longitudinal epithelial growth

Lateral = Lateral epithelial growth

∆ =change in

*

= p<0.05, ** p<0.01, ***p<0.001

# = number

TEBs = terminal end buds

TE = differentiating duct ends

Source: Macon et al. 2011

White et al. (2011) examined the extended consequences of PFOA-induced altered mammary 

gland development in a multigenerational study in CD-1 mice. Pregnant mice (F0, n = 10–12 

dams per group) were gavage-dosed with 0, 1, and 5 mg PFOA/kg/day from GD 1–17. A 

separate group of pregnant mice (n = 7–10 dams per group) was gavage-dosed with either 0 or 

1 mg PFOA/kg/day from GD 1–17 and received drinking water containing 5 ppb PFOA 

beginning on GD 7. F1 females and F2 offspring from the second group continued to receive 

drinking water that contained 5 ppb PFOA until the end of the study, except during F1 breeding 

and early gestation, to simulate a chronic low-dose exposure. Only the F0 dams were given 

Perfluorooctanoic acid (PFOA) – May 2016 3-95
PFOA by gavage. Total doses were not calculated for the groups receiving drinking water with 

5 ppb PFOA. Table 3-23 shows the array of dosing regimens used in the study and the estimated 

average daily PFOA intake by F0 dams. The average daily intake from the chronic water 

exposures were calculated from total weekly water consumption, divided by the number of days 

per week (values given in supplemental materials; intake by the F1 animals was not calculated). 

Table 3-23. Dosing Regimens Used in the Multigeneration Study of CD-1 Mice

F0 Dams F0 Dams → F1 Offspring

F1 Dams → F2 

Offspring

F2 Offspring

Treatment Gavage Drinking water Gavage + 

drinking water

Drinking water Drinking 

water

Dose 0, 1, or 5 

mg/kg/day

0+5 ppb 1+5 ppb 5 ppb 5 ppb

Duration

Gavage

Drinking 

water

GD 1-17

None

GD 1-17

GD 7-LD 22

GD 1-17

GD 7-LD 22

None

Through LD 22

None

Through PND

63

Total Daily PFOA 

intake to dams 

from gavage and 

drinking water

Not relevant (0 

mg/kg/day)

36 μg/day (1 

mg/kg/day)

187 μg/day (5 

mg/kg/day

0.054 μg/day 

(gestation)

0.105 μg/day 

(lactation)

37 + 0.051 

μg/day 

(gestation)

0 + 0.130 μg/day 

(lactation)

Not calculated Not calculated

Source: White et al. 2011

F0 females were sacrificed on PND 22. F1 offspring were weaned on PND 22 and bred at 7–

8 weeks of age. F2 litters were maintained through PND 63. Groups of F1 and F2 offspring 

(n = 1–2 offspring per litter from 5–7 litters per group) were sacrificed on PND 22, 42, and 63. A 

group of F2 offspring (n = 6–10 per group) also was sacrificed on PND 10. A lactational 

challenge experiment was performed on PND 10 with F1 dams and F2 offspring. Mammary 

glands were evaluated from F0 dams on PND 22, from F1 dams on PNDs 10 and 22, and from 

F1 and F2 female offspring on PNDs 10 (F2 only), 22, 42, and 63. Mammary gland whole 

mounts were scored qualitatively as described above.

Exposure to 5 mg PFOA/kg/day significantly increased prenatal loss in F0 mice and 

significantly decreased the number of live offspring and the postnatal survival of the viable pups. 

Maternal weight gain and number of implants did not differ among F0 the groups. There was no 

indication of toxicity in F1 adult females. Exposure to PFOA did not affect prenatal loss or 

postnatal survival, although F1 females that had been exposed in utero to 5 mg/kg/day had 

significantly fewer implants.

On PND 22, F1 pup body weight was similar across all treated and control groups. F1 

offspring body weight at PND 42 was significantly reduced for those whose dams received 

5 mg/kg/day; at PND 63, body weight was significantly reduced for offspring from dams given 

1 mg/kg/day plus 5 ppb in the drinking water compared to offspring from dams given 

1 mg/kg/day. Liver-to-body weight ratios were significantly increased at 1 mg/kg/day on PND 

22 and at 5 mg/kg on PNDs 22 and 42. For the F2 pups, a significant reduction in body weight 

was observed in control plus 5 ppb drinking water PFOA offspring on PND 1, but there was no 

difference by PND 3. F2 offspring from the 1 mg/kg/day and 1 mg/kg/day plus 5-ppb drinking 

water PFOA groups had increased body weight compared to controls on PNDs 14, 17, and 22. 

Liver-to-body weight ratios were no different across the groups. 

Perfluorooctanoic acid (PFOA) – May 2016 3-96
Mammary gland developmental scores for the three generations of females are summarized 

in Table 3-24. At PND 22, control F0 dams displayed weaning-induced mammary involution. At 

PND 22, the mammary glands of all PFOA-exposed F0 dams, including the control dams 

receiving 5 ppb PFOA in drinking water, resembled glands of mice at or near the peak of 

lactation (~PND 10). The F1 dams examined on PNDs 10 and 22 had significantly lower 

developmental scores on PND 10, but that was no longer evident at PND 22, except for those 

exposed in utero to 5 mg/kg/day. 

Table 3-24. Mammary Gland Scores from Three Generations of CD-1 Female Mice

Group Control

Control 

+5 ppb 1 mg/kg/day

1 mg/kg/day + 

5 ppb 5 mg/kg/day

F0 dams on PND 22 2.4 ± 0.2 3.4 ± 0.1* 3.0 ± 0.2* 3.2 ± 0.2* 3.9 ± 0.1*

F1 as pups PND 63 3.8 ± 0.2 2.6 ± 0.4* 2.9 ± 0.2* 2.0 ± 0.3*# 2.2 ± 0.2*

F1 as dams on PND 10 4.0 ± 0.0 2.8 ± 0.5* 2.5 ± 0.2* 2.0 ± 0.2* 2.5 ± 0.2*

F1 as dams on PND 22 2.1 ± 0.3 2.2 ± 0.2 1.9 ± 0.4 1.5 ± 0.2* 3.2 ± 0.3*

F2 PND 10 2.8 ± 0.3 3.0 ± 0.2 1.9 ± 0.3 2.6 ± 0.2 2.0 ± 0.2

F2 PND 22 3.1 ± 0.4 1.9 ± 0.3 2.3 ± 0.1 2.3 ± 0.2 2.0 ± 0.2

F2 PND 42 3.5 ± 0.2 2.5 ± 0.4* 3.4 ± 0.2 2.4 ± 0.2*# 3.3 ± 0.4

F2 PND 63 3.4 ± 0.2 3.5 ± 0.2 2.4 ± 0.2* 2.6 ± 0.5 2.6 ± 0.4

Notes: n = 4–11. * p<0.05 compared with control.

# p<0.05 compared with 1 mg/kg/day.

F1 and F2 animals represented in each data set are different. They represent members of litters within each group at different 

stages of development.

In the F1 female offspring not used for breeding, the mammary glands of all mice exposed to 

PFOA were significantly delayed in development on PNDs 22, 42, and 63. For the F2 female 

offspring, some differences in mammary gland scores were observed between the groups, but 

most were not significantly different from controls.

In the lactational challenge experiment, dams were removed from their litters for 3 hours, 

then returned to their litters and allowed to nurse for 30 mins. The time from the dam’s return to 

the litter and nursing initiation was recorded. The litters were weighed before and after nursing to 

estimate volume of milk produced. The results from the lactational challenge on PND 10 for the 

F1 dams showed a slight dose-related trend for decreased milk production (measured in grams) 

over a 30-min period (differences from controls not identified as significant), but no clear 

differences in time to initiate nursing (measured in seconds). As discussed above, morphological 

differences were seen in developmental scores for the treated F1 dams on PND 10 and were 

generally no longer evident at PND 22. 

White et al. (2011) demonstrated that no significant dose-related differences were found in 

the ability of the CD-1 mice given 1 mg/kg/day to provide nourishment to their young as 

reflected in measurements of body weight in F1 and F2 pups across a 63-day postnatal period. 

There were body weight effects in the pups from dams given 5 mg/kg/day and in pups from 

dams that received 1 mg/kg/day by gavage with 5 ppb in the drinking water. 

In the study by Albrecht et al. (2013) discussed earlier, groups of female wild-type, PPARαnull, and PPARα-humanized mice on a SV/129 genetic background were given 0 and 3 mg 

PFOA/kg on GD 1–17 by oral gavage. Controls received the water vehicle. The study was 

Perfluorooctanoic acid (PFOA) – May 2016 3-97
designed with the goal of identifying the contribution of PPARα activation to the responses 

evaluated. Mammary gland structure was one of the endpoints evaluated. Females were either 

sacrificed on GD 18 (n = 5–8 per group) or allowed to give birth and then sacrificed, along with 

their litters (n = 8–14), on PND 20. The left and right fourth and fifth mammary glands were 

removed, spread on a glass slide, and stained. Ductal length and terminal end buds were 

quantified in the offspring of from three to nine dams. There was no significant difference in the 

measurements for either parameter at either timepoint for the offspring of PFOA-treated animals 

compared to the controls. In the case of the wild-type mice, the terminal end bud measurements 

were 2.1 ± 0.01 terminal end buds/gland for the control and 2.2 ± 0.2 terminal end buds/gland 

based on the mean for three control litters and four PFOA-exposed litters. For the ductal lengths, 

the values were 2.4 ± 0.3 millimeter (mm) for the control and 2.4 ± 0.4 mm for the PFOAexposed animals. There was no qualitative component of the scoring approach used by Albrecht 

et al. (2013). The fewest number of terminal end buds and the longest ductal length measurement 

were those for the animals with the hPPARα.

To examine the impact of differences in mouse strains, Tucker et al. (2015) conducted a 

study of the effects of gestational exposure on mammary gland development as measured at 

prepubertal time points. Doses of 0, 0.01, 0.1, 0.3, and 1 mg/kg/day were administered to timed 

pregnant CD-1 and C57Bl/6 mice by gavage on GD 1–17. After parturition, the number of pups 

was reduced so that there were ultimately four to eight CD-1 litters per treatment block and three 

to seven B57BL/6 litters per treatment. Endpoints monitored included body weight; net body 

weight; absolute and relative liver weight on PNDs 21, 35, and 56; neonatal developmental 

endpoints (e.g., vaginal opening, first estrus); and serum estradiol and progesterone (P) 

measurement; and as well as mammary gland development scores. Qualitative assessment of 

mammary gland scores was as described above. Different treatment blocks monitored different 

endpoints at different times. Serum POA levels were measured at PNDs 21, 35, and 56 for the 

CD-1 mice (n = 4–12) and at PND 21 and 61 for the C57BL/6 mice (n = 2–6). At each time 

point, the serum concentration increased with dose and decreased with duration. 

There were no measures that were significantly (p<0.05) different from controls for the CD-1 

anthropometric parameters, except relative liver weight on PND 56 at 0.3 mg/g/day and on PND 

21 at 1 mg/kg/day. Net body weight was significantly increased (p<0.05) at PNDs 21 and 35 in 

the 1-mg/kg/day group. No significant differences were observed in the C57Bl/6 mice at any 

dose or duration. There were no significant differences for postnatal developmental endpoints, 

estradiol, or P in either mouse strain. There was a trend towards decreasing mammary gland 

developmental scores with dose for both strains of mice. In the CD-1 mice, mammary gland 

developmental scores were significantly reduced at ≥0.01 mg/kg/day on PND 35 and at 

≥0.1 mg/kg/day on PND 21 compared to scores in the controls. However, in the C57Bl/6 mice, 

mammary gland developmental scores were significantly reduced only at 0.3 and 1.0 mg/kg/day 

on PND 21 compared to scores in the controls.

Serum P was higher in the control and treated CD-1 mice on PND 56 than at the other two 

time points but lacked dose response; estradiol was relatively consistent across time points. For 

the C57BL/6 mice, the estradiol levels at PND 61 were higher in all treated groups but lacked 

dose-response; P changed little with time and was similar between treated and control groups. 

The LOAEL was 0.01 mg/kg/day for aberrant mammary gland development in the CD-1 mice 

and 0.3 mg/kg/day for the C57BL/6 mice. The CD-1 mice lacked a NOAEL. The NOAEL for 

the C57/BL/6 mice was 0.1 mg/kg/day. Although both strains experienced delayed prepubertal 

mammary gland development, there were no significant changes in other postnatal 

Perfluorooctanoic acid (PFOA) – May 2016 3-98
developmental events. The relevance of the mammary gland changes at maturity in the absence 

of any postlactational PFOA exposure is uncertain, especially as it relates to humans. 

Direct peripubertal exposures

C. Yang et al. (2009) gavage-dosed 21-day-old female BALB/c mice (5 per group) with 0, 1, 

5, and 10 mg PFOA/kg/day for 5 days per week for 4 weeks to determine the effects of 

peripubertal PFOA exposure on puberty and mammary gland development. At necropsy, uteri 

and livers were weighed and processed for histological examination. Mammary glands were 

collected and processed for histological and whole-mount examination. A significant decrease in 

body weight was observed following exposure to 10 mg/kg/day. The mammary glands of female 

BALB/c mice treated with 5 or 10 mg/kg/day had reduced ductal length, decreased number of 

terminal end buds, and decreased stimulated terminal ducts compared to the mammary glands of 

control mice. BrdU incorporation into the mammary gland revealed a significantly lower number 

of proliferating cells in the ducts and terminal end buds/terminal ducts at 5 mg/kg/day (not tested 

at 10 mg/kg/day). Absolute and relative liver weight was significantly increased in all treated 

BALB/c mice. The absolute and relative uterine weight was significantly decreased in all treated 

mice compared to uterine weight in control mice. Vaginal opening was significantly delayed in 

mice dosed with 1 mg/kg/day and did not occur at 5 or 10 mg/kg/day. The LOAEL was 1 

mg/kg/day based on delayed vaginal opening, increased liver weight, and decreased uterine 

weight; and no NOAEL was established. 

C. Yang et al. (2009) also dosed 21-day-old female C57BL/6 mice in the same manner as the 

BALB/c mice and examined the effects of PFOA on mammary gland development and vaginal 

opening. The body weight effects were similar in both strains with 10 mg/kg/day causing 

significantly reduced body weight. At 5 mg/kg/day, PFOA had a stimulatory effect on the 

mammary glands. There was a significant increase in the number of terminal end buds and 

stimulated terminal ducts. Ductal length was not affected. Mammary gland development was 

inhibited in mice dosed with 10 mg/kg/day, with no terminal end buds or stimulated terminal 

ducts present and very little ductal growth. Absolute and relative liver weight was significantly 

increased in all treated mice. The absolute and relative uterine weight was significantly increased 

in C57BL/6 mice dosed with 1 mg/kg/day and significantly decreased in C57BL/6 mice dosed 

with 10 mg/kg/day. There was no difference in uterine weights between mice treated with 

5 mg/kg/day and control mice. Vaginal opening was delayed in C57BL/6 mice dosed with 

5 mg/kg/day and did not occur in mice dosed with 10 mg/kg/day. The LOAEL was 1 mg/kg/day 

based on increased liver and uterine weights, and no NOAEL was established. 

Y. Zhao et al. (2010) conducted several experiments in C57BL/6 mice to determine the 

potential mechanism by which peripubertal PFOA exposure resulted in the stimulation of 

mammary gland development observed by C. Yang et al. (2009). In experiments to determine if 

PFOA has a hormonal effect on mammary gland development, C57BL/6 mice (n = 10 per group) 

were OVX at 3 weeks of age, allowed 1 week to recover, and treated with 0 and 5 mg PFOA/kg 

bw/day for 4 weeks. Abdominal and inguinal mammary glands were collected at sacrifice, 

prepared as whole mounts, and scored for growth and development. The mammary glands of the 

OVX control and PFOA-treated OVX mice were similarly stunted in growth as evidenced by no 

outgrowth of ducts or presence of terminal end buds. This was in contrast to the stimulatory 

effect of PFOA observed by C. Yang et al. (2009) in intact mice.

In experiments to determine if PFOA-affected mammary glands respond to hormone 

treatment, intact C57BL/6 mice were dosed with 0 or 5 mg/kg bw/day of PFOA for 4 weeks 

Perfluorooctanoic acid (PFOA) – May 2016 3-99
starting at 21 days of age. After the last dose, the mice were OVX, allowed to recover for 1 

week, and injected subcutaneous for 5 days with E2 (1 μg/0.2 ml per mouse), P (1 mg/0.2 ml per 

mouse), or both (E+P, 1 μg+1 mg/0.2 ml per mouse). The mice were sacrificed 24 hours after the 

last hormone injection. Abdominal and inguinal mammary glands were collected at sacrifice, 

prepared as whole mounts, and scored for growth and development. In the mammary glands of 

mice treated with PFOA and estradiol, stimulated terminal ducts were observed, and in PFOAtreated mice given P or E+P, stimulated terminal ducts and an increased number of side branches 

were observed. The results showed that PFOA increased the mouse mammary gland response to 

exogenous estrogen and P. 

In experiments to determine if PFOA-induced mammary gland development stimulation is 

related to PPARα expression and the impact of PFOA on steroid hormones and growth factors, 

female C57BL/6 and PPARα-null C57BL/6 mice (n = 5–10 mice per group) were gavage-dosed 

with 0 or 5 mg/kg bw/day of PFOA 5 days per weeks for 4 weeks starting at 21 days of age (Y. 

Zhao et al. 2010). Vaginal opening was monitored daily and estrous cycle state was determined 

at sacrifice after 4 weeks of treatment. At necropsy, blood was collected for measurement of 

serum steroid hormones and binding proteins. Portions of the mammary glands, ovaries, and 

livers were collected and processed for histological examination. RNA was extracted from the 

livers for quantitative RT-PCR and PCR array for selected genes related to metabolism of drugs, 

toxic chemicals, hormones, and micronutrients. Portions of mammary glands were used in 

western blot analysis of several enzymes, local growth factors, and receptors, including 

aromatase—which aids in converting testosterone to estradiol and androstenedione to estrone, 

hydroxysteroid 17β dehydrogenase 1 (HSD17β1)—which aids in converting estrone to estradiol, 

and hydroxysteroid 3β dehydrogenase 1 (HSD3β1)—which aids in converting pregnenolone to P 

and androstenedione to testosterone. Growth factors critically involved in mammary gland 

development, including amphiregulin (Areg), insulin like growth factor I (IGF-I), and hepatocyte 

growth factor (HGFα), and markers of cell proliferation (e.g., cyclin D1 and PCNA) were 

analyzed by western blot. Areg mediates estrogen receptor α (ERα) function and is a ligand for 

the epidermal growth factor receptor (EGFR). These receptors also were analyzed by western 

blot. 

The mammary glands of PPARα-null mice treated with PFOA had an increased number of 

terminal end buds and stimulated terminal ducts compared to control PPARα-null mice. Protein 

levels of Areg, IGF-I, HGFα, ERα, and EGFR were significantly increased (p<0.05) in PFOAtreated C57BL/6 mice; and Areg, HGFα, ERα, and EGFR were significantly increased (p<0.05) 

in PFOA-treated PPARα-null mice. Cyclin D1 and PCNA were significantly increased (p<0.05) 

in C57BL/6 and PPARα-null mice treated with PFOA compared to levels in control mice.

Immunofluorescent staining of the mammary glands for ERα and Areg showed a significant 

increase (p<0.05) in Areg positive luminal epithelial cells and Areg and ERα double positive 

staining cells in C57BL/6 and PPARα-null mice treated with PFOA compared to control mice.

The results show that the stimulatory effect of PFOA on mammary gland development is 

independent of PPARα expression and suggest that PFOA increases the levels of steroid 

hormones, growth factors, and receptors, which promote mammary gland cell proliferation. 

Estradiol levels were similar between intact control and treated wild-type mice, but P levels 

were significantly increased (p<0.05) in PFOA-treated mice in proestrus and estrus compared to 

control mice in the same stages of the estrous cycle. Serum SHBG and albumin levels were not 

significantly changed by treatment with PFOA. 

Perfluorooctanoic acid (PFOA) – May 2016 3-100
The effect of PFOA on aromatase, HSD17β1, and HSD3β1 activity in the ovaries of 

C57BL/6 PPARα-null mice was examined. In C57BL/6 mice, HSD17β1 and HSD3β1 proteins 

were significantly increased (p<0.05), and in PPARα-null mice, HSD17β1 protein was 

significantly increased. Aromatase levels were not affected by PFOA. The results suggest that 

PFOA might increase serum steroid hormone levels in the ovaries.

Due to the increased P levels observed in PFOA-treated mice, the expression of liver 

metabolic enzymes was analyzed. Liver metabolic function might affect steroid hormone serum 

levels, which play a role in mammary gland development. In PPARα-null and C57BL/6 mice 

treated with PFOA, detoxification enzymes in the liver, including glutathione s-transferase α1, 

μ3, and μ4, were upregulated (p<0.05). Expression of liver drug metabolic enzymes, including 

CYP1a1, CYP1a2, and HSD17β2, was significantly downregulated (p<0.05) in C57BL/6 mice 

treated with PFOA, but expression in PFOA-treated PPARα-null mice was comparable to that in 

control mice. Hydroxysteroid 17β dehydrogenase 4, an enzyme that converts estradiol to estrone, 

was significantly upregulated (p<0.05) in C57BL/6 mice treated with PFOA. The results 

suggested that PFOA-induced expression changes in liver enzymes might not contribute to 

PFOA-induced mammary gland development stimulation. 

Inhalation Exposure 

Staples et al. (1984) exposed Sprague-Dawley rats to PFOA using whole-body dust 

inhalation for 6 hours per day on GD 6–15. The MMAD of the particles ranged from 1.4 to 

3.4 μm and the GSD ranged from 4.3 to 6.0. The study was carried out in two trials with each 

trial including two experiments. In experiment 1, the dams were sacrificed on GD 21 prior to 

parturition, and in experiment 2, the dams were allowed to litter and were sacrificed on PND 23; 

offspring were sacrificed on PND 35. In the trial 1 (both experiments), dams (n = 12) were 

exposed to 0, 0.1, 1, and 25 mg/m3

. In trial 2, the high dose was reduced to 10 mg/m3. In 

experiment 1 of trial 2, dams numbered 12–15 per group and two additional groups (6 dams per 

group) were added and were pair-fed at 10 and 25 mg/m3. In experiment 2 of trial 2, only six 

control and six dams dosed at 10 mg/m3 were allowed to litter.

In experiment 1, the dams were weighed on GDs 1, 6, 9, 13, 16, and 21 and observed daily 

for abnormal clinical signs. On GD 21, the dams were sacrificed by cervical dislocation and 

examined for any gross abnormalities, liver weights were recorded, and the reproductive status 

of each animal was evaluated. The ovaries, uterus, and contents were examined for the number 

of corpora lutea, live and dead fetuses, resorptions, and implantation sites. Pups (live and dead) 

were counted, weighed and sexed, and examined for external, visceral, and skeletal alterations. 

The heads of all control and high-dosed group fetuses were examined for visceral alterations and 

macroscopic and microscopic evaluations were conducted of the eyes.

Treatment-related clinical signs of maternal toxicity occurred at 10 and 25 mg/m3 and 

consisted of wet abdomens, chromodacryorrhea, chromorhinorrhea, a general unkempt 

appearance, and lethargy in four dams at the end of the exposure period (high-concentration 

group only). Three out of 12 dams died during treatment at 25 mg/m3 (on GDs 12, 13, and 17). 

Food consumption was significantly reduced at 10 and 25 mg/m3; however, no significant 

differences were noted between treated and pair-fed groups. Significant reductions in body 

weight also were observed at these concentrations, with statistical significance at the high

concentration only. Likewise, statistically significant increases in mean liver weights (p<0.05) 

were seen in the high-concentration group. The NOAEL and LOAEL for maternal toxicity were 

1 and 10 mg/m3

, respectively.

Perfluorooctanoic acid (PFOA) – May 2016 3-101
No effects were observed on the maintenance of pregnancy or the incidence of resorptions. 

Mean fetal body weights were significantly decreased in the 25 mg/m3

 PFOA group (p = 0.002) 

and in the pair-fed control group (p = 0.001). Interpretation of the decreased fetal body weight is 

difficult given the high incidence of mortality in the dams. The NOAEL and LOAEL for 

developmental toxicity were 10 and 25 mg/m3

, respectively.

In experiment 2, in which the dams were allowed to litter, the procedure was the same as for 

experiment 1 until GD 21. Two days before the expected day of parturition, each dam was 

housed in an individual cage. The date of parturition was noted and designated PND 1. Dams 

were weighed and examined for clinical signs on PNDs 1, 7, 14, and 22. On PND 23 all dams 

were sacrificed. Pups were counted, weighed, and examined for external alterations. At birth, 

each pup was subsequently weighed and then inspected for adverse clinical signs on PNDs 4, 7, 

14, and 22. The eyes of the pups were also examined on PNDs 15 and 17. Pups were sacrificed 

on PND 35 and examined for visceral and skeletal alterations. 

Clinical signs of maternal toxicity seen at 10 and 25 mg/m3 were similar in type and 

incidence to those described for trial 1. Maternal body weight gain during treatment at 25 mg/m3

was less than controls, although the difference was not statistically significant. In addition, two 

out of 12 dams died during treatment at 25 mg/m3

. No other treatment-related effects were 

reported, nor were any adverse effects noted for any of the measurements of reproductive 

performance. The NOAEL and LOAEL for maternal toxicity were 1 and 10 mg/m3, respectively.

Signs of developmental toxicity in this group consisted of statistically significant reductions 

in pup body weight on PND 1 (6.1 g at 25 mg/m3

 versus 6.8 g in controls, p = 0.02). On PNDs 4 

and 22, pup body weight continued to remain lower than controls, although the difference was 

not statistically significant. No significant effects were reported following external examination 

of the pups or with ophthalmoscopic examination of the eyes. The NOAEL and LOAEL for 

developmental toxicity were 10 and 25 mg/m3

, respectively.

Dermal Exposure

No data on the developmental effects of dermal exposures to PFOA were identified in the 

literature.

3.2.8 Chronic Toxicity 

Oral Exposure 

Monkey. Male cynomolgus monkeys (n = 4 or 6 per dose) were administered PFOA by oral 

capsule containing 0, 3, 10, or 30/20 mg/kg/day for 26 weeks (Butenhoff et al. 2002). Dosing of 

animals in the 30-mg/kg/day dose group ceased after 12 days and decreased to 20 mg/kg/day 

when reinstated on day 22 because of low food consumption, decreased body weight, and 

decreased feces. Sacrifice of the surviving monkeys, except for two control monkeys and two 

monkeys from the mid-dose group (recovery animals) occurred at 26 weeks. The animals in the 

recovery groups were sacrificed 13 weeks later.

Animals were observed twice daily for mortality and moribundity and were examined at least 

once daily for signs of poor health or abnormal behavior. Ophthalmic examinations were 

performed before treatment began and at weeks 26 and 40. Body weight, food consumption, 

clinical hematology, clinical chemistry, urinalysis, serum hormone levels, and PFOA levels in 

Perfluorooctanoic acid (PFOA) – May 2016 3-102
blood and tissue were assessed throughout the study. One animal from the 30/20-mg/kg/day dose 

group was sacrificed in moribund condition on day 29 with signs of dosing injury and liver 

lesions. One animal from the 3-mg/kg/day dose was sacrificed (day 137) with signs of hind limb 

paralysis, ataxia and hypoactive behavior, few feces, and no food consumption. Treatment of the 

remaining three animals given 30/20 mg/kg/day was halted on days 43, 66, and 81, respectively, 

because of thin appearance, few or no feces, low or no food consumption, and weight loss, but 

the animals appeared to recover from compound-related effects within 3 weeks after cessation 

of treatment. No significant changes in mean body weight were observed at doses of 3 or 

10 mg/kg/day. 

Serum hormone levels (i.e., estrone, estradiol, estriol, testosterone, TSH, FT4, total T4, and 

CCK) were not significantly altered throughout the study. However, FT3 and total T3 levels 

were significantly decreased (p<0.05) from weeks 5 to 10 and at week 27 in the 30/20-mg/kg/day 

dose group compared to controls.

At terminal sacrifice (26 weeks), mean absolute liver weight was significantly increased in 

all dose groups and the relative liver-to-body weight ratio was significantly increased for the 

High-Dose Group. Final Body Weight And Liver Weight Data Are Presented In Table 3-25. 

Table 3-25. Liver Weight Data in Monkeys Administered PFOA for 6 Months

Dose Body Weight Absolute Liver Wt (g) Relative Liver Wt (%)

0 mg/kg (n = 4) 3947 ± 591 60.2 ± 6.9 1.5 ± 0.1

3 mg/kg (n = 3) 4486 ± 30 81.8 ± 2.8* 1.8 ± 0.1

10 mg/kg (n = 4) 4447 ± 498 83.2 ± 9.7* 1.9 ± 0.1

30/20 mg/kg (n = 2) 3925 ± 583 90.4 ± 4.2* 2.4 ± 0.5*

Source: Butenhoff et al. 2002

Note: * Significantly different from control, p<0.01.

The cause of the increase in liver weight was suggested to be hepatocellular hypertrophy 

(indicated by decreased hepatic DNA content), which was hypothesized to result from 

mitochondrial proliferation based on an increase in hepatic succinate dehydrogenase activity.

The two animals given 20 mg/kg/day had significantly decreased hepatic DNA content, and 

increased succinate dehydrogenase and palmitoyl-CoA oxidase activities; glucose-6-phophatase 

activity was slightly decreased in all treated groups, but a dose-response was not shown. These 

data are shown in Table 3-26. Succinate dehydrogenase activity was highly variable in animals 

given 3 mg/kg/day despite this group having the most consistent liver PFOA concentrations. 

Although serum steady-state had been attained by 4-6 weeks of dosing (Table 3-26 (see section 

2.2, Distribution), liver PFOA levels ranged from 11.3-18.5, 6.29-21.9, and 16-83.3 μg/g tissue 

in the 3, 10, and 20 mg/kg/day groups, respectively. 

Because administration of PFOA to rats has been shown to result in liver, Leydig cell tumors 

(LCTs), and pancreatic acinar cell tumors (PACTs), Butenhoff et al. (2002) analyzed markers of 

tumor formation in the monkey study just described. In the liver, a twofold increase in hepatic 

palmitoyl-CoA oxidase activity was observed in the 30/20-mg/kg/day group, consistent with 

reports for species that are not particularly responsive to PPARα agonists. Replicative DNA 

synthesis in the liver, an indication of cell proliferation, was not altered in the treated animals. It 

also has been proposed that changes associated with the PACTs in rats include increased serum 

CCK concentrations and indications of cholestasis, including increases in ALP, bilirubin, and 

bile acids. None of these changes were observed in the cynomolgus monkeys. There were also 

no significant changes in estradiol, estriol, or testosterone in the monkeys. Each of these factors 

Perfluorooctanoic acid (PFOA) – May 2016 3-103
is associated with LCTs in rats. There were no changes in replicative DNA synthesis in the 

pancreas or testes.

Table 3-26. Subcellular Liver Enzyme Activities and Liver PFOA Concentrations

Endpoint 0 mg/kg/day 3 mg/kg/day 10 mg/kg/day 20 mg/kg/day

DNA (mg/g liver) 1.44 ± 0.28 1.23 ± 0.89 1.25 ± 0.37 1.02 ± 0.17*

Succinate dehydrogenase (μmol 

cytochrome c reduced/min/g liver)

0.21 ± 0.15 1.77 ± 1.59 0.55 ± 0.14 1.37 ± 0.73*

Palmitoyl-CoA oxidation 

(μmol/min/g liver)

0.53 ± 0.12 0.47 ± 0.13 0.90 ± 0.29 1.36 ± 0.34*

Acid phosphatase (μmol/min/g 

liver)

0.78 ± 0.10 0.81 ± 0.11 0.80 ± 0.14 0.55 ± 0.10

Glucose-6-phophatase (μmol/min/g 

liver)

12.32 ± 3.11 6.02 ± 0.33* 10.17 ± 0.63 8.83 ± 1.41

PFOA liver level (μg/g tissue)

(individual animal)

0.09

<LOQ

0.23

<LOQ

15.2

18.5

11.3

18.3 (sacrificed 

day 137)

21.9

6.29

8.86

18.8

16.0

83.3

Source: Butenhoff et al. 2002

Notes: 

* Significantly different from control, p<0.05.

<LOQ: less than the lower LOQ of 0.019 μg/g.

After a 13-week recovery period, there were no treatment-related effects on terminal body 

weights or on absolute or relative organ weights, suggesting that the treatment-related liver 

weight changes were reversible. There were no treatment-related macroscopic or microscopic 

changes at the recovery sacrifice. Toxicological significance of the parameters in Table 3-26 are 

difficult to evaluate because of the small number of animals and the high variability in the 

enzyme measurements and the liver PFOA levels.

Rat. The chronic toxicity of PFOA was investigated in a 2-year study in rats by Butenhoff et al. 

(2012); this study was conducted from April 1981 through May 1983. Sprague-Dawley (Crl:CD 

BR) rats (50 per gender) were fed diets containing 0, 30, and 300 ppm PFOA (0, 1.3, and 14.2 

mg/kg/day for males; 0, 1.6, and 16.1 mg/kg/day for females). Groups of 15 additional rats per 

gender were fed 0 or 300 ppm PFOA and evaluated at the 1-year interim sacrifice. All animals 

were observed daily throughout the 2-year dosing period. Periodic observations included body 

weights and feed consumption, hematology, serum chemistry, urinalysis, gross pathology, organ 

weights, and histopathology. Animals were sacrificed after 1 and 2 years of dosing. Organ 

weights were determined after each sacrifice and tissues subjected to histological examination.

No dose- or treatment-related clinical signs of toxicity were observed in males or females.

Significantly decreased body weight gains occurred in high-dose male and female rats compared 

to the controls. The body weight changes correlated with slight decreases in feed consumption 

during the study. Survival rate was increased for high-dose males and females compared with 

their respective controls, likely because of the lower body weights for these treated groups. No 

consistent dose-related changes over time were observed in hematology parameters of males and 

females. Clinical chemistry changes included slight (less than twofold) but significant increases 

in ALT, AST, and ALP in both treated male groups from 3 to 18 months, but only in the highdose males at 24 months (Table 3-27); no differences were observed in females. No dose- or 

treatment-related differences in absolute and relative organ weights were found between the 

treated and control groups at 2 years. 

Perfluorooctanoic acid (PFOA) – May 2016 3-104
Table 3-27. Clinical Chemistry Values from Male Rats Given PFOA for 2 Years 

Endpoint

Diet Level 

(ppm) 3 Months 6 Months 12 Months 18 Months 24 Months

ALT (IU/L) 0 21.4 ± 2.67 24.1 ± 3.75 33.5 ± 19.45 34.1 ± 10.68 33.4 ± 8.1

30 34.5 ± 15.33* 53.3 ± 29.34* 77.6 ± 56.59* 59.7 ± 33.41* 42.5 ± 10

300 31.9 ± 21.94* 54.8 ± 29.26* 106.1 ± 70* 84.3 ± 55.95* 61.8 ± 20.13*

AST (IU/L) 0 45.3 ± 7.26 49.7 ± 14.98 79.1 ± 44.61 99.1 ± 68.14 64.9 ± 25.76

30 59.7 ± 22.47 92.1 ± 45.6* 124.4 ± 94.04* 116.4 ± 57.99 68.0 ± 17.64

300 58.2 ± 27.23 87.8 ± 34.83* 132.7 ± 76.84* 123.3 ± 62.98 95.7 ± 29.76*

ALP (mg/dL) 0 91.1 ± 26.22 97.1 ± 40.41 150.8 ± 43.94 85.2 ± 33.76 70.1 ± 25.53

30 138.7 ± 33.14* 146.9 ± 37.13* 128.3 ± 41.75 112.5 ± 32.61 81.2 ± 26.2

300 153.5 ± 31.84* 147.3 ± 34.85* 166.5 ± 59.28* 184.4 ± 73.37* 113.5 ± 22.84*

Source: Butenhoff et al. 2012

Note:

* Significantly different from control, p ≤ 0.05. 

Incidence of selected microscopic lesions is detailed in Table 3-28; severity scores were not 

given for any type of lesion. Significantly increased incidence of lesions in the liver was

observed in the high-dose male group. At 1 year, diffuse hepatocellular hypertrophy, portal 

mononuclear cell infiltration, and hepatocellular necrosis were seen. At 2 years, significant 

increases in hepatocellular hypertrophy were observed in the males and females in the high-dose 

group. Hepatic cystoid degeneration, a condition characterized by areas of multilocular 

microcysts in the liver parenchyma, also was significantly increased in high-dose males. The 

incidence of hepatocellular necrosis did not increase for the high-dose males at the end of the 

study compared with the interim rate.

Among the high-dose males, histological changes were noted in tissues other than the liver. 

Small but statistically significant increases in vascular mineralization of the testes and of 

pulmonary hemorrhage probably were not caused by treatment with PFOA. In the lung, while the 

incidence of alveolar macrophages was increased, that of perivascular mononuclear infiltrate and 

of pneumonia were decreased and vascular mineralization was a common finding in treated and 

control animals.

The LOAEL for male rats is 300 ppm (14.2 mg/kg/day) based on a decrease in body weight 

gain and histological changes in the liver. The LOAEL for female rats is 300 ppm 

(16.1 mg/kg/day) based on decreased body weight gain. The NOAEL for both genders is 30 ppm 

(1.3 mg/kg/day for males and 1.6 mg/kg/kg for females). 

Biegel et al. (2001) conducted a 2-year mechanistic study in which male Crl:CD BR (CD) 

rats (n = 156 per group) were fed a diet containing 0 or 300 ppm PFOA (0 or 13.6 mg/kg/day). 

Interim sacrifices were conducted every 3 months up to 21 months for measurements of liver and 

testes weights, peroxisome proliferation, and cell replication. Serum samples were collected and 

reproductive hormones measured. 

Perfluorooctanoic acid (PFOA) – May 2016 3-105
Table 3-28. Incidence of Nonneoplastic Lesions in Rats Given PFOA for 2 Years

Lesion 0 ppm 30 ppm 300 ppm

Males

Liver

Cystoid degeneration

Hepatocellular hypertrophy

[incidence at 1 year]

Mononuclear cell infiltrate

[incidence at 1 year]

Necrosis

[incidence at 1 year]

4/50

0/50

[0/15]

37/50

[7/15]

3/50

[0/15]

7/50

6/50

[-] 

32/50

[-] 

5/50

[-]

28/50*

40/50*

[12/15]

48/50*

[13/15]

5/50

[6/15]

Lung

Alveolar macrophages

Hemorrhage

Mononuclear infiltrate

10/49

10/49

21/49

16/50

14/49

3/49*

31/49*

22/50*

7/50*

Testes

Vascular mineralization 0/49 3/50 9/50*

Females

Liver

Cystoid degeneration

Hepatocellular hypertrophy

Mononuclear cell infiltrate

Necrosis

0/50

0/50

19/50

5/50

0/50

1/50

11/50

6/50

0/50

8/50*

19/50

2/50

Source: Butenhoff et al. 2012

Notes: * Significantly different from control, p≤0.05.

- Not examined; interim sacrifice not done on animals at 30 ppm.

Body weight was significantly decreased from days 8 through 630 in PFOA-exposed rats. In 

the treated group, relative liver weights and hepatic β-oxidation activity were statistically 

significantly increased at all time points between 1 and 21 months when compared to the 

controls. Absolute testis weights were significantly increased only at 24 months. No hepatic or 

Leydig cell proliferation was observed at any sampling times. The incidence of Leydig cell 

hyperplasia was significantly increased in PFOA-exposed rats (46% versus 14% in the control 

group). Pancreatic acinar cell proliferation was significantly increased at 15, 18, and 21 months. 

The incidence of acinar cell hyperplasia was 30/76 (39%) compared to the incidence in the 

control group of 14/80 (18%). There were no significant differences in serum testosterone or 

prolactin in the PFOA-treated rats when compared to the controls. Serum FSH was significantly 

increased at 6 months, and LH was significantly increased at 6 and 18 months. There were 

significant increases in serum estradiol concentrations in the treated rats at 1, 3, 6, 9, and 12 

months. 

3.2.9 Carcinogenicity

Oral Exposure 

Rat. Tissues from the animals in the Butenhoff et al. study (2012) were evaluated for neoplastic 

and preneoplastic formations; this study was conducted from April 1981 through May 1983. 

Hepatocellular carcinomas were observed at 6% (3/49), 2% (1/50), and 10% (5/50) in the 

control, low-, and high-dose male rats, respectively. None were observed in females in the 

control and low-dose groups, but a 2% (1/50) incidence was observed for female rats in the highPerfluorooctanoic acid (PFOA) – May 2016 3-106
dose group. The differences between the treated and control groups were not significantly 

different. No liver adenomas were observed.

At the 1-year sacrifice, testicular masses were found in 7/50 (14%) high-dose and 2/50 (4%) 

low-dose rats, but not in any of the controls. A significant increase (p<0.05) in the incidence of 

testicular (Leydig) cell adenomas was observed in the high-dose male rats at the end of the study.

The LCT incidence in the control, low-, and high-dose groups was 0/50 (0%), 2/50 (4%), and 

7/50 (14%), respectively. The increase also was statistically significant when compared to the 

historical control incidence of 0.82% observed in 1,340 Sprague-Dawley control male rats used 

in 17 carcinogenicity studies (Chandra et al. 1992). In a published workshop report on LCTs, 

Clegg et al. (1997) identified the spontaneous incidence of LCTs in 2-year-old Sprague-Dawley

rats as approximately 5%. 

A statistically significant, dose-related increase in the incidence of ovarian tubular 

hyperplasia was found in female rats at the 2-year sacrifice. The incidence of this lesion in the 

control, low-, and high-dose groups was 0%, 14%, and 32%, respectively. The biological 

significance of this effect at the time of the initial evaluation was unknown, as there was no 

evidence of progression to tumors.

Slides of the ovaries from the Butenhoff et al. study (2012)—originally conducted from April 

1981 through May 1983—were reevaluated by Mann and Frame (2004) with emphasis placed on 

the proliferative lesions of the ovary. Using more recently published nomenclature, the ovarian 

lesions were diagnosed and graded as gonadal stromal hyperplasia and/or adenomas, which 

corresponded to the diagnoses of tubular hyperplasia or tubular adenoma by the original study 

pathologist. The data are summarized in Table 3-29. No statistically significant increases in 

hyperplasia (total number), adenomas, or hyperplasia/adenoma combined were seen in treated 

groups compared to controls. There was some evidence of an increase in size of stromal lesions 

observed at the 300-ppm group; however, adenomas occurred in greater incidence in the control 

group than in either of the treated groups. Results of this follow-up evaluation indicated that rats 

sacrificed at the 1-year interim sacrifice, as well as rats that died prior to the interim sacrifice, 

were not considered at risk for tumor development. 

Table 3-29. Incidence of Ovarian Stromal Hyperplasia and Adenoma in Rats 

Group 0 ppm 30 ppm 300 ppm

No. examined 45 47 46

Hyperplasia (Total) 8 16 15

Grade 1 6 7 5

Grade 2 2 3 1

Grade 3 0 5 6

Grade 4 0 1 3

Adenoma 4 0 2

Adenoma and/or Hyperplasia 12 16 17

Source: Mann and Frame 2004

Mammary gland tumors also were observed in the Butenhoff et al. (2012) bioassay. In the 

original analysis of mammary tissues from female rats, the incidence of fibroadenoma of the 

mammary gland in the female 300-ppm group (48%) was greater than that in either of the 

concurrent control groups (22%). It also was similar to the incidence in the 30-ppm group (42%), 

but considered to be within the norm for background variation of this lesion in Sprague-Dawley 

Perfluorooctanoic acid (PFOA) – May 2016 3-107
rats based on the published literature. As a result of questions raised about this conclusion, a 

pathology working group (PWG) was commissioned to review the female mammary tumor 

findings, blinded to treatment status, using current diagnostic criteria (Hardisty et al. 2010). 

Table 3-30 compares the original mammary gland tumor findings to those of the PWG. 

Table 3-30. Mammary Gland Tumor Incidence Comparison 

0 ppm 30 ppm 300 ppm

Butenhoff Hardisty Butenhoff Hardisty Butenhoff Hardisty

Number necropsied 50 50 50 50 50 50

Lobular hyperplasia (%) 6

(12%)

0

(0%)

3

(6%)

2

(4%)

2

(4%)

0

(0%)

Adenocarcinoma (%) 8

(16%)

9

(18%)

14

(28%)

16

(32%)

5

(10%)

5

(10%)

Fibroadenomaa (%) 10

(20%)

18

(36%)

19

(38%)

22

(44%)

21

(42%)

23

(46%)

Adenoma (%) 3

(6%)

1

(2%)

0

(0%)

0

(0%)

0

(0%)

0

(0%)

Source: Hardisty 2005; Hardisty et al. 2010 

Notes: 

a

 Includes fibroadenoma, multiple counts. 

The principal differences between the original reported findings and the PWG results relate 

to changes in the mammary gland that were initially reported as lobular hyperplasia, which the 

PWG felt had features more characteristic of mammary gland fibroadenoma (Table 3-30). As a 

result, the numbers of rats with benign tumors (adenoma and fibroadenoma) were reclassified 

from 13 to 19 in the control group, from 19 to 22 in the 30-ppm group, and from 21 to 23 in the 

300-ppm group. Although the incidence of neoplasms varied among the control and treated 

groups, there were no statistically significant differences when evaluated using the Fisher’s exact 

test for pairwise comparison for fibroadenoma, adenocarcinoma, total benign neoplasms, and 

total malignant neoplasms. The morphologic appearance, overall incidence, and distribution of 

the neoplasms observed in treated and control groups were similar, resulting in a conclusion that 

they are not related to compound administration. 

A 2-year mechanistic study in male Crl:CD BR (CD) rats (Biegel et al. 2001; Cook et al. 

1992) resulted in liver tumors, LCTs, and PACTs. The rats (n = 156 per group) were fed diets 

containing 0 ppm (ad libitum control and control pair-fed to the PFOA-exposed rats) or 300-ppm 

PFOA (13.6 mg/kg intake). Rats were euthanized at interim time points of 1, 3, 6, 9, 12, 15, 18, 

and 21 months. All rats surviving the 24-month test period were necropsied for microscopic 

examination of various organs (e.g., kidneys, liver, testes, brain, heart, spleen). The incidence of 

liver adenomas in the ad libitum control, pair-fed control, and treated groups was 3% (2/80), 1% 

(1/79), and 13% (10/76), respectively. In the Butenhoff et al. study (2012), no hepatic adenomas 

were observed. The incidence for liver carcinomas was 0% (0/80), 3% (2/79), and 0% (0/76) in 

the ad libitum control, pair-fed control, and treated groups, respectively. 

There was a significant increase in the incidence of Leydig cell adenomas in the treated 

rats—11% (8/76) when compared to the pair-fed control rats (3%, 2/78)—supporting the 

observations from the Butenhoff et al. study (2012). The incidence in ad libitum control rats was 

0% (0/80). In addition, the treated group had a significant increase in the incidence of liver 

adenomas and pancreatic acinar cell adenomas when compared to the pair-fed and ad libitum 

control groups. The incidence for the pancreatic acinar cell adenomas was 0% (0/80) in the 

Perfluorooctanoic acid (PFOA) – May 2016 3-108
treated rats, 1% (1/79) in the pair-fed control rats, and 9% (7/76) in the control rats. The 

incidence of pancreatic acinar cell carcinoma was 1% (1/76) in the treated rats, 0% (0/79) in the 

pair-fed control rats, and 0% (0/80) in the control rats. 

In Butenhoff et al. (2012), there was no reported increase in the incidence of PACTs. 

However, the incidence of pancreatic acinar hyperplasia in the male rats was 0/33, 2/34, and 1/43 

in the control, 30-, and 300-ppm groups, respectively. To resolve this discrepancy, the 

histological slides from both studies were reviewed by independent pathologists. This review of 

the microscopic lesions of the pancreas in the two studies indicated that PFOA produced 

increased incidence of proliferative acinar cell lesions of the pancreas in the rats of both studies 

at the dietary concentration of 300 ppm. The differences observed were quantitative rather than 

qualitative; more and larger focal proliferative acinar cell lesions and greater tendency for 

progression of lesions to adenoma of the pancreas were observed in the Biegel et al. study (2001) 

than in the Butenhoff et al. study (2012). The difference between pancreatic acinar hyperplasia 

(Butenhoff et al. 2012) and adenomas (Biegel et al. 2001) in the rat was a reflection of arbitrary 

diagnostic criteria and nomenclature by different pathologists. The basis for the quantitative 

difference in the lesions observed is not known, but was believed most likely to have been 

caused by the difference in the diets used in the two laboratories (Frame and McConnell 2003).

Mouse. Filgo et al. (2015) reported on tumor development in females from three strains of mice 

(CD-1, SV-139, and SV-129 PPARα knock-out [KO]) at 18 months with exposures that occurred 

only during development (gestation and lactation). The animals were from separate experiments 

initially carried out by EPA and published as Hines et al. (2009) and Abbott et al. (2007). The 

Filgo et al. (2015) analysis focused on the mature offspring from the earlier publications and was 

carried out at the National Institute for Environmental Health Sciences (NIEHS). Dosing 

regimens differed for the individual strains as did the doses and the number of animals per dose 

group. Some of the animals in the original studies had died before the 18-month sacrifice at 

NIEHS. After sacrifice, the livers were recovered for analysis. The tissue sections were reviewed 

by a team of board-certified veterinary pathologists. Table 3-31 summarizes the tumor results. 

Table 3-31. Liver Tumors in Three Strains of Mice at 18 Months with Exposure to PFOA 

Only during Gestation and Lactation

Strain

0 

mg/kg/day

0.01 

mg/kg/day

0.1 

mg/kg/day

0.3 

mg/kg/day

0.6 

mg/kg/day

1 

mg/kg/day

3 

mg/kg/day

5 

mg/kg/day

Number of Tumors / Total Number Tested

Tumor Type

CD-1 1/29 L 1/29 HCA 1/37 HCA 6/26

HCA(4), 

HCC, L

NT 2/31

HcyS, L

NT 6/21

HmS(2), 

HCA, 

HCC, 

HcyS, L

SV-129 0/10 NT 1/10

HcyS

0/8 0/6 0/8 NT NT

SV-129-

PPARαKO

0/6 NT 1/10

HCA

2/10

HCA. ICT

NT 1/9

ICT

2/9

HCA

NT

Notes: 

HCA = hepatocellular adenoma, HCC = hepatocellular carcinoma, HcyS = histocytic sarcoma, HmS = hemangeosarcomas, ICT 

= Ito cell tumor, L = lymphoma, NT = not tested.

Perfluorooctanoic acid (PFOA) – May 2016 3-109
It is difficult to draw conclusions regarding the carcinogenicity of PFOA in mice based on 

the data collected because of the small number of animals evaluated in both studies of SV-129 

mice and the lack of PFOA exposure between PND 21 and 18 months for all dose groups. As 

was the case for liver tumors in the Butenhoff et al. study (2012), there is a lack of dose-response 

for total liver tumors, although the four hepatocellular adenomas seen at 0.3 mg/kg/day in CD-1 

mice were significantly greater (p<0.05) than the control. Tumor types varied across the dose 

groups. The authors also reported on preneoplastic basophilic, and eosinophilic foci were 

observed in the CD-1 mice but did not show a response to dose.

An interesting histological finding in both the CD-1 and SV-129 mice was a trend for 

increased Ito cell atrophy and lesion severity across the doses (Filgo et al. 2015). Since Ito cells 

accumulate fat in the liver sinusoids, this observation provides additional support for hepatic 

steatosis as a condition of concern following developmental PFOA exposure. There was an

increase in severity with dose for the Ito cell fat deposits for all but the high-dose group. The Ito 

cell lesion was present in the SV-129 mice, but was not associated with tumors. CD-1 mice had a 

significant increase in Ito cell hypertrophy at 5 mg/kg/day compared to controls, but there was a 

lack of dose-response. The authors concluded that liver damage from PFOA exposure occurring 

early in development is not totally linked to PPAR-α and could progress as animals aged without 

continued dosing, thus compromising liver function and possibly leading to tumor development. 

Inhalation and Dermal Exposures 

No data on the tumorigenic effects of chronic inhalation or dermal exposures to PFOA were 

identified in the literature.

3.3 Other Key Data

3.3.1 Mutagenicity and Genotoxicity

PFOA has been tested for genotoxicity in a variety of in vivo and in vitro assays. The data 

from the in vitro studies are summarized in Table 3-32. 

PFOA was tested in a cell transformation and cytotoxicity assay conducted in C3H10T1⁄2

mouse embryo fibroblasts. The cell transformation was determined as both colony 

transformation and foci transformation. There was no evidence of transformation at any of the 

dose levels tested in either the colony or foci assay methods (Garry and Nelson 1981). 

PFOA was tested twice (Lawlor 1995, 1996) for its ability to induce mutation in the 

Salmonella – E. coli/mammalian-microsome reverse mutation assay. The tests were performed 

both with and without metabolic activation. A single positive response seen in S. typhimurium

TA1537 when tested without metabolic activation was not reproducible. PFOA did not induce 

mutation in either S. typhimurium or E. coli when tested either with or without metabolic 

activation. PFOA did not induce chromosomal aberrations in human lymphocytes when tested 

with and without metabolic activation up to cytotoxic concentrations (Murli 1996a; NOTOX 

2000). Sadhu (2002) reported that PFOA did not induce gene mutation when tested with or 

without metabolic activation in the K-1 line of CHO cells in culture.

Perfluorooctanoic acid (PFOA) – May 2016 3-110
Table 3-32. Genotoxicity of PFOA In Vitro

Test System End-point With Activation Without Activation Reference

C3H10T1⁄2 mouse 

embryo fibroblasts

Cell Transformation NA - Garry and Nelson 

1981

C3H 10T1⁄2 mouse 

embryo fibroblasts

Cytotoxicity NA - Garry and Nelson 

1981

S. typhimurium

TA1537

Gene Mutation - +

(not reproducible)

Lawlor 1995, 1996

E. coli Gene Mutation - - Lawlor 1995, 1996

CHO cells Chromosomal 

Aberrations +, + +, - Murli 1996b, 1996c

CHO cells Polyploidy +, + +, - Murli 1996b, 1996c

Human lymphocytes Chromosomal 

Aberrations - - Murli 1996c; NOTOX 2000

K-1 CHO cells Gene Mutation - - Sadhu 2002

S. typhimurium

TA98, TA100, 

TA102, TA104

Gene Mutation

- - 

Freire et al. 2008

Note: NA= not applicable. 

Murli (1996b, 1996c) tested PFOA twice for its ability to induce chromosomal aberrations in 

CHO cells. In the first assay, PFOA induced both chromosomal aberrations and polyploidy in 

both the presence and absence of metabolic activation. In the second assay, no significant 

increases in chromosomal aberrations were observed without activation. However, when tested 

with metabolic activation, PFOA induced significant increases in chromosomal aberrations and 

in polyploidy (Murli 1996b). The effects were observed only at toxic concentrations (EFSA 

2008). 

PFOA did not display mutagenic activity with or without metabolic activation in 

S. typhimurium strains TA98, TA100, TA102, or TA104 (Freire et al. 2008). 

The data summarized in Table 3-32 suggest that PFOA is not a mutagen. A single positive 

result in S. typhimurium was not reproducible by the same authors and was not replicated in 

other studies. Potential chromosomal effects were found in CHO cells at toxic concentrations, 

but not in human lymphocytes. 

Governini et al. (2015) collected semen samples from 59 healthy nonsmoking patients 

attending a Center for Couple Sterility conference at the University in Siena, Italy. The subjects 

were divided into those that were normozoospermic (13) and those that were oligoasthenoteratozoospermic (46). PFOA was present in 75% of the seminal plasma samples and only 16% of the 

blood samples. Conversely, PFOS was present in 25% of the seminal plasma samples and 84% 

of the serum samples. Sperm were evaluated for the presence of aneuploidy and diploidy, and 

sperm DNA was evaluated for fragmentation using the terminal deoxynucleotidyl transferase 

dUTP nick end labeling (TUNEL) assay. The frequencies of aneuploidy and diploidy were 

significantly greater in the PFAS-positive samples than in the PFAS-negative samples (P<0.001 

and P<0.05, respectively), suggesting the possibility for errors in cell division. The levels of 

fragmented chromatin were significantly increased (P<0.001) for the PFAS-positive group 

compared with the PFAS-negative group. 

Perfluorooctanoic acid (PFOA) – May 2016 3-111
PFOA was tested twice in the in vivo mouse micronucleus assay. PFOA did not induce any 

significant increases in micronuclei and was considered negative under the conditions of this 

assay (Murli 1995, 1996d).

G. Zhao et al. (2010) used AL cells to determine the mutagenicity of PFOA to mammalian 

cells. AL cells are a human-hamster hybrid containing CHO-K1 chromosomes and a single copy 

of human chromosome 11. The significance of human chromosome 11 is that it encodes for 

expression of the human cell surface protein CD59. AL and mitochondria-deficient AL cells were 

incubated with 0, 1, 10, 100, and 200 μmol PFOA for up to 16 days and used in the 3-(4,5-

dimethylthiazol-2-yl)-2,5-diphenyltetrazolium bromide (MTT) viability, mutation, or caspace 

assays. Reactive oxygen species (ROS), nitric oxide, and superoxide anion production were 

measured in the cells, and the effects of ROS/reactive nitrogen species quenchers [0.5% dimethyl 

sulfoxide (DMSO) and 0.2 mM NG-methyl-L-arginine, respectively] on mutagenicity and 

caspace activities were determined. At 100 and 200 μmol PFOA, AL cell viability was 

significantly decreased after incubation for 1, 4, 8, and 16 days. CD59 mutation frequencies were 

increased in AL cells after a 16-day incubation with 200 μmol PFOA. There was no increase in 

mutations in mitochondria-deficient AL cells after incubation with 100 or 200 μmol PFOA.

Production of ROS, nitric oxide, and superoxide anion was significantly increased at 100 and 

200 μmol PFOA after incubation of AL cells for 1, 4, and 16 days. Incubation with DMSO to 

inhibit ROS production significantly decreased the CD59 mutation frequency caused by 

200 μmol PFOA after the 16-day incubation. In contrast, mitochondria-deficient AL cells had no 

increase in ROS or superoxide production after incubation with up to 200 μmol PFOA for 

16 days. 

To assess whether PFOA could induce the apoptotic pathway, caspase-3/7 and caspase-9 

were examined in intact AL cells (mitochondria-deficient cells were not examined). The highest 

concentration significantly increased caspase 3/7 and 9 activities after 1- and 4-day incubations. 

Incubation with 0.5% DMSO and 0.2 mM NG-methyl-L-arginine significantly decreased the 

increased caspace activity induced by 200 μmol PFOA. The results led the authors to suggest 

that mitochondrial-dependent ROS might play an important role in PFOA-induced mutagenicity 

and that induction of caspase activities might be mediated by reactive oxygen and nitrogen 

species.

3.3.2 Immunotoxicity

The impact of PFOA on the immune system has been the subject of considerable research, 

primarily in mice. A number of the early studies by Yang et al. (2000, 2001, 2002a) used highdose exposures of 0.02% to 0.05% PFOA. Later studies by DeWitt et al. (2008, 2009, 2015) and 

Loveless et al. (2008) used a range of doses from < 1 mg/kg/day to 30 mg/kg/day. Most of the 

studies focused on responses associated with the spleen and thymus. Some of the effects 

observed were PPARα-associated, but others are totally or partially independent. There is 

evidence for full or partial reversal of effects in those studies that incorporated a recovery phase.

One study of immunotoxicity used the dermal route of exposure (Fairley et al. 2007). 

Rat. Loveless et al. (2008) administered 0, 0.3, 1, 10, and 30 mg linear PFOA/kg by oral gavage 

to groups of male CD rats (n = 10 per group) for 29 days. The animals received a dose of SRBC 

on day 23. A separate group of high-dose rats were injected with water instead of SRBC. Rat 

body weight was recorded on days 0, 3, and 6–28. At necropsy, blood was collected for 

evaluation of immune system parameters. Cell counts were determined for the thymus and 

Perfluorooctanoic acid (PFOA) – May 2016 3-112
spleen. Total spleen and thymocyte cell counts and organ weights in exposed rats were 

comparable to control. Microscopic examination of the thymus, mesenteric lymph nodes, and 

popluteal lymph nodes revealed no effects in treated rats resulting from PFOA exposure. There 

was no difference observed in immunoglobulin (IgM) titers between treated and control rats. The 

immunological NOAEL was 30 mg/kg/day. 

Mouse. Yang et al. (2000, 2001, 2002a, 2002b) completed a series of studies investigating the 

immunotoxic effects of PFOA. In the first study, Yang et al. (2000) examined the liver, spleen, 

and thymus effects of several known PPARα activators, including PFOA. The researchers 

administered 0.02% PFOA (~40 mg/kg/day) to male C57BL/6 mice in the diet for 2, 5, 7, and 

10 days. At the end of the feeding period, mice were sacrificed and the liver, spleen, and thymus 

were weighed. Administration of PFOA resulted in a significant increase in liver weight relative 

to control even at day 2. Following 5 days of administration, significant decreases in thymus and 

spleen weight were noted. 

A second component of the Yang et al. study (2000) examined the effect of 0.02% PFOA in 

the diet on the cellularity, cell surface phenotype, and cell cycle of thymocytes and splenocytes. 

After 7 days, significant decreases in the total number of thymocytes (85%) and splenocytes 

(80%) were observed. There is a pattern to the development of thymocytes that should be 

considered when evaluating the impact of chemicals on their differentiation. Early thymocytes 

formed in the bone marrow do not express CD4 or CD8 (CD4-

CD8-). In the thymus, they 

differentiate and express both CD4 and CD8 (CD4+CD8+). They also undergo proliferation and 

downregulation of either the CD4 or CD8 protein expression to become either a CD4 or CD8 

thymocyte (Yang et al. 2000). Following exposure to PFOA, the number of thymocytes 

expressing neither CD4 nor CD8 decreased by 57%; the number expressing both CD4 and CD8 

decreased by 95%; the number expressing only CD4 decreased by 64% while those expressing 

only CD8 decreased by 72%. As detected by cell cycle flow cytometry analyses, thymocyte 

proliferation was inhibited based on the number of cells in each stage of the cell cycle.

T-cells (CD3±) and B-cells (CD19±) in the spleen decreased by 75% and 86%, respectively. 

Splenic T-cells are lymphocytes produced in the thymus that carry the CD3+ surface protein 

marking them as T-cells for exportation to the spleen. There are several classes of T-cells that are 

characterized by surface proteins. Yang et al. (2000) found significant decreases in helper 

CD3+T-cells with CD4+ surface proteins (78%) and cytotoxic CD3+T-cells with CD8+ surface 

proteins (74%). The authors suggested that, unlike the CD3+T-cells that originate in the thymus, 

the decrease in CD19+B-cells of the spleen reflects decreased differentiation and maturation in 

the bone marrow where they are formed. 

In the final phase of the Yang et al. study (2000), the effects of in vitro exposure of 

thymocytes and splenocytes to PFOA were examined. The in vitro studies showed spontaneous 

apoptosis occurring in splenocytes and thymocytes after 8 or 24 hours of culturing in the 

presence of varying concentrations of PFOA (50, 100, and 200 μmol). However, under the 

exposure conditions, PFOA did not appear to significantly alter the cell cycle. The only dose 

tested (~40 mg/kg/day) was a LOAEL for its effects on the immunoactive products of the 

thymus and spleen. Recovery can occur with the cessation of exposure as illustrated by the Yang 

et al. study (2001) described below. 

Yang et al. (2001) reported on their examination of the immunosuppressive effects of PFOA.

As was the case in their earlier publication (Yang et al. 2000), the 2001 report includes several 

components. A diet of 0.02% PFOA (~40 mg/kg/day) was fed to C57BL/6 mice for 2–10 days. 

Perfluorooctanoic acid (PFOA) – May 2016 3-113
One group of animals was exposed to PFOA each day until day of sacrifice on days 2, 5, 7, and 

10. At sacrifice, body, liver, and spleen weights were recorded. A second group of animals was

dosed according to the same schedule, but dosing ceased after day 7, and the animals were fed 

normal diets for 2–10 days to monitor recovery from the effects of exposure. In the recovery 

group, animals were sacrificed after 2-, 5-, and 10-day recovery periods. 

The mice that received 0.02% PFOA for up to 10 days experienced significant increases in 

liver weight compared to controls beginning at day 2. Significant decreases in thymus and spleen 

weights were observed starting on day 5. Body weight increased for the first 2 days of the study 

and decreased continuously for the remainder of the exposure period. The activity of palmitoylCoA and lauryl-CoA, biomarkers for PPARα activation and peroxisome proliferation, also were 

increased significantly and increasingly across the exposure period. The impact of PFOA 

exposure was similar to that observed in the Yang et al. study (2000). After administration for 

7 days, the number of thymocytes expressing neither CD4 nor CD8 decreased by 65% following 

exposure to PFOA; the number expressing both CD4 and CD8 decreased by 95%; and the 

number expressing either CD4 or CD8 decreased by 65% and 75%, respectively. T-cell (CD3+) 

splenocytes and B-cell (CD19+) splenocytes decreased by 65% and 75%, respectively. As 

detected by cell cycle flow cytometry analyses, thymocyte but not splenocyte proliferation was 

inhibited. 

The animals that participated in the recovery portion of this study rapidly regained their body 

weight starting on the second day after withdrawal of PFOA. However, the liver weight failed to 

recover even after 10 days. Thymus weight recovery started on day 2 and was completed by day 

10. The spleen weights returned to normal by day 2 post-withdrawal. The increases in thymus 

and spleen weight during recovery were paralleled by increases in total thymocyte and 

splenocyte counts. Thymocyte recovery was apparent on day 5 and complete by day 10, although 

during the first two days of the recovery period, further decreases in the CD4+CD8+, CD4+ and 

CD8+ cells were observed. Flow cytometry evaluation of the distribution of the cells across the 

cell cycle in the recovery group animals demonstrated increases in cell proliferation following 

removal of PFOA from the diet. However, final cell counts did not reach the control values for 

the thymocyte (CD4+ and CD8+) or splenocyte (CD3+ and CD19+) populations evaluated.

In the second component of the Yang et al. study (2001), C57BL/6 mice were administered 

diets consisting of 0.001%–0.05% PFOA (w/w) for 10 days. These doses are equivalent to 

approximately 2.0–100 mg/kg/day. There was a dose-related decrease in spleen and thymus 

weights and a dose-related increase in liver weights accompanied by a corresponding increase of 

palmitoyl-CoA and lauryl-CoA activity. Enzyme activity was significantly increased for all 

doses. Spleen and thymus weights were significantly decreased at doses ≥ 0.01% and higher but 

not at the lower doses; the increases in liver weights were significantly increased for the 0.02% 

and 0.05% doses. With the testing of a broader range of doses, ~20 mg/kg/day was found to be a 

LOAEL for effects on the thymus and spleen and the ~6 mg/kg/day dose a NOAEL. 

Yang et al. (2002a) examined the possible involvement of PPARα in the immunomodulation 

exerted by PFOA. This study made use of transgenic PPARα-null mice (Sv/129), which are 

homozygous with regards to a functional mutation in the PPARα gene. These mice do not exhibit 

peroxisome proliferation or hepatomegaly and hepatocarcinogenesis even after exposure to 

peroxisome proliferators. The mice were fed a diet consisting of 0.02% PFOA (w/w) 

(~40 mg/kg/day) for 7 days. At the end of the feeding period, mice were sacrificed and the liver, 

spleen, and thymus were removed and weighed. The effect of PFOA on peroxisome 

proliferation, cell cycle, and lymphoproliferation was ascertained.

Perfluorooctanoic acid (PFOA) – May 2016 3-114
The results showed that, in contrast to wild-type mice, feeding PPARα-null mice PFOA 

resulted in no significant decrease in body weight. Liver weight in PPARα-null mice fed the 

PFOA diet was significantly increased when compared to control PPARα-null mice, but not 

when compared to wild-type PFOA-exposed mice. Peroxisome proliferation, as measured by 

fatty acid oxidation, was totally lacking in PPARα-null mice. Also, in contrast to wild-type mice, 

feeding PPARα-null mice PFOA resulted in no significant decrease in the weight of the spleen or 

the number of splenocytes. 

There was a decrease in weight and cellularity of the thymus in the PPARα-null mice 

compared to the PPARα-null control mice, but it was not as dramatic as that in the PFOAexposed wild-type mice. In addition, the decreases in the size of the CD4+CD8+ population of 

thymus cells and the number of thymus cells in the S and G2/M phases of the cell cycle were 

lower in PPARα-null mice than they were in the PFOA-exposed wild-type mice, but higher than 

in the PPARα-null control mice. PFOA treatment caused no significant change in splenocyte 

proliferation in PPARα-null mice in response to mitogen exposure, but did show a response in 

the PFOA-exposed wild-type mice as described above.

The series of studies published by Yang et al. (2000, 2001, 2002a) link many of the effects of 

the liver, thymus, and spleen in PFOA-exposed mice to the activation of PPARα. However, there 

were some impacts on the thymus and liver that were independent of PPARα receptor activation. 

PPARα-null mice still showed increases in liver weight and effects on the thymus (small 

decrements in organ weight, thymocyte cellularity, and proliferative cell cycle) following a 7-day 

exposure to approximately 40 mg/kg/day PFOA that were independent of PPARα. 

Yang and colleagues extended their studies of the immunotoxicity of PFOA in a feeding 

study designed to examine the effects of PFOA on specific humoral immune responses in mice 

(Yang et al. 2002b). For this study, 0.02 % PFOA was administered to male C57BL/6 mice for 

10 days. The animals were then evaluated via plaque-forming cell (PFC) and serum antibody 

assays for their ability to generate an immune response to horse red blood cells (HRBCs). 

Ex vivo and in vitro splenic lymphocyte proliferation assays also were performed. The results 

showed that mice fed normal chow had a strong humoral response to challenge the HRBCs, as 

measured by the PFC assay. In contrast, mice fed PFOA responded to HRBC immunization with 

no increase in HRBC-specific PFCs, relative to unimmunized controls. However, in experiments 

where PFOA-treated mice received normal chow following HRBC immunization, there was a 

significant recovery of the numbers of specific PFCs stimulated. The suppression of the humoral 

immune response by PFOA was confirmed by analysis of the serum anti-HRBC response.

In ex vivo experiments, splenocytes isolated from control mice responded to both 

concanavalinA (ConA) and lipopolysaccharide (LPS) with lymphocyte proliferation, as 

measured by thymidine incorporation. However, treating mice with 0.02% PFOA for 7 days 

attenuated the proliferation. In a set of in vitro experiments, PFOA (1–200 μmol) added to the 

culture medium of splenocytes cultured from untreated mice did not cause an alteration of 

lymphocyte proliferation in response to LPS or ConA. 

DeWitt et al. (2008) expanded the repertoire of studies of the immunological effects of 

PFOA by examining various aspects of humoral (antibody production) and cellular immunity.

The first component of their publication had many similarities with the Yang et al. study (2001). 

Adult female C57BL/6J mice (n = 40 per endpoint and 8 per group) were exposed to a single 

daily dose of 30 mg PFOA/kg/day in distilled water by gavage for 10 continuous days. After 10 

continuous days of exposure, half of the mice continued receiving PFOA from day 11 through 

Perfluorooctanoic acid (PFOA) – May 2016 3-115
day 15 (constant group) while the other half received distilled water from day 11 through day 15 

(recovery group). On day 11, 16 mice per group were immunized with sheep red blood cells 

(SRBC) and eight mice per group were injected with BSA. Sacrifices took place on day 16 

(1 day postexposure period) and day 31 (15 days postexposure period). Vehicle and cage 

controls also were included in the study. All groups were monitored for the following effects:

• Body weight and organ weights (day 16, day 31) 

• Serum IgM levels (day 16) 

• Delayed-type hypersensitivity (DTH) foot-pad response to BSA (day 26) 

• Serum IgG levels after booster immunization with SRBC on day 20 (day 31) 

The results for body and organ weights were comparable to those in the Yang et al. study 

(2001). Body weight was significantly decreased from days 8–11 for both PFOA-treated groups 

and on day 16 for mice in the constant exposure group. By day 31, there were no body weight 

differences between the groups. Relative liver weight was significantly elevated in both PFOAtreated groups on days 16 and 31. Absolute and relative spleen and thymus weights of animals in 

both PFOA groups were significantly decreased compared to control groups on day 16. By day 

31, thymus and spleen weights were not statistically different between control and treated mice.

IgM levels following immunization with SRBC were reduced by up to 20% compared to controls 

on postexposure day 1 in both the recovery and constant exposure groups. There were no 

significant differences from controls for SRBC-specific IgG levels and for DTH foot-pad 

responses to the BSA challenge.

The C57BL/6 mice used for the continuous-dosing versus recovery component of the DeWitt 

et al. study (2008) were found to develop ulcerative dermatitis following the PFOA exposure. It 

was determined that this effect was a genetic susceptibility in the strain, and they were not used 

for the dose-response component of the study; the C57BL/6N strain was used in its place. 

Two studies of dose-response were included in the DeWitt et al. (2008) publication. Groups 

of 16 female C57BL/6N mice were given 0, 3.75, 7.5, 15, and 30 mg PFOA/kg/day in the 

drinking water for 15 days during the first experiment. In the second experiment, the doses were

0, 0.94, 1.88, 3.75, and 7.5 mg PFOA/kg/day administered for 15 days in the drinking water. The 

immunological sensitization and postdose monitoring were identical to that used in the constantdosing versus recovery experiment. 

In the first experiment, body weight was significantly decreased from day 8–16 at 30 mg/kg 

PFOA and on day 16 at 15 mg/kg PFOA. As observed previously, liver weights were 

significantly elevated at day 16 and day 31 at all doses. Absolute and relative spleen and thymus 

weights were significantly decreased at ≥15 mg/kg PFOA on day 16. With the exception of the 

absolute thymus weight at 15 mg/kg PFOA, all spleen and thymus weights were similar to 

weights in controls 15 days after dosing. The IgM response to SRBC was significantly reduced at 

≥3.75 mg/kg PFOA in a direct dose-related manner. The IgG response to SRBC challenge was 

slightly but significantly elevated at 3.75 and 7.5 mg/kg PFOA but similar to that of the control 

level at the higher doses. Thus, there was a direct response of IgM, but not IgG, to dose across 

the dose levels. There was no significant change in the DTH response at any dose. The LOAEL 

from the first experiment was 3.75 mg/kg/day dose based on decreased IgM and increased IgG 

response to SRBC immunization and increased liver weights (p<0.05). 

The second dose-response experiment confirmed the 3.75 mg/kg/day dose as the 

immunological LOAEL on the basis of significantly decreased spleen weight, decreased IgM 

Perfluorooctanoic acid (PFOA) – May 2016 3-116
levels on day 16, and increased IgG levels on day 31. The immunological NOAEL was 

1.88 mg/kg/day. BMD analysis of IgM serum titer data gave a lower bound 95% confidence 

limit of 1.75 mg/kg/day on a BMD (one SD) of 3.06 mg/kg/day. Liver weight was significantly 

increased at all doses on days 16 and 31. The LOAEL for increased liver weight was 0.94 mg/kg 

PFOA.

As mentioned earlier, some of the immunological responses observed in the studies of 

immunotoxicity are linked to PPARα activation by PFOA. DeWitt et al. (2015) published results 

for a study in female PPARα KO mice (B6.129S4-Ppartm1GonzN12 mice) and compared them to 

the response of female C57BL/6-Tac wild-type mice. Both T-cell-dependent and T-cellindependent antibody production were evaluated. The doses used in the study of the T-celldependent responses were 0, 7.5, and 30 mg PFOA/kg/day dissolved in deionized drinking water

for 14 days. On day 11, the animals were injected with SRBCs to stimulate an immune response.

PFOA dosing continued for 4 more days (15 days dosed); the following day, the animals were 

sacrificed. Body weight was significantly decreased only in wild-type mice at 30 mg/kg/day. 

Relative spleen weights were significantly decreased (P<0.05) in the wild-type but not the KO

mice at 30 mg/kg/day of PFOA. Relative thymus weights were significantly decreased in the 

wild-type mice at 7.5 mg/kg/day, but not in the KO mice at either dose or the wild-type mice 

receiving 30 mg/kg/day. There was a significant (P<0.05) reduction in the IgM antibody 

response to the SRBC injection at 30 mg/kg/day for both the wild-type and KO animals (n = 6), 

indicating that the response was not totally related to PPAR-α activation. The NOAEL in wildtype and KO animals was 7.5 mg/kg/day and the LOAEL 30 mg/kg/day based on decreased Tcell-dependent IgM antibody response to SRBC. 

To evaluate T-cell-independent responses to PFOA, groups of eight C57BL/6N mice were 

given doses of 0, 0.94, 1.88, 3.75, and 7.5 mg/kg/day in their drinking water. On day 11, they 

were injected with the T-cell-independent antigen dinitrophenyl ficol. At sacrifice (day 16), 

blood was collected for analysis of IgM antibody titers. There was a significant decrease

(p<0.05) in the antibody response by 9.4–10.7% in the animals receiving doses ≥ 1.88 mg/kg/day 

of PFOA. The NOAEL for the T-cell-independent response to dinitrophenyl ficol was 

0.94 mg/kg/day of PFOA and the LOAEL was 1.88 mg/kg/day. Thus, suppression of the T-cellindependent response occurred at a lower dose (1.88 mg/kg/day) than the dose resulting in 

suppression of the T-cell-dependent response (7.5 mg/kg/day). 

The authors looked at changes in lymphocyte populations at 10, 13, and 15 days of exposure 

in the female C57BL/6N mice and saw no significant dose-dependent changes in lymphocyte 

cell types. They concluded that both sets of responses were due to changes in cellular function 

rather than lymphocytotoxicity (DeWitt et al. 2015). 

Loveless et al. (2008) administered 0, 0.3, 1, 10, and 30 mg linear PFOA/kg by oral gavage 

to groups of male CD-1 mice (n = 20 per group) for 29 days. The animals received a dose of 

SRBC on day 24. A separate group of high-dose mice was injected with water instead of SRBC.

Mice were weighed daily. At necropsy, blood was collected for evaluation of immunity

parameters. Cell counts were determined for the thymus and spleen. 

Absolute and relative spleen and thymus weights were significantly decreased at 

≥10 mg/kg/day. The relative spleen weight of mice dosed with 1 mg/kg/day also was 

significantly decreased compared to control animals. Spleen and thymus cell counts were 

significantly decreased and minimal to severe lymphoid depletion/atrophy of the thymus was 

observed at ≥10 mg/kg/day. IgM titers were significantly decreased at ≥10 mg/kg/day. Serum 

Perfluorooctanoic acid (PFOA) – May 2016 3-117
CORT was significantly increased at 10 mg/kg/day and elevated (not statistically significant) at 

30 mg/kg/day. When IgM and CORT were plotted against each other, a negative correlation 

coefficient suggested that increasing CORT levels decreased the ability to make SRBC antibody. 

The LOAEL was 1 mg/kg/day based on decreased spleen weight, and the NOAEL was 

0.3 mg/kg/day. Mice appeared to be more susceptible than rats to immunosuppression from 

PFOA.

Loveless et al. (2008) hypothesized that at least a portion of the thymic response to PFOA 

might be related to physiological stress and increased levels of CORT hormones. DeWitt et al. 

(2009) investigated this hypothesis by comparing the immunological response of 

adrenalectomized (ADX) C57BL/6N female mice to that of sham-operated female mice from the 

same strain. The animals were dosed with 0, 3.75, 7.5, and 15 mg PFOA/kg/day in the drinking 

water for 10 days. Body weight was recorded on dosing days 0, 4, and 8, plus 2 and 5 days 

postdosing. On exposure days 5 and 10, blood and serum were collected for analysis of a broad 

array of clinical chemistry parameters, including activity of liver enzymes indicative of cellular 

damage (e.g., ALP, AST, ALT, LDH, GGT, and SDH), serum lipids (cholesterol and 

triglycerides), and CORT. A baseline measure of CORT was determined from serum samples 

collected before dosing began. One day after cessation of exposure, the mice were immunized 

with SRBC. Four days later, serum was collected and the levels of CORT and IgM were 

determined. 

Body weight in the sham-operated mice declined during dosing in the highest dose group but 

recovered by 5 days postdosing. In the ADX mice, body weight declined during dosing at 7.5 

and 15 mg PFOA/kg/day, but recovered in mice receiving 7.5 mg PFOA/kg/day by 5 days 

postdosing. There were significant increases in ALT, AST, LDH, and SDH at the highest dose 

for the ADX mice indicative of damage to hepatic cell membranes (Table 3-33).

Serum levels of triglycerides significantly decreased compared to controls, with all doses for 

the sham-operated mice on day 5 of dosing but only for the 7.5- and 15-mg/kg/day doses in the 

ADX mice. Cholesterol levels were significantly decreased (p<0.05) in the sham-operated mice 

at the highest dose, but no differences in cholesterol levels were observed in the ADX mice.

Table 3-33. Selected Clinical Chemistry Parameters in Mice Treated with PFOA for 5 Days 

Dose (mg/kg/day) ALT AST LDH SDH

Sham-Operated

0 39.52±2.50 121.56±17.96 320.57±29.84 46.43±1.03

3.75 43.88±0.93 104.07±10.24 293.92±68.65 39.31±3.32

7.5 56.96±6.78 95.55±10.22 262.71±35.60 39.02±7.77

15 62.57±3.15 89.07±1.30 191.76±22.25 46.87±1.46

ADX

0 26.96±1.78 73.53±4.70 176.50±19.32 33.05±1.58

3.75 29.67±1.62 76.58±3.38 222.69±19.18 37.95±2.35

7.5 39.04±2.59 83.79±8.94 320.45±53.34 46.35±1.42

15 94.23±31.66* 126.47±16.39* 435.57±81.42* 77.61±19.89*

Source: DeWitt et al. 2009

Note: *

 = p<0.05 versus corresponding sham control or ADX control group. 

Perfluorooctanoic acid (PFOA) – May 2016 3-118
After 10 days, there were no significant differences in liver enzymes for the ADX or sham 

mice. However, there was a dose-related trend towards increased levels of liver enzymes for the 

PFOA-exposed sham-operated animals and for LDH in the PFOA-exposed ADX animals 

(Table 3-34).

Table 3-34. Selected Clinical Chemistry Parameters in Mice Treated with

PFOA for 10 Days 

Dose (mg/kg/day) ALT AST LDH SDH

Sham-Operated

0 51.51±14.62 93.30±6.33 333.48±86.86 54.60±16.72

3.75 79.26±33.87 123.73±15.20 404.14±59.89 45.50±10.15

7.5 135.57±38.18 142.66±15.59 490.44±69.14 80.71±14.59

15 344.53±235.63 242.92±117.62 595.01±137.37 89.20±26.03

ADX

0 128.22±24.80 106.00±8.86 236.96±30.23 61.88±8.87

3.75 282.23±193.54 217.10±3.48 379.61±80.67 68.78±24.88

7.5 89.79±21.54 99.78±12.59 574.65±236.38 52.07±11.98

15 261.14±75.95 181.40±32.94 614.05±144.95 101.93±24.00

Source: DeWitt et al. 2009

At the end of dosing, corticosteroid levels in the sham-operated animals were greatly 

elevated compared to the levels in the control animals at all doses, and the difference was 

statistically significant at the highest dose. By 5 days postdosing, the CORT levels had declined 

for all doses but were still elevated compared to controls for the 7.5- and 15-mg/kg/day groups. 

In the animals lacking their adrenal glands, there were no statistically significant differences in 

the hormone levels. IgM levels were significantly lower than controls at the highest dose for the 

sham-operated animals and at the two highest dose groups for the ADX mice. However, when 

comparing the sham mice to the ADX mice, the only significant difference in IgM was found for 

the 7.5-mg/kg/day animals. On the basis of data, it appears that stress-related CORT production 

did not have a major impact on the IgM response to the SRBC inoculation. 

Son et al. (2009) administered 0, 2, 10, 50, and 250 ppm PFOA (equivalent to 0, 0.49, 2.64, 

17.63, and 47.21 mg/kg PFOA) in the drinking water to 4-week-old male ICR mice for 21 days 

to determine if PFOA alters T lymphocyte phenotypes and cytokine expression in mice. The 

spleen, thymus, and trunk blood were collected at sacrifice. Sections of the spleen and thymus 

were processed for histological examination. Splenic and thymic expression of mRNA from 

proinflammatory cytokines—including tumor necrosis factor-α, interleukin-1β, and interleukin6, and the proto-oncogene c-myc—were analyzed using RT-PCR. Flow cytometry was used to 

phenotype the splenic and thymic lymphocyte populations. 

Spleen and thymus weights were slightly decreased in mice treated with PFOA. Enlargement 

with marked hyperplasia of the white pulp and increased cellular density of the lymphoid 

follicles were observed in spleens at 250 ppm. In the thymus, decreased cortex and medulla 

thickness and densely arranged cortex lymphoid cells were observed at 250 ppm. Tumor necrosis 

factor-α, interleukin-1β, interleukin-6, and c-myc expression were significantly elevated at 

250ppm in the spleen. Interleukin-1β expression also was elevated at 50 ppm in the spleen. In the 

thymus, c-myc expression was significantly elevated by treatment with 50 and 250 ppm PFOA.

Perfluorooctanoic acid (PFOA) – May 2016 3-119
The splenic and thymic lymphocyte population was altered by PFOA treatment, as shown in 

Table 3-35. A 50% decrease was observed in splenic CD8+ lymphocytes at all PFOA doses, and 

increases in splenic CD4+ lymphocytes of 43% and 106% at 50 and 250 ppm PFOA, 

respectively, were observed. In the thymus, a 110% increase was observed in thymic CD8+

lymphocytes at 250 ppm, but thymic CD4+ lymphocyte populations were not affected and 

CD4+CD8+ populations were decreased at 50 and 250 ppm PFOA. The lowest dose tested 

(0.49 mg/kg/day) was a LOAEL for CD4- and CD8+ splenocytes. 

Table 3-35. Impact of PFOA on Splenic and Thymic Lymphocyte Populations

Dose (mg/kg/day)

Spleen 0.49 2.64 17.63 47.21

CD4-CD8- ↑ - - ↓

CD4+CD8- - - ↑ ↑

CD4-CD8+ ↓ ↓ ↓ ↓

CD4+CD8+ ↓ ↓ - -

Thymus

CD4-CD8- - - - ↑

CD4+CD8- - - - -

CD4-CD8+ - - - ↑

CD4+CD8+ - - ↓ ↓

Source: Son et al. 2009

Notes:

↑ Significantly increased compared to control (p < 0.05). 

↓ Significantly decreased compared to control (p < 0.05). 

- Not significantly different from control. 

Qazi et al. (2009) investigated the impact of PFOA on the innate immune system. Adult male 

C57BL/6 (H-2b

) mice were administered 0.001% and 0.02% PFOA (~2 or 40 mg/kg) in the diet 

(w/w) for 10 days. After the last dose, all mice were sacrificed. Sacrifice was delayed for a 

subset of the animals until 2 hours after they had received an LPS injection to stimulate an 

immunological response. Blood, peritoneal exudate cells, liver, epididymal fat, spleen, thymus, 

and bone marrow were recovered. The blood, peritoneal exudate, bone marrow, and spleen were 

evaluated for total and differential white blood cell counts and concentrations of tumor necrosis 

factor (TNF-α) and interleukin 6 (IL-6). 

Consistent with other studies of the 0.02% dose, there was a significant increase in liver 

weight after the 10-day exposure. Body weight, thymus weight, spleen weight, and epididymal 

fat depots were decreased. Food consumption in these animals was reduced by 35%, which 

might have played a role in the reduced body, organ, and tissue weights. Compared to the 

controls, there was a significant decrease in total white cells, lymphocytes, and neutrophils at 

0.02% PFOA. This same dose was associated with a decrease in total white cell count in bone 

marrow and spleen, and an increase in the proportion found as macrophages in the bone marrow, 

spleen, and peritoneal cavity. Although the total number of macrophages was not reduced in the 

peritoneal cavity and spleen, it was reduced in the bone marrow. The increase in the proportion 

of macrophages reflects a decrease in other white cell populations. There was significant increase 

in the concentration of IL-6 in all of the 0.02 % dosed animals, but only those receiving the LPS 

injections showed a significant increase in TNF-α. The 0.001% dose (about 2 mg/kg/day) was a 

NOAEL.

Perfluorooctanoic acid (PFOA) – May 2016 3-120
The data available on immunological responses of animals following oral exposure to PFOA 

are extensive, especially as they apply to mice. A number of the studies used doses of about 

40 mg/kg/day. However, studies conducted at lower doses find effects on the spleen and/or 

thymus at doses from 0.5 to 2 mg/kg/day. Activation of PPARα appears to augment the response 

based on studies in PPARα-null mice but is not necessary (Yang et al. 2002a). There are 

differences between mice and rats with mice showing a response at a lower dose (Loveless et al. 

2008). Cessation of dosing can reverse some of the observed effects in mice (Yang et al. 2001). 

Inhalation Exposure 

No data on the effects of inhalation exposure on immunological endpoints were identified in 

the literature.

Dermal Exposure

Fairley et al. (2007) carried out a complex study of toxicity and respiratory hypersensitivity 

to ovalbumin (OVA) as impacted by dermal exposure to PFOA dissolved in acetone compared to 

acetone alone. There were several phases to the study. In the first phase, a range-finding study, 

PFOA was applied to each ear of female BALB/c mice (n = 5 per group) at doses of 0, 0.01%, 

0.1%, 0.25%, 0.5%, 1.0%, and 1.5% PFOA (equivalent to 0, 0.25, 2.5, 6.25, 12.5, 25, and 

50 mg/kg/day) for 4 days. Six days after last inoculation, the animals were sacrificed. The liver, 

spleen, and thymus were recovered and weighed. A significant increase in liver weight was 

observed at ≥6.25 mg/kg PFOA. Spleen weight was significantly decreased in mice dosed with 

25 mg/kg and 50 mg/kg PFOA, and thymus weight was significantly decreased in mice at the 

highest dose (p<0.05). The cell counts in the spleen were significantly decreased compared to 

control at all doses and for the highest two doses in the thymus. The LOAEL was 

6.25 mg/kg/day based on a statistically significant increase in liver weight (p<0.01), and the 

NOAEL was 2.5 mg/kg/day. 

In the second phase of the Fairley et al. study (2007), groups of 5–15 animals were dosed 

dermally on the ears for 4 days with doses of 0, 0.5%, 0.75%, 1.0%, and 1.5% PFOA (equivalent 

to 0, 12.5, 18.75, 25, and 50 mg/kg/day). On days 1 and 10, they were injected i.p. with either 

2.0 mg alum or 7.5 μg OVA and 2.0 mg alum in a phosphate-buffered saline solution (100μL).

Four days after the last inoculation, the animals were sacrificed and blood was collected by 

cardiac puncture. Liver, spleen, and thymus were recovered and weighed; spleen and thymus 

cellularities were determined. A significant (p<0.01) increase in liver weight and decrease in 

spleen weight and spleen cellularity occurred at all doses. Thymus weight and cellularity were 

significantly decreased (p<0.01) at ≥18.75 and ≥ 25 mg/kg/day, respectively. There were no 

significant differences in the CD4+, CD8+, CD4-

8-, or CD3e T-cells. CD3e protein is expressed 

by both thymocytes and mature T-cells.

Levels of IgE and OVA-specific IgE were measured in the control and dosed animals by 

enzyme-linked immunosorbent assay. IgE is the immunoglobulin that is best correlated with

respiratory allergic responses. It functions to stimulate mast cells and basophils to release 

histamine and other mediators of inflammation (Saladin 2004). The IgE response was increased 

in a dose-related fashion compared to the OVA control for all the PFOA-treated animals; the 

increase was significant (p<0.05 or 0.01) at doses ≥18.75 mg/kg/day. The OVA-specific IgE 

response did not demonstrate a direct response to dose, but there was a significant increase 

(p<0.05) for the 18.75- and 25-mg/kg/day groups. The OVA-specific response for the three

highest dose groups was practically indistinguishable from the OVA control. 

Perfluorooctanoic acid (PFOA) – May 2016 3-121
The dermal LOAEL was 12.5 mg/kg/day based on increased liver weight and decreased 

spleen weight and cellularity. No NOAEL was established.

In the third part of the Fairley et al. study (2007), mice (n = 5) were dosed dermally via their 

ears for 4 days as described above (0, 12.5, 18.75, 25, and 50 mg/kg/day). On days 19 and 26 

after the start of dosing, they were challenged by pharyngeal aspiration of 250 μg OVA in the 

phosphate-buffered saline vehicle and sacrificed 24 hours after the last challenge. There was a 

dose-related decrease in number of splenocytes carrying the B220+ marker (expressed on B-cells, 

activated B-cells, and subsets of T- and natural killer- [NK-] cells) compared to the OVA 

controls. The changes were significantly different for the 25-mg/kg/day (p<0.05) and 

37.5-mg/kg/day (p<0.01) groups. 

After the day 19 challenge, the mice (n = 5) were placed in a plethysmography chamber for 

measurement of enhanced pause airway respiration (PenH) values. PenH values reflect volume 

of air in the lungs. Once in the chamber, they were challenged with nebulized methacholine for 

3 mins followed by 2 mins of fresh air. The PenH measures were recorded every 30 seconds over 

the next 5 hours. The area under the plasma concentration time curve (AUC) for the PenH

measures was determined after correction for baseline (acetone control, no OVA or PFOA). An 

AUC of 1.6 was considered to be a positive response. Twenty-four hours later, blood was drawn 

from the abdominal artery and the mice were sacrificed. The lungs were recovered for 

histological analysis. An increase in antigen-specific hyperactivity response to PFOA, in both the 

PenH values and the number of animals responding, was observed at doses up to about 

25 mg/kg/day. The PenH AUC was significantly (p<0.05) increased in mice treated with 

25 mg/kg/day PFOA and OVA compared to the OVA control mice, but there was no significant 

difference between the OVA control and the animals exposed to 50 mg/kg/day PFOA and OVA. 

The LOAEL for the PenH response was 25 mg/kg/day, and the NOAEL was 18.75 mg/kg/day. 

Histopathological evaluation of the lungs revealed macrophage and eosinophil infiltration in 

response to the challenge with 250 μg OVA by pharyngeal aspiration. The severity of the 

response increased with increasing concentrations of PFOA. Eosinophils and macrophages were 

found in the interstitial, peribronchiole, and perivascular areas. Neutrophils, lymphocytes, and 

some multinucleated giant cells also were observed. Increased secretory matter, sloughing of 

epithelium, and secretory cell necrosis were observed in mice exposed to all concentrations of 

PFOA and OVA. The response was not observed in the mice exposed to only PFOA. The authors 

concluded that dermal exposure to PFOA was immunotoxic and enhanced the airway 

hypersensitivity response to OVA suggesting that PFOA may augment the IgE response to 

environmental allergens.

In vitro. In a pilot study, Brieger et al. (2011) examined the effects of PFOA on human 

leukocytes. Peripheral blood mononuclear cells (PBMCs) were obtained from the blood of 11 

voluntary donors (n = 6 females, 5 males). PBMCs were incubated with varying concentrations 

of PFOA followed by assays for cell viability, proliferation, and NK cell activity. The human 

promyelocytic leukemia cell line, HL-60, was used in cell viability and monocyte differentiation 

assays. The various components of the assays employed are identified as follows: 

• In the cell viability assay, the PBMCs were incubated with 0–500 μg/mL for 24, 48, and 

72 hours, and HL-60 cells were incubated with 0–125 μg/mL PFOA for 24 hours.

• In the proliferation assay, the PBMCs were incubated with 0–100 μg/mL PFOA for 24 

hours, labeled with 6-carboxyfluorescein succinimidyl ester (CFSE), stimulated with 

Perfluorooctanoic acid (PFOA) – May 2016 3-122
ConA, a T-cell mitogen (5 μg/mL to half of all samples), and incubated for an additional 

72 hours. 

• For the NK assays, PBMCs were incubated with 0–100 μg/mL PFOA for 24 hours 

followed in incubation of 3 hours with K562 target cells (12.5:1 ratio) labeled with 

CFSE.

• In the monocyte differentiation assay, HL-60 cells were incubated with 0–100 μg/mL 

PFOA for 72 hours. Half of each sample was stimulated with 25 nmol calcitrol, 

1α,25-dehydroxyvitamin D3 (1,25D3) 24 hours into the incubation period. Expression 

of CD11b and CD14 were measured as markers of differentiation.

• Whole blood was incubated with 0–100 μg/mL PFOA in the presence or absence of 

25 μg/mL phytohemagglutinin (PHA), T-cell cytokine secretion stimulator for 48 hours 

in quantification assays for the cytokines TNF-α and IL-6. LPS (0 or 250 ng/mL) was 

added to whole blood incubated with 0.1–100 μg/mL PFOA either 4 or 24 hours prior to 

the end of the 48-hour incubation period to determine TNF-α and IL-6 release.

The plasma concentrations of PFOA were 3.3, 1.56, and 4.19 ng/mL for all, female, and male 

volunteers, respectively. Exposure to 31.3 and 62.5 μg/mL PFOA significantly increased PBMC 

viability at the 72-hour endpoint, and 62.5 μg/mL PFOA significantly increased cell viability at 

24 hours. Exposure to 250 and 500 μg/mL PFOA significantly decreased cell viability at all time 

endpoints. Exposure to PFOA did not affect HL-60 cell viability. A trend towards slightly 

augmented proliferation was observed following incubation with PFOA. Of the nine samples 

used, cells from six donors had slightly increased proliferation and t had no response. In cells 

incubated with ConA and 100 μg/mL PFOA, a nonsignificant decrease in the number of 

proliferating cells was observed. PFOA decreased NK cell activity approximately 16% (not 

statistically significant). In the presence of 1,25D3, 100 μg/mL PFOA significantly increased the 

percentage of HL-60 cells expressing CD11b and CD14. There were no differences in monocyte 

differentiation in the absence of 1,25D3. 

In whole blood, exposure to PFOA for 48 hours caused a slight increase in TNF-α and IL-6 

levels. In the presence of PHA, a slight dose-dependent decrease in TNF-α and IL-6 was 

observed. There was a slight dose-dependent decrease in TNF-α release when LPS was added 

4 hours before the end of the incubation period and a slight dose-dependent increase in IL-6 

release when LPS was added 24 hours prior to the end of incubation. The authors also looked at 

the correlation between basal PFOA concentration and cytokine release. A significant association 

was observed between PFOA concentration and the release of LPS-induced TNF-α and IL-6 by 

peripheral monocytes. The authors suggested that the trends observed at the lower concentrations 

might show an impact on human immunity with a larger population. 

Ahuja et al. (2009) examined the effects of PFOA on the production and activation of human 

monocyte-derived dendritic cells. These cells are responsible for a primary immune system 

response by activating lymphocytes and secreting cytokines. Peripheral monocytes and immature 

dendritic cells were incubated with 200 μmol PFOA from day 0 to day 6 or 8 to determine the 

impact on phenotype and cytokine secretion. Maturation stimulus (i.e., prostaglandin E2, tumor 

necrosis factor, interleukin 1β, and IL-6) was added during the last 48 hours of incubation to 

induce dendritic cell maturation. Mixed leukocyte reaction assays were conducted to determine if 

immature dendritic cells could stimulate T-cells. Cytokine (HLA-DR, CD25, CD80, CD83, and 

CD86) expression was measured as a determination of maturity. HLA-DR is a cytokine that 

presents antigens to elicit T-cell response. CD25, 80, 83, and 86 are cell receptors that act as coreceptors in T-cell activation; and interleukin 12p40 and 10 stimulate T-cells. Mature cytokinePerfluorooctanoic acid (PFOA) – May 2016 3-123
activated dendritic cells secrete interleukin 12p40 and interleukin 10 as antiinflammatory 

cytokines. 

In peripheral monocytes incubated with only PFOA from day 0–6 or day 0–8, expression of 

HLA-DR and CD86 was increased compared to expression in control cells, indicating that 

immature dendritic cells were present. In the mixed leukocyte reaction assay, the ability to 

stimulate T-cells was not different between immature dendritic cells generated in the absence or 

presence of PFOA.

To determine if PFOA impacted the differentiation of immature dendritic cells to mature 

dendritic cells, immature dendritic cells were incubated with 200 μmol of PFOA for 6 days and 

the maturation stimulus was added for the final 2 days of incubation. There were no differences 

in cytokine (CD25, CD80, CD83, and CD86) expression between cells incubated with PFOA and 

control cells. Expression of interleukin 12p40 and interleukin 10 was significantly inhibited by 

PFOA in mature cytokine-activated dendritic cells, even in the presence of maturation stimulus 

during the last 48 hours of incubation. The result suggested that exposure to PFOA during 

generation of dendritic cells affected the phenotype and cytokine production of human dendritic 

cells and could lead to immunomodulation in the development of the immune response. 

3.3.3 Hormone Disruption

Thyroid. Martin et al. (2007) administered 20 mg PFOA/kg to adult male Sprague-Dawley rats 

(n = 4 or 5) for 1, 3, or 5 days by oral gavage and determined the impact of PFOA on hormone 

levels. Blood was collected via cardiac puncture and the serum was analyzed for cholesterol, 

testosterone, FT4 and total T4, and total T3. RNA extracted from the livers was used for gene 

expression profiling, genomic signatures, and pathway analyses to determine a mechanism of 

toxicity.

Following a 1-day, 3-day, and 5-day dose, a significant decrease (p<0.05) was observed in 

serum cholesterol (~↓45-72%), total T4 (~↓83%), FT4 (~↓80%), and total T3 (~↓25-48%).

Serum testosterone was significantly decreased (p<0.05, ~↓70%) following a 3-day and 5-day 

PFOA dose. PFOA treatment was matched to hepatotoxicity-related genomic signatures, as well 

as signatures for hepatocellular hypertrophy, hypocholesterolemia, hypolipidemia, and 

peroxisome proliferation. PPARα nuclear regulated genes were induced by PFOA treatment.

Genes associated with the thyroid hormone release and synthesis pathway including Dio3, which 

catalyzes the inactivation of T3, and Dio1, which deiodinates prohormone T4 to bioactivate T3, 

were affected by PFOA. Treatment with PFOA resulted in significantly upregulated expression 

of Dio3 and downregulated expression of Dio1 (p<0.05). Expression of HMG-CoA reductase 

(involved in cholesterol biosynthesis) was significantly upregulated and cholesterol biosynthesis 

was downregulated in a manner consistent with PPARγ agonists. The authors suggested a link 

between PFOA, PPAR, and thyroid hormone homeostasis based on work by Miller et al. (2001), 

who observed decreased serum T4 and T3 levels and increased hepatic proliferation following 

exposure to peroxisome proliferators. They also noted that PFOA exhibited similarities to 

compounds that induce xenobiotic metabolizing enzymes through PPARγ and CAR. The 

20-mg/kg/day dose was a LOAEL for the effects monitored after a 5-day exposure.

Reproductive Hormones. Cook et al. (1992) gavage-dosed male CD rats (n = 15 per group) for 

14 days with 0, 1, 10, 25, and 50 mg PFOA/kg/day to examine the possibility that an endocrinerelated mechanism might explain Leydig cell adenomas observed in rats. A separate control 

group was pair-fed to the 50-mg/kg/day group. Blood and testicular interstitial fluid were 

Perfluorooctanoic acid (PFOA) – May 2016 3-124
collected at necropsy for hormone analysis including testosterone, estradiol, and LH. A separate 

group of rats was dosed with 0 and 50 mg PFOA/kg/day for 14 days and challenged with 

100 IUs of human chorionic gonadotropin (hCG) or 2 mg naloxone/kg 1 hour prior to necropsy 

to induce testosterone concentrations. Blood was collected and analyzed for testosterone and LH.

Serum from rats challenged with 100 IUs hCG also was analyzed for P, 17 

α-hydroxyprogesterone, and androstenedione.

The relative liver weight at 10, 25, and 50 mg PFOA/kg/day was significantly increased 

(p<0.05). The accessory sex organ unit relative weight was significantly decreased (p<0.05) at 

25 and 50 mg PFOA/kg/day compared to those weights in control rats. The relative weights of 

the liver, accessory sex organ unit, and ventral prostate were significantly decreased at the 

highest dose compared to the pair-fed control. 

Serum estradiol was significantly increased at ≥10 mg PFOA/kg compared to the control. No 

differences were observed in testosterone and LH between the treated rats and control. In the 

challenge experiment, serum testosterone was significantly decreased (p<0.05) by treatment with 

50 mg PFOA/kg after challenge with 100 IUs hCG. No differences in testosterone concentration 

were observed in the naloxone-challenged rats, and no differences in LH were observed after 

either challenge. In the hCG-challenged rats, androstenedione was significantly reduced at 50 mg 

PFOA/kg, but no differences in concentrations were observed in P or 17 α-hydroxyprogesterone 

between control and treated rats. The authors suggested that the observed decreased serum 

testosterone levels could be due to decreased conversion of 17 α-hydroxyprogesterone to 

androstenedione as a result of increased serum estradiol levels. The LOAEL was 10 mg/kg based 

on increased liver weight and increased serum estradiol levels, and the NOAEL was 1 mg/kg. 

Biegel et al. (1995) conducted in vitro, in vivo, and ex vivo studies to determine the effects of 

PFOA on Leydig cell function. In the in vitro study, Leydig cells were cultured with ± 2 IUs 

hCG (for final 3 hours) and 0, 100, 200, 250, 500, 700, and 1000 μmol PFOA for a total of 

5 hours and then analyzed for testosterone concentration. Leydig cells also were incubated 

±500 μmol PFOA and analyzed for testosterone and estradiol at 0, 0.5, 1, 3, 6, 12, 24, and 

48 hours. 

In the in vitro studies, there was no effect of PFOA treatment on testosterone in Leydig cells 

cultured without hCG. In cells cultured with hCG, PFOA caused a dose-dependent decrease in 

testosterone production. At 100 μmol PFOA plus hCG, the testosterone concentration was 

significantly increased compared to cells treated with only 100 μmol PFOA. Cytotoxicity 

occurred at ≥ 750 μmol PFOA. In the time course experiment, 500 μmol PFOA significantly 

inhibited hCG-stimulated release of testosterone at time points of at least 3 hours compared to 

control. Estradiol levels of PFOA-treated Leydig cells at 48 hours were statistically greater than 

the control. 

Male CD rats were gavage-dosed for 14 days with 0, 0 pair-fed, or 25 mg PFOA/kg and 

necropsied on day 15. Blood and testicular interstitial fluid were collected for hormone analysis.

Liver samples were collected for analysis of peroxisomal β-oxidation and microsomal aromatase 

activities. Serum estradiol was significantly increased (p<0.05) by 25 mg PFOA/kg when 

compared to the ad libitum and pair-fed control rats. Testicular interstitial fluid testosterone 

concentration was significantly decreased (p<0.05) and microsomal aromatase activity, and 

peroxisomal β-oxidation activity were significantly increased (p<0.05) in PFOA-treated rats 

compared to the pair-fed control rats.

Perfluorooctanoic acid (PFOA) – May 2016 3-125
Leydig cells from the treated rats in the in vivo study were isolated and cultured for analysis 

of testosterone concentration for the ex vivo study. An increase of 8.6-fold in testosterone 

production (p<0.05) was observed in Leydig cells isolated from PFOA-treated rats. The authors 

suggested that the increased serum estradiol levels resulted from liver aromatase induction by 

PFOA, and that PFOA could directly affect Leydig cell function.

Liu et al. (1996) treated adult male Crl:CD(BR) rats (n = 15 per group) with 0, 0 pair-fed, 

0.2, 2, 20, and 40 mg PFOA/kg for 14 days by oral gavage to determine the impact of PFOA on 

aromatase activity. At necropsy on day 15, blood was collected for serum estradiol 

determination. Liver samples were collected for determination of microsomal aromatase activity 

and total P450 concentration. The testes were collected and testicular aromatase was determined.

The body weight of rats treated with ≥ 20 mg PFOA/kg was significantly decreased (p<0.05) 

compared to the control rats. Pair-fed control rats also had significantly decreased body weight 

compared to the control rats. Body weight was not different between the pair-fed control rats and 

rats dosed with 40 mg/kg PFOA. Absolute and relative liver weights were significantly increased 

(p<0.05) at ≥ 2 mg PFOA/kg. Relative testes weight was significantly increased at ≥20 mg 

PFOA/kg, but the differences were due to decreased body weight. There were no differences

observed in testicular aromatase activity. In the remaining analysis, the pair-fed control group 

was similar to the ad libitum control group. The protein yield of hepatic microsomes was 

significantly increased at ≥0.2 mg PFOA/kg, and hepatic aromatase activity, total hepatic 

aromatase activity adjusted for liver and body weight effects, and serum estradiol were 

significantly increased (p<0.05) at ≥2 mg PFOA/kg. The maximum increase in total hepatic 

aromatase activity was 16-fold and the increase was twofold for serum estradiol. A significant 

correlation (p<0.0001) was observed between total hepatic aromatase activity and serum 

estradiol. The aromatase activity in liver microsomes isolated from control rats and incubated for 

2 hours with PFOA was significantly decreased at ≥ 100 μmol. The authors estimated the half 

maximal effective concentration (EC50) values for the outcomes, and they are shown in Table 

3-36. Liu et al. (1996) concluded that the PFOA-increased protein yields suggested induction of 

the ER resulting in aromatase induction, which led to increased serum estradiol. However, PFOA

also inhibited aromatase activity, which would explain why serum estradiol was only increased 

up to twofold. 

Table 3-36. Estimated EC50 Values

Parameters EC50 (mg PFOA/kg)

Hepatic microsome protein yield 0.53

Hepatic microsomal aromatase activity 0.76

Absolute liver weight 1.07

Relative liver weight 1.56

Serum estradiol 3.24

Terminal body weight 11.65

Source: Liu et al. 1996

Note: EC50= half-maximum response. 

A separate component of the Liu et al. study (1996) examined the effect of PFOA on 

aromatase activity in cultured hepatocytes and is discussed below. Aromatase is a cytochrome 

P450 enzyme localized to the ER that catalyzes the conversion of androgens to estrogens. The 

cultured hepatocytes isolated from control male CD rats were incubated with 0–1000 μmol

PFOA and the aromatase activity was evaluated after 18, 42, and 66 hours (Liu et al. 1996). 

Perfluorooctanoic acid (PFOA) – May 2016 3-126
Compared to aromatase activity in time-matched control cultures, PFOA caused a decrease in 

aromatase activity after 18 and 42 hours incubation with hepatocytes and an increase after the 

66-hour incubation period. 

In their study examining the impact of PFOA on aromatase activity, Liu et al. (1996) also 

examined the impact of PFOA on peroxisome β-oxidation and cytochrome P450 activities. Male 

Crl:CD BR (CD) rats (n = 15 per group) were orally dosed with 0, 0 pair-fed, 0.2, 2, 20, and 

40 mg PFOA/kg for 14 days. Liver samples were collected for determination of microsomal total 

cytochrome P450 concentration and peroxisome β-oxidation activity. Total cytochrome P450 

was significantly increased (p<0.05) at ≥20 mg PFOA/kg and β-oxidation activity was increased 

at ≥2 mg PFOA/kg. The estimated EC50s for total cytochrome P450 and β-oxidation were 18.18 

and 2.19 mg PFOA/kg, respectively. The LOAEL was 2 mg/kg based on increased liver weight, 

serum estradiol, and hepatic aromatase activity, and the NOAEL was 0.2 mg/kg. 

Hines et al. (2009) examined the roles that exposure to PFOA and ovarian hormones might 

play in animals exposed during gestation compared to during their adult years. Timed-pregnant 

CD-1 mice were gavage-dosed in two blocks on GDs 1–17, but not thereafter. Block 1 animals 

were dosed with 0, 1, 3, and 5 mg PFOA/kg, and block 2 animals were dosed with 0, 0.01, 0.1, 

0.3, 1, and 5 mg PFOA/kg/day. At birth, pups were pooled within each block and dose group and 

randomly redistributed among the dams (10 pups per litter). Offspring were weaned at 3 weeks, 

and a subset of females from each dose group (0, 0.01, 0.1, 0.3, 1, and 5 mg PFOA/kg/day) was 

OVX at weaning or the day after weaning. All animals were observed until they reached 18 

months of age. 

Body weight was recorded weekly for the first 9 months of age, followed by monthly body 

weight recordings over the next 9 months. As the animals matured, they were evaluated for the 

endpoints listed in Table 3-37. A group of naive 8-week-old adult mice were dosed for 17 days 

with 0, 1, and 5 mg PFOA/kg/day to compare the impact of exposure in adult animals to those 

occurring during gestation. At 18 months of age, the mice were sacrificed. Blood, retroperitoneal 

abdominal fat, interscapular brown fat, organs, and abnormal growths were collected at 

necropsy. 

Table 3-37. Data Collection for Female Mice Gestationally Exposed to PFOA

Test Age at Test Dose (mg/kg/day) Group

Glucose tolerance test 15-16 weeks 0, 1, 5 Intact

Serum leptin and insulin 21-33 weeks 0, 0.01, 0.1, 0.3, 1 Intact, OVX

Body mass composition 42 weeks 0, 0.01, 0.1, 0.3, 1 Intact

Glucose tolerance test 17 months 0, 0.1, 1, 5 Intact

Food consumption 17 months 0, 0.1, 1, 5 Intact

Serum estradiol 18 months 0, 0.01, 0.1, 0.3, 1, 5 Intact

Source: Hines et al. 2009

Body weight of offspring born to dams exposed to 5 mg PFOA/kg was significantly 

decreased (p<0.05) on PND 1 and through 18 months of age compared to control offspring body 

weight. At weaning, the body weight of offspring born to dams exposed to 1 mg PFOA/kg/day 

was significantly decreased (p<0.05) compared to control offspring body weight. A significant 

increase (p<0.05) in body weight, due to more rapid weight gain after week 10, compared to 

intact control body weight, was observed in intact mice exposed to 0.01–0.3 mg PFOA/kg/day 

Perfluorooctanoic acid (PFOA) – May 2016 3-127
during gestation. Body weight of intact mice gestationally exposed to 0.01–0.3 mg PFOA/kg/day 

was comparable to body weight of control mice at 18 months. 

Due to the increased weight gain observed in intact mice exposed to PFOA during gestation, 

glucose tolerance tests were carried out along with determination of serum insulin concentration.

In cases of insulin resistance, plasma glucose and insulin levels are elevated and the insulin 

response is lessened. Insulin resistance also has been associated with excess abdominal fat.

Serum leptin levels also were determined as increased leptin levels have been associated with a 

leptin-resistance mechanism of action (MOA) for increased weight gain in humans. Body mass 

composition was used to determine if there were differences in body fat between the intact 

groups, and feed consumption was recorded to determine if consumption played a role in body 

weight differences in intact control and intact gestationally exposed mice. Serum estradiol was 

measured to determine if PFOA impacted hormone levels at 18 months in intact control and 

intact gestationally exposed mice.

Glucose tolerance testing showed no statistically significant differences in baseline glucose 

or response to glucose challenge at 15–16 weeks or at 17 months. At 21 and 31 weeks of age, a 

significant increase in serum leptin and insulin levels was observed in intact mice exposed to 

0.01 and 0.1 mg PFOA/kg/day. No statistically significant difference was observed between the 

fat-to-lean ratio of intact control and intact gestationally exposed animals at 42 weeks of age. No 

significant difference was observed in food consumption between intact control and intact 

gestationally exposed animals at 42 weeks of age. Serum estradiol levels were not different 

between intact control and intact gestationally exposed animals at 18 months. 

Exposure to PFOA as an adult did not result in body weight differences among the groups at 

18 months of age. The body weight of intact mice gestationally exposed to 1 mg PFOA/kg/day 

was significantly increased (p<0.05) compared to adult mice exposed to 1 mg PFOA/kg/day. No 

other differences in body weight among the groups were observed. 

No significant differences among the groups were observed in survival during the 18-month 

study. At necropsy, abdominal white fat was significantly decreased (p<0.05) at 1 and 5 mg 

PFOA/kg/day in gestationally exposed intact mice compared to intact control mice. Interscapular 

brown fat was significantly increased (p<0.05) at 1 and 3 mg PFOA/kg/day in gestationally 

exposed intact mice and in gestationally exposed OVX mice at 1 mg PFOA/kg/day. Relative 

spleen weight was significantly decreased (p<0.05) at 3 mg PFOA/kg/day in gestationally 

exposed intact mice and at 1 and 5 mg PFOA/mg (p = 0.05-0.07) in gestationally exposed OVX 

mice. Relative liver weight was not different between the groups. No differences were observed 

at 18 months of age in tissue weight in mice exposed to PFOA as adults. At 1 mg PFOA/kg/day, 

white and brown fat weight was significantly increased in gestationally exposed intact mice 

compared to adult-exposed mice exposed to 1 mg PFOA/kg/day. 

The authors concluded that developmental exposure to low doses and high doses of PFOA 

resulted in different phenotypes in mice. At low doses, increased weight, increased serum 

insulin, and increased serum leptin were observed in adult mice. At high doses the animals 

displayed decreased weight in early and late life, decreased white fat, increased brown fat, and 

decreased spleen weight. Under the conditions of the study, the developmental LOAEL was 0.01 

mg PFOA/kg based on increased weight gain and increased serum insulin and leptin levels. No 

developmental NOAEL was established. The adult NOAEL was 5 mg PFOA/kg, and no LOAEL 

was established.

Perfluorooctanoic acid (PFOA) – May 2016 3-128
Adrenal Hormones. Thottassery et al. (1992) exposed intact or ADX male Sprague-Dawley rats 

to a single dose of 150 mg/kg PFOA in corn oil to determine the role of adrenal hormones on 

liver enlargement and peroxisomal proliferation. ADX rats were dosed 2 days after surgery with 

PFOA (ADX PFOA), CORT (ADX CORT), or both (ADX CORT PFOA). A group of intact and 

ADX rats received only the vehicle and served as controls. The animals were sacrificed 48 hours 

after dosing with PFOA or vehicle. Assays were conducted to determine DNA levels and 

changes in peroxisomal β-oxidation, catalase, and ornithine decarboxylase activities. An increase 

in ornithine decarboxylase activity has been associated with proliferation of many different cell 

types. An increase of ornithine decarboxylase in the livers of animals exposed to PFOA would 

suggest that the increased liver weight observed in PFOA-exposed animals was the result of 

hyperplasia. Ornithine decarboxylase was determined by measuring liberated CO2 from DL-[1- 14C] ornithine hydrochloride in all animals except those in the ADX CORT PFOA group. 

Relative liver weight in intact rats treated with PFOA was significantly increased compared 

to control (36%, p<0.05). Relative liver weight in rats in the ADX PFOA group was significantly 

increased compared to rats in the ADX vehicle group (16%, p<0.05). Relative liver weight in rats 

in the ADX CORT PFOA group was significantly increased compared to rats in the ADX CORT 

group (32%, p<0.05). Hepatic DNA levels were significantly decreased p<0.001) in intact rats 

treated with PFOA and in rats in the ADX CORT PFOA group.

Ornithine decarboxylase activity was significantly increased in the rats in the ADX PFOA 

group compared to rats in the ADX group (170.5 pmole CO2/hr/mg protein, versus 30.5 pmole 

CO2/hr/mg protein, p<0.001), but no different between the intact rats treated with PFOA and the 

intact rats treated with the vehicle.

PFOA increased whole liver peroxisomal β-oxidation activity by a similar amount and was 

not different among the groups. In intact rats and rats in the ADX CORT PFOA group, exposure 

to PFOA increased whole liver catalase activity, but exposure did not increase activity in the rats 

in the ADX PFOA group. Based on the results, the authors concluded that adrenal hormones 

were not required to induce peroxisomal β-oxidation activity in PFOA-exposed rats, but are 

required to increase catalase activity. They also concluded that the enlarged livers of PFOAexposed animals were the result of hypertrophy rather than hyperplasia based on decreased 

hepatic DNA content and lack of increased ornithine decarboxylase activity. 

3.3.4 Physiological or Mechanistic Studies

Gene Expression. Rosen et al. (2007) examined the gene expression profile in the lung and liver 

of mouse fetuses exposed to PFOA. Pregnant CD-1 mice were gavage-dosed with 0, 1, 3, 5, and 

10 mg PFOA/kg/day on GD 1–17. Dams were sacrificed on GD 18, and three fetuses per litter 

were processed for total RNA from portions of the liver and lung. Global gene expression was 

analyzed using Affymetrix gene chips.

A dose-related increase was observed in the number of genes altered by PFOA exposure in 

both the liver and lung. A greater number of genes in the liver were altered compared the number 

of genes altered in the lung. Analysis of the genes by canonical pathway or biological function 

showed that most of the altered genes in both the liver and lung were associated with lipid 

homeostasis. In the fetal lung, the two highest doses of PFOA altered genes associated with fatty 

acid catabolism. In the fetal liver, all doses of PFOA were associated with genes involved in fatty 

acid catabolism, lipid transport, cholesterol biosynthesis, bile acid biosynthesis, lipoprotein 

metabolism, steroid metabolism, retinol metabolism, inflammation, phospholipid metabolism, 

Perfluorooctanoic acid (PFOA) – May 2016 3-129
glucose metabolism, proteosome activation, and ketogenesis. Although PPARα is known to at 

least partly regulate the expression of genes for the pathways or biological functions involved in 

lipid homeostasis, PFOA might independently activate other nuclear receptors, influencing the 

metabolic responses observed. 

Rosen et al. (2008a) described the gene profiles in liver tissue from wild-type 129S1/SvlmJ 

mice (7–8 per group) and PPARα-null mice (129S4/SvJae-PPARαtm1Gonz/J, 6–8 per group) dosed 

for 7 days with 0, 1, and 3 mg PFOA/kg or 50 mg Wyeth 14,643, a PPARα agonist (Wolf et al. 

2008a). RNA was isolated from the tissues and gene expression analyzed using Applied 

Biosystems Mouse Genome Survey Microarrays. RT-PCR was used to evaluate selected genes.

In both wild-type and PPARα-null mice exposed to PFOA, the number of significant and 

fully annotated genes used to evaluate the data for relevance to canonical pathway or biological 

function were fewer at 1 mg/kg than at 3 mg/kg PFOA. However, 85% of the altered genes at 

1 mg/kg PFOA also were altered at 3 mg/kg PFOA. 

PPARα target genes including acyl-CoA oxidase 1 (Acox1), Me1, Slc27a1, Hsd17b4, Hadha, 

Hadhb, and Pdk4 were upregulated in PFOA- and Wyeth 14,643-treated wild-type mice, but not 

in PPARα-null mice. Pdk4 was downregulated in PPARα-null mice exposed to PFOA but not in 

PPARα-null mice exposed to Wyeth 14,643. Principal components analysis showed that genes 

activated in PFOA-treated PPARα-null mice were similar to those in PFOA-treated wild-type 

mice at 3 mg PFOA/kg, suggesting that many of the responses were not completely linked to 

PPARα. 

In wild-type PFOA- and Wyeth 14,643-treated mice, alterations were observed in genes 

associated with fatty acid metabolism (mostly upregulated), inflammatory response (mostly 

downregulated), cell cycle control (mostly upregulated), peroxisome biogenesis (mostly 

upregulated), and proteasome structure and organization (mostly upregulated). In genes 

associated with xenobiotic metabolism, the response was different between PFOA- and Wyeth 

14,643-treated wild-type mice. Many of the Cyp2 genes were upregulated by PFOA and 

downregulated by Wyeth 14,643. In PPARα-null PFOA-treated mice, genes associated with fatty 

acid metabolism, inflammation, xenobiotic metabolism, and cell cycle control were altered in a 

manner similar to the changes observed in PFOA-treated wild-type mice.

RT-PCR generally revealed good agreement with microarray analysis. However, expression 

of Ehhadh, a PPARα-regulated gene, was upregulated in PFOA-treated wild-type mice but not in 

PFOA-treated PPARα-null mice in microarray analysis. In contrast, expression of Ehhadh was 

upregulated in all PFOA-treated mice in RT-PCR analysis. The authors concluded that PFOA 

induces transcriptional changes mediated through PPARα activation, and it also alters gene 

expression independently of PPARα. They noted that PFOA had multiple modes of action and 

can function as a biologically active xenobiotic in the absence of PPARα. 

Rosen et al. (2008b) described the transcript profiles in the livers of adult mice exposed to 

PFOA. Tissues from several different studies were analyzed. The samples included liver tissue 

from: 

• male wild-type (strain 129S1/SvlmJ) and PPARα-null (strain 129S4/SvJae) mice dosed 

with 3 mg/kg/day PFOA for 7 days (from Wolf et al. 2008a); 

• male wild-type and PPARα-null mice (strain SV129/C57BL/6) gavage-dosed or fed diets 

containing Wyeth 14,643 (PPARα agonist); 

Perfluorooctanoic acid (PFOA) – May 2016 3-130
• female wild-type and CAR-null (strain C57BL/6x129Sv) gavage-dosed with CAR 

activators phenobarbital (PB) or 1,4-bis[2-(3,5-dichloropyridyloxy)] benzene 

(TCPOBOP); and 

• wild-type and Nrf-null ICR mice gavage-dosed with the Nrf activator dithiol-3-thione.

RNA was isolated from the tissues and gene expression was analyzed using Affymetrix full 

genome mouse chips. Rosetta Resolver software was used to identify significantly altered genes.

Exposure to 3 mg/kg PFOA for 7 days upregulated 641 genes and downregulated 451 genes 

in wild-type mice compared to 104 upregulated genes and 52 downregulated genes in PPARαnull mice. A total of 117 genes were regulated similarly in both strains, and 29 upregulated genes 

and 11 downregulated genes were unique to PPARα-null mice.

The gene expression profile of wild-type and PPARα-null mice exposed to PFOA for 7 days 

or Wyeth 14,643 for 12 hours, or 3 or 7 days were compared. Four groups of altered genes were 

identified based on their expression in wild-type and PPARα-null PFOA-exposed mice compared 

to genes from Wyeth 14,643-treated mice. The first group consisted of genes (397) regulated by 

both PFOA and Wyeth 14,643 in wild-type mice. They had a common direction and magnitude 

of change and were characterized as being involved in lipid homeostasis, inflammation, cell 

proliferation, or proteome maintenance genes. Group II consisted of genes in wild-type mice (51) 

regulated solely by PFOA; most were involved in amino acid metabolism. Of the 81 genes 

altered by PFOA exposure in PPARα-null mice (Group III), 62 had similar expression in wildtype mice and many were involved in lipid metabolism. Regulation of these genes also was 

observed in Wyeth 14,643 wild-type mice. Group IV genes (19) were altered by PFOA only in 

PPARα-null mice; most were xenobiotic metabolizing enzymes.

By comparing the gene expression patterns between PFOA and Wyeth 14,643, the authors 

concluded that:

• PPARα is required for a majority of the transcriptional changes observed in the mouse 

liver following PF also are regulated by other peroxisome proliferators in a PPARαdependent manner; and 

• PFOA impacts some PPARα-independent genes including ones involved in lipid 

homeostasis (upregulated), amino acid metabolism (downregulated), and xenobiotic 

metabolism (upregulated). 

The transcription profiles of PFOA exposed wild-type and PPARα-null mice were compared 

to the transcription profile of PB- or TCPOBOP-exposed wild-type and CAR-null mice and 

dithiol-3-thione-exposed wild-type and Nrf2-null mice to determine if PFOA activated CAR or 

Nrf2. A similar pattern was observed in the modified gene expression of PFOA-exposed PPARαnull mice and PB- (0.86 Pearson’s correlation) or TCPOBOP- (0.84 Pearson’s correlation) 

exposed wild-type mice, but no pattern was observed in gene expression of dithiol-3-thioneexposed mice (≤0.06 Pearson’s correlation) and PFOA-exposed PPARα-null mice. These results 

suggest that some genes altered by PFOA exposure in PPARα-null mice are regulated by CAR 

but not by Nrf2.

Bjork and Wallace (2009) examined the PPARα-dependent transcriptional activation 

potential of PFOA in rodent and human hepatic liver cells. Primary rat and human hepatocytes 

and HEPG2/C3A cells were incubated with 0, 5, 10, 20, 30, 50, 100, and 200 μmol PFOA for 

24 hours. Expression of Acox, Cyp4a1 (rat), Cyp4a11 (human), acyl-CoA thioesterase (Cte–rat, 

Acot1–human), and DNA damage inducible transcript (Ddit3) were determined by quantitative 

Perfluorooctanoic acid (PFOA) – May 2016 3-131
RT-PCR. These genes are inducible by peroxisome proliferators, except Ddit3, which is induced 

in the presence of direct or indirect DNA damage. Exposure to ≥ 100 μmol PFOA significantly 

increased Ddit3 mRNA expression in primary rat hepatocytes. At the highest dose, Ddit3 was 

significantly increased in human hepatocytes and HepG2/C3A cells. Expression of Acox was 

significantly induced by 5, 10, 20, and 30 μmol PFOA, and Cte/Acot1 was significantly induced 

at ≥20 μmol PFOA in rat hepatocytes only. Expression of Cyp4a1/11 was significantly induced 

in rat hepatocytes at 5–-50 μmol and in human hepatocytes at 20–50 μmol. The authors 

concluded that induction of peroxisome-related fatty acid oxidation gene expression is not 

observed in primary human liver cells or in transformed human liver cells in vitro.

Nakamura et al. (2009) investigated the differences in PFOA response between mice and 

humans using a humanized PPARα transgenic mouse line (hPPARα). The study design and 

whole animal toxicity data are described in section 3.2.2. Male 8-week-old wild-type (mPPARα) 

mice, PPARα-null mice, and hPPARα mice were gavage-dosed with 0, 0.1, and 0.3 mg/kg PFOA

(n = 4–6 per group) for 2 weeks and sacrificed 18–20 hours following the last dose. Livers were 

collected and analyzed for mRNA (RT-QPCR) and protein levels (western blot analysis) of 

PPARα and related genes (retinoid X receptor alpha [RXRα], peroxisomal bifunctional protein 

[PH], peroxisomal thiolase [PT], very long chain acyl-CoA dehydrogenase [VLCAD], medium 

chain acyl-CoA dehydrogenase [MCAD], and cytochrome P450 4a10 [CYP4A10]). RXRα forms 

a heterodimer with PPARα to control transcription of genes affecting lipid metabolism. 

CYP4A10 also plays a role in lipid metabolism. Treatment with peroxisome proliferators caused 

an increase in both PH and PT. MCAD and VLCAD are mitochondrial fatty acid metabolizing 

enzymes whose gene expression is mediated by PPARα (Aoyama et al. 1998). The results of 

mRNA expression impacted by PFOA exposure are shown in Table 3-38. 

Treatment with PFOA did not alter mRNA expression or protein expression of PPARα, 

RXRα, or MCAD in mPPARα mice. At 0.1 mg/kg PFOA, mRNA expression of CYP4A10 was 

significantly increased (p<0.05) in mPPARα mice compared to control mPPARα mice.

Treatment with 0.3 mg/kg PFOA resulted in significantly increased (p<0.05) mRNA expression 

of CYP4A10 and mRNA and protein expression of PH, PT, and VLCAD in mPPARα mice when 

compared to control mPPARα mice.

Table 3-38. mRNA Expression of Hepatic PPARα and Related Genes

mPPARα PPARα-null hPPARα

0 mg/kg

0.1 

mg/kg

0.3 

mg/kg 0 mg/kg

0.1 

mg/kg

0.3 

mg/kg 0 mg/kg

0.1 

mg/kg

0.3 

mg/kg

PPARα - - - NA NA NA -*↑ -*↑ -*↑

RXRα - - - - - - - - -

PH - - +↑ - -*↓ -*↓ - -*↓ -*↓

PT - - +↑ - -*↓ -*↓ -*↑ - -

VLCAD - - +↑ -*↓ -*↓ -*↓ - -*↓ -*↓

MCAD - - - - -*↓ - - -*↓ +↑

CYP4A10 - +↑ +↑ - -*↓ -*↓ -*↑ -*↓ -*↓

Source: Nakamura et al. 2009

Notes:

- Not different from respective control. 

+ Significantly different from respective control. 

* Significantly different from mPPARα mice treated with same concentration. 

↓ Decreased expression relative to respective control or mPPARα mice at same concentration. 

↑ Increased expression relative to respective control or mPPARα mice at same concentration. 

Perfluorooctanoic acid (PFOA) – May 2016 3-132
As expected, mRNA and protein expression of PPARα was absent in PPARα-null mice.

Treatment with 0.1 and 0.3 mg/kg PFOA did not alter mRNA or protein expression for any genes 

investigated compared to control PPARα-null mice. VLCAD mRNA expression and PT protein 

expression in control PPARα-null mice was significantly decreased (p<0.05) compared to 

mPPARα control mice. VLCAD mRNA and protein expression of PFOA treated PPARα-null 

mice was significantly decreased (p<0.05) compared to mPPARα mice treated with the same 

doses. Following treatment with 0.1 mg/kg PFOA, MCAD mRNA expression was decreased 

(p<0.05) compared to mPPARα mice treated with the same dose. When compared to mPPARα 

mice treated with the same dose, mRNA and protein expression of PH and PT was significantly 

decreased (p<0.05) in PPARα-null mice, as was CYP4A10 mRNA expression. 

Treatment with PFOA did not alter mRNA or protein expression of PPARα, RXRα, PH, PT, 

or VLCAD in hPPARα mice compared with their respective controls. Expression of CYP4A10 

mRNA also was not altered by PFOA treatment. MCAD mRNA and protein expression were 

significantly increased (p<0.5) in hPPARα mice treated with 0.3 mg/kg PFOA compared to 

hPPARα control mice. Expression of PPARα mRNA and protein levels were significantly higher 

(p<0.05) in all hPPARα mice than in mPPARα mice given the same concentration of PFOA. 

Treatment of hPPARα mice with 0.1 and 0.3 mg/kg PFOA caused a decrease (p<0.05) in mRNA 

expression of PH, VLCAD, and CYP4A10 compared to mPPARα mice at the same dose. Only 

hPPARα mice treated with 0.3 mg/kg PFOA had decreased protein expression of PH and 

VLCAD compared to mPPARα mice given the same treatment. An important finding from this 

study was the significant downregulation of some genes in PPARα-null and hPPARα mice that 

are significantly upregulated by PPARα in the control animals. In the animals with the 

humanized PPARα gene or no PPARα gene, there was a response, but the response was the 

opposite of what occurred with normal mouse PPARα activation. In the null and humanized 

mice, the significantly decreased alterations in gene expression occurred at 0.1 mg/kg/day; this 

dose level had no change in expression for all but one gene in the normal mice and increased 

expression, rather than decreased expression, at 0.3 mg/kg/day (see Table 3-38). 

Treatment with 0.3 mg/kg PFOA caused activation of PPARα in mouse, but not in 

humanized PPARα mice. The results suggest that the functional activation of human PPARα 

could be weaker than that of mice as expression of human PPARα in mice was greater than the 

expression of mouse PPARα. Higher concentrations of PFOA might be needed to cause 

activation of human PPARα in hPPARα mice.

To further evaluate the developmental effects potentially mediated by PPARα, groups of 

female wild-type, PPARα-null, and PPARα-humanized mice were given 0 and 3 mg PFOA/kg 

on GDs 1–17 by oral gavage (Albrecht et al. 2013). The study design and developmental toxicity 

data are described in section 3.2.5. Females were either sacrificed on GD 18 (n = 5–8 per group) 

or allowed to give birth and then sacrificed, along with their litters (n = 8–14), on PND 20. 

Livers from dams, fetuses, and pups were collected for measurement of mRNAs encoding the 

PPARα target genes Cyp4a10 and Acox1, the CAR target gene (Cyp2b10), and the PXR target 

gene (Cyp3a11). 

On GD 18, maternal liver samples from treated groups showed increased expression of 

Acox1 in wild-type mice and Cyp4a10 in wild-type and humanized mice. Expression of 

Cyp2b10 and Cyp3a11 were increased following PFOA administration in all three genotypes. On 

PND 20, maternal liver samples from treated groups showed increased expression of Acox1 in 

wild-type mice; expression of Cyp2b10 was unchanged in all groups; and expression of Cyp3a11 

was increased following PFOA administration in all three genotypes. 

Perfluorooctanoic acid (PFOA) – May 2016 3-133
For fetuses on GD 18, liver samples from treated groups showed increased expression of 

Acox1 and Cyp4a10 in wild-type and humanized mice. Expression of Cyp2b10 was unchanged 

following maternal PFOA administration in all three genotypes, while expression of Cyp3a11 

was increased in humanized fetal liver. On PND 20, pup liver samples from treated dams showed 

increased expression of Acox1 and Cyp4a10 in wild-type mice; expression of Cyp2b10 was 

increased in all genotypes; and expression of Cyp3a11 was increased following maternal PFOA 

administration in wild-type and humanized pups. Thus, expression of PPARα target genes that 

modulate lipid metabolism was increased in both wild-type and humanized mice coincident with 

increased liver weight and microscopic lesions; however, the neonatal mortality was observed 

only in wild-type offspring (Albrecht et al. 2013). 

Walters et al. (2009) examined the impact of PFOA on mitochondrial biogenesis and gene 

transcription in adult male Sprague-Dawley rats orally dosed with 0 or 30 mg/kg PFOA for 

28 days. At sacrifice, a portion of the midlobe region of the livers was collected. Liver DNA and 

RNA were isolated for RT-PCR of genes in the peroxisome proliferator-activated receptor 

gamma coactivator 1α- (Pgc-1α-) mediated pathway of mitochondrial biogenesis: Pgc-1α, 

estrogen-related receptor α (Errα), nuclear respiratory factor 1 (Nrf1) and Nrf2, transcription 

factor A (Tfam), cytochrome c oxidase subunit II and IV (Cox II and Cox IV), NADH 

dehydrogenase 2 (Nd2), and NADH dehydrogenase iron-sulfur protein 8 (Ndufs8). In 

mitochondrial biogenesis, Pgc-1α and Errα increase expression of the transcription factors Nrf1 

and Nrf2. The Nrf transcription factors promote expression of Tfam, which is required for 

mitochondrial DNA replication and transcription. Within the mitochondrial membrane, oxidative 

phosphorylation proteins (Cox II and IV, Nds, and Ndufs8) catalyze the transfer of electrons 

and/or pump protons from the matrix to the intermembrane space. Western blotting was used to 

analyze protein expression of Pgc-1α, Tfam, Cox II, and Cox IV.

Mitochondrial DNA in rats treated with PFOA was significantly increased (p<0.05) 

compared to control rats. In PFOA-treated rats, the expression of Pgc-1α, Errα, Nrf1, Nrf2, and 

Tfam was significantly increased 1.3–2.2-fold (p<0.05), and expression of Cox II, Cox IV, Nd2, 

and Ndufs8 was significantly increased 2–9-fold (p<0.05) compared to controls. Protein 

expression of Pgc-1α was increased, and expression of Cox II and Cox IV were decreased in 

PFOA-treated rats. Protein expression of Tfam was not affected by treatment with PFOA. The 

results suggested that PFOA induced mitochondrial biogenesis at the transcriptional level by 

activation of the Pgc-1α pathway, confirming the potential for effects on mitochondria but not 

clarifying whether those effects are in some way linked to PPARα activation. 

Elcombe et al. (2010) examined the expression of some cytochrome P450 isoforms in the 

livers of male Sprague-Dawley rats fed diets containing 300 ppm PFOA or 50 ppm Wyeth 

14,643 for 1, 7, or 28 days. The isoforms included those involved in activation of PPARα 

(CYP4A1), CAR (CYP2B1/2), and PXR (CYP3A1). All three isoforms were induced by PFOA 

exposure. CYP2B1/2 and CYP4A1 were induced after 1 day of exposure to PFOA. CYP3A1 was 

induced in all PFOA-exposed rats after 7 days of exposure. Treatment with Wyeth 14,643 caused 

the induction of CYP4A1 only. 

PPAR Activation. Takacs and Abbott (2007) evaluated the potential for PFOA to activate 

PPARs, using a transient transfection cell assay. Cos-1 cells, derived from the kidney cells of the 

African green monkey, were transfected with mouse or human PPARα, PPARβ/δ, or PPARγ 

reporter plasmids and exposed to 0.5–100 μmol PFOA or 0.5–100 μmol PFOA and MK-886 

(PPARα antagonist) or GW9662 (PPARγ antagonist). An antagonist for PPARβ/δ was not 

available. The three types, PPARα, β/δ, and γ, are encoded by different genes, expressed in many 

Perfluorooctanoic acid (PFOA) – May 2016 3-134
tissues, and have specific roles during development as well as in the adult. The results are shown 

in Table 3-39. PFOA activated PPARα in a dose-dependent manner with a significant increase in 

activity observed at 10, 20, 30, and 40 μmol for the mouse receptor and 30 and 40 μmol for the 

human receptor compared to the negative control. The presence of the PPARα antagonist 

MK-886 prevented the activity increase resulting from PFOA exposure alone in mouse and 

human PPARα constructs.

Table 3-39. Activation of Mouse and Human PPAR by PFOA

PPARα PPARβ/δ PPARγ

PFOA 

(μm) Mouse Human

PFOA 

(μm) Mouse Human

PFOA 

(μm) Mouse Human

0 - - 0 - - 0 - -

0.5 - - 10 - - 1 - -

1 - - 15 - - 5 - -

3 - - 20 - - 10 - -

5 - - 30 - - 20 - -

10 + - 40 + - 30 - -

15 - - 50 + - 40 - -

20 + - 60 + - 50 - -

30 + + 70 + - 75 - -

40 + + 80 + - 100 - -

Source: Takacs and Abbott 2007

Notes:

+ Significant increase in activity between treated and control. 

- No difference in activity between treated and control. 

Activity of mouse PPARβ/δ was significantly increased after exposure to 40–80 μmol PFOA 

compared to the negative control. Activity of human PPARβ/δ was not increased by PFOA 

exposure. Activity of mouse and human PPARγ were not increased by exposure to PFOA. PFOA 

was found to activate mouse and human PPARα and mouse PPAR β/δ under the conditions in 

this study. 

Biomarkers for Peroxisome Proliferation. Pastoor et al. (1987) dosed male Crl:CD (SD) BR 

rats for 1, 3, and 7 days with 0 or 50 mg PFOA/kg/day. Hepatic DNA content, cytochrome P450 

content, UDP-glucuronyltransferase, glutathione S-transferase, benzphetamine N-demethylase 

activity (marker for smooth ER proliferation), and ethoxyresorufin O-deethylase activity (marker 

for cytochrome P450 induction via the aryl hydrocarbon receptor) were measured from rats 

dosed 1 and 3 days. Liver microsomes were prepared from rats dosed for 3 days for CAT and 

CPT activity assays. CAT served as a marker for peroxisome proliferation and CPT was a 

marker for mitochondrial proliferation. Incorporation of [14C]acetate into hepatic lipids was used 

to determine the effect of PFOA on hepatic lipid metabolism. Plasma TC and triacylglycerides 

was determined from rats dosed for 7 days.

Hepatic DNA content was not increased in treated rats when compared to content in control 

rats. Cytochrome P450 was significantly increased (p<0.05) and ethoxyresorufin O-deethylase 

activity was significantly decreased (p<0.05) after treatment for 1 and 3 days. Benzphetamine 

N-demethylase activity was significantly increased (p<0.05) after treatment with PFOA for 

3 days. CAT activity increased 12-fold (p<0.05) and CPT increased twofold (p<0.05) after a 

3-day treatment with 50 mg PFOA/kg. No differences were observed among the groups for the 

Perfluorooctanoic acid (PFOA) – May 2016 3-135
other enzymes. No differences were observed between rats treated for 7 days and control rats in 

plasma TC or triacylglycerol. Although a significant increase (p<0.05) was observed for 

[

14C]acetate incorporation into triacylglycerols, cholesteryl esters, and polar lipids, there was no 

difference in the distribution of the incorporated label between control and treated rats. The 

authors concluded that the lack of increased DNA content, proliferation of smooth ER, and 

peroxisome proliferation pointed to increased liver weight due to hepatocyte hypertrophy. 

Gap Junction Intercellular Communication. Upham et al. (1998, 2009) examined the effects 

of perfluorinated fatty acids on gap junction intercellular communication (GJIC) in male Fischer 

344 rats fed diets containing 0 or 0.02% PFOA (intake 37.9 mg/kg/day) for 1 week and in 

WB-F344 rat liver epithelial cells. The chain lengths of the perfluorinated fatty acids ranged 

from 2–10, 16, and 18 carbons. Liver weight in the rats fed diets containing 0.02% PFOA was 

significantly increased compared to control rat liver weight. No differences were observed in 

serum AST, ALT, and ALP. PFOA significantly inhibited GJIC in the livers of rats after 

treatment for 1 week. In WB-F344 cells, GJIC was inhibited by perfluorinated fatty acids with 

7–10 carbons within 15 mins of incubation. The inhibition was reversible with full recovery 

occurring within 30 mins of PFOA removal from media. Extracellular receptor kinase was 

activated by PFOA within 5 mins of incubation in the cells. Preincubation of cells with the 

phosphatidylcholine-specific phospholipase C inhibitor D609 partially prevented GJIC inhibition 

by PFOA. The authors concluded that PFOA, having an 8-carbon chain, inhibited GJIC by 

activation of extracellular receptor kinase and phosphatidylcholine-specific phospholipase C, but 

noted that other mechanisms might be involved. 

Production of ROS. Takagi et al. (1991) fed male Fischer 344 rats diets containing 0, 10, and 

20 mg PFOA/kg for 2 weeks to determine the formation of 8-OH-dG (marker of oxidative DNA 

damage). Livers and kidneys were removed at necropsy and DNA was isolated from each organ 

and analyzed. The relative liver and kidney weights were significantly increased (p<0.05) in the 

treated rats compared to the control. A significant increase in 8-OH-dG liver levels was observed 

at ≥ 10 mg PFOA/kg. There were no significant differences in 8-OH-dG kidney levels between 

PFOA-treated and control rats. The authors concluded that PFOA could cause organ-specific 

oxidative DNA damage. 

Hu and Hu (2009) exposed human hepatoma cells, HepG2, to PFOA to evaluate cytotoxic 

effects. Cells also were exposed to a mixture of PFOA and PFOS to determine antagonistic or 

synergic effects. The cells were exposed to 0, 50, 100, 150, and 200 μmol PFOA or to 0, 50, 150, 

and 200 μmol each of PFOA and PFOS. A group of cells also were exposed to 0, 50, 100, 150, 

and 200 μmol PFOS. The cells were cultured for 24, 48, and 72 hours. Cell viability, apoptosis, 

ROS, mitochondrial membrane potential, antioxidant enzymes, glutathione content, and 

differential expression of apoptosis gene regulators p53, Bax, Bcl-2, caspace-3, and caspace-9 

genes were evaluated.

Exposure to PFOA or PFOS caused a dose-dependent decrease in viability of HepG2 cells. A 

nonsignificant dose-dependent increase in apoptosis was observed in the cells cultured with 

PFOA. However, the combination of PFOA and PFOS showed a significant dose-dependent 

increase (p<0.05) in apoptosis. Intracellular ROS were significantly increased (p<0.05) in cells 

cultured with 100, 150, and 200 μmol PFOA or PFOS. HepG2 cells exposed to the mixture of 

100 and 200 μmol PFOA and PFOS exhibited a decline in fluorescence intensity in the 

mitochondrial membrane potential assay, indicating that mitochondrial pathways were involved 

in the apoptosis observed. Exposure to 100 μmol PFOA significantly decreased (p<0.05) 

glutathione concentration and glutathione peroxidase activity; and 150 μmol PFOA significantly 

Perfluorooctanoic acid (PFOA) – May 2016 3-136
increased (p<0.05) the activities of SOD, catalase, and glutathione reductase, and significantly 

decreased (p<0.05) glutathione peroxidase activity, and glutathione concentration in HepG2 

cells. The trend was the same at 200 μmol PFOA, with the exception of GST activity being 

significantly decreased (p<0.05). 

Exposure to PFOA did not change p53, Bax, or caspace-3 expression in HepG2 cells. 

Expression of Bcl-2 was downregulated and caspace-9 was upregulated in a dose-dependent 

manner in HepG2 cells following exposure to 50–200 μ/Mol PFOA. The authors proposed that 

PFOA and PFOS induced cell apoptosis by overwhelming the homeostasis of antioxidative 

systems, increasing ROS, impacting mitochondria, and changing gene expression of apoptosis 

gene regulators.

Eriksen et al. (2010) examined ability of PFOA to generate ROS and induce oxidative DNA 

damage in human HepG2 cells. Cell were incubated with 0, 0.4, 4, 40, 200, 400, 1,000, and 

2,000 μmol PFOA and 2’,7’-dichlorofluorescein diacetate. Hydrogen peroxide, H2O2, was used 

as a positive control. A fluorescence spectrophotometer was used to measure ROS production 

every 15 mins during the 3-hour incubation period in all cultures. The comet assay was used to 

measure DNA damage in cells exposed to 0, 100, and 400 μmol PFOA for 24 hours. 

Cytotoxicity was determined by measuring the level of lactate dehydrogenase activity in the cell 

medium. Exposure to PFOA caused a dose-independent increase (all doses p<0.05) in ROS 

production in HepG2 cells. Compared to ROS production in negative control cells, PFOA 

induced a 1.52-fold increase in production. There was no difference in oxidative DNA damage 

and lactase dehydrogenase activity between PFOA-treated cells and negative control cells. The 

authors concluded that oxidative stress and DNA damage were probably not relevant to potential 

adverse effects of PFOA.

Protein Binding. The ability of PFOA to bind to serum proteins for distribution is discussed in 

section 2.2. PFAS protein binding also can impact cellular function in cases in which the proteins 

in question are transporters (serum albumin and fatty acid binding protein) or enzymes (lysine 

decarboxylase) as well as membrane receptors (e.g., members of the PPAR family) and thyroid 

hormone receptors. The mechanistic studies of the nuclear PPARα membrane receptors are 

described in section 3.3.4. 

Ren et al. (2015) examined the relative binding affinities of 16 PFASs for the human thyroid 

hormone receptor’s α ligand binding domain (TRα-LBD) using a fluorescence competitive 

binding assay. Solutions of 1 μmol TRα-LBD were prepared in DMSO. Changes in TRα-LBD 

tryptophan fluorescence after binding to 10-μmol T3 in the absence or presence of the PFAS was 

used to determine the binding properties of the PFAS. IC50 values were calculated by linear 

extrapolation between two responses located in the vicinity of a 50% inhibition level. All the 

PFASs had a lower affinity for the receptor than T3. Affinity of PFOA was less than that for 

PFDA, PFUnA, PFNA, and PFOS.

ToxCast Assay Results. The Toxicity Forecaster (ToxCast) database is a large high-throughput 

screening compilation of public in vitro and in vivo assays on over 9,000 chemicals (USEPA 

2015). PFOA was tested in 1,084 assays and was active in 40 (USEPA 2015). Assays with less 

than 50% efficacy reported or overfitting issues are not included in the summary of results that 

follows. 

Perfluorooctanoic acid (PFOA) – May 2016 3-137
Three of the acceptable ToxCast active cytotoxicity assays evaluated the impact of PFOA. 

All three of these assays are based on one cell type. If there was no cytotoxicity reported for a 

specific cell type, the AC50 (the minimum concentration with 50% cytotoxicity activity) was 

used for comparison when reporting the ToxCast results. The lowest recorded AC50 (109 μmol) 

measured the degradation of microtubules in liver cells at 109 μmol and the highest recorded 

(123 μmol) measured general cytotoxicity in liver cells.

PFOA activated two of the 21 estrogen related assays in ToxCast; both were ESR1-related. 

Estrogen and its receptors are essential for sexual development and reproductive function, but 

also play a role in other tissues such as bone. PFOA induced estrogen response element and 

inhibited ESR1 at concentrations lower than their AC50 values with concentrations of 33.8 μmol

and 47.4 μmol, respectively. This implies that PFOA could have some estrogenic potential; 

however, due to the small fraction of estrogenic assays activated (10%), any activity is likely 

weak.

PFOA activated PPARs, PXR, CAR, and retinoic acid receptor (RAR) assays within the 

ToxCast program. From the PPAR assays, PFOA induced the DNA sequences for PPARα, 

PPARγ, and the peroxisome proliferator hormone response element (PPRE) and antagonized the 

PPARγ receptor. The only PPAR assay AC50 value that was above the cell-specific AC50 was 

that for PPARγ antagonism at 5.91 μmol. However, it is possible that cytotoxicity occurs due to 

PPAR induction, or that PPAR cytotoxicity leads to PPAR induction confounding interpretation 

of the outcome. PFOA induced DNA sequences for PXR (AC50 9.42 μmol) at a concentration 

lower than the cell-specific AC50. CAR and RAR alpha antagonism also was observed, but the 

concentrations of 17.57 μmol and 28.45 μmol, respectively, were not below the cell-specific 

cytotoxicity value. PPAR, PXR, CAR, and RAR pathways are all nuclear receptors that can form 

heterodimers with one another to induce translation of various genes. Some of these genes are 

important for development, reproduction, and waste degradation, and could play a role in PFOAinduced cancer.

The ToxCast program examined Cytochrome P450 (CYP) activation associated with PFOA 

exposure. Although PFOA is not metabolically active, it was found to activate four CYPs: 

CYP2C18, CYP2C19, and CYP2C9 in human cells and CYP2C11 in rat cells. All of the CYP 

assays were activated at concentrations lower than the lowest AC50 (109 μmol) but lacked cellspecific AC50s. The CYP2C class of CYPs is involved in the metabolism of xenobiotics such as 

the following drugs: the antiseizure medication diazepam, beta blocker propranolol, and selective 

serotonergic reuptake inhibitor citalopram. Though there is no evidence of metabolism of PFOA 

by these CYPs, it is possible that it acts as a competitive or allosteric inhibitor for known 

substrates of the CYPs activated. This coupled with PFOA’s high affinity for binding to albumin 

could significantly alter the PKs of various pharmaceutical bound to serum albumin, thus 

potentially playing a role in increasing systemic toxicity of some pharmaceuticals by increasing 

the free serum concentration.

PFOA failed to cause toxicity in the in vivo fish model for neurological and developmental 

toxicity. This is important because PFOA induces developmental toxicity in mice and rats 

in vivo. 

Perfluorooctanoic acid (PFOA) – May 2016 3-138
3.3.5 Structure-Activity Relationship

Bjork and Wallace (2009) compared the PPARα-dependent transcriptional activation 

potential of linear perfluorocarboxylic and sulfonic acids in rodent and human hepatic liver cells. 

The PFAAs tested included perfluorinated carboxylic acids with carbon chain lengths of 2–8 and 

perfluorinated sulfonic acids with chain lengths of 4–8. Primary rat and human hepatocytes and 

HEPG2/C3A cells were incubated with 0 and 25 μmol perfluorinated compounds for 24 hours. 

Expression of Acox, Cyp4a1 (rat), Cyp4a11 (human), Cte/Acot1, and Ddit3 (GADD153) 

transcripts were determined by quantitative RT-PCR. All the genes are inducible by peroxisome 

proliferators except Ddit3, which is induced in the presence of direct or indirect DNA damage.

Perfluorinated compounds induced mRNA expression of either Acox or Cte/Acot1 only in rat 

hepatocytes, and the degree of stimulation of gene expression increased with increasing carbon 

number. The Cyp4a11 gene was not expressed or stimulated by any of the PFAAs in 

HepG2/C3A cells. However, this gene expression was stimulated by perfluorinated exposure in 

both rat and human hepatocytes with the perfluorocarboxylates showing a chain-lengthdependent structure activity relationship. The study results suggest that the PPARα-related 

changes in gene expression induced by perfluorinated compounds in primary rat hepatocytes are

directly related to the carbon chain length and appear to be stronger for the carboxylic acids 

(i.e., PFOA) than the sulfonates (i.e., PFOS). There was no induction in expression of Acox and 

Cte/Acot 1 in either primary or transformed human liver cells in culture. The authors suggested 

that the PPARα mediated peroxisome proliferation observed in rodent liver might not be relevant 

as an indicator to human risk. 

Wolf et al. (2008b) tested PFAAs, including PFOA, to determine if mouse and human 

PPARα activity could be induced in transiently transfected COS-1 cell assays. COS-1 cells were 

transfected with either a mouse or human PPAR-α receptor-luciferase reporter plasmid and, after 

24 hours, were exposed to either negative controls (water or 0.1% DMSO), a positive control 

(Wyeth 14,643), or PFOA at 0.5–100 μmol. Other concentrations of PFAAs were used but not 

provided in this report. At the end of 24 hours of exposure, the luciferase activity was measured.

The positive and negative controls had the expected results. A lowest observed effect 

concentration (LOEC) and no observed effect concentration (NOEC) were determined for each 

PFAA. In the study, the mouse PPARα was more responsive than the human. Also, carboxylates 

induced higher mouse and human PPARα activity than the sulfonates. In this study, the NOEC 

for PFOA was 0.5 μmol in the mouse and 5 μmol in humans; the LOEC was 1 μmol

(0.43 μg/mL) in the mouse and 10 μmol (4.3 μg/mL) in humans. 

A similar study included additional PFAAs (Wolf et al. 2012). Transfected cells were 

incubated with PFAAs at concentrations of 0.5 to 100 μmol, vehicle (water or 0.1% DMSO as 

negative control) or with 10 μmol Wyeth 14,643 (positive control) on each plate. Assays were 

performed with three identical plates per compound per species with nine concentrations per 

plate and eight wells per concentration. Cell viability was assessed using the Cell Titer Blue cell 

viability kit and read in a fluorometer. The positive and negative controls had the expected 

results. All cells transfected with either human and mouse PPARα responded to the PFAAs.

Again, the carboxylates were stronger inducers than the sulfonates, and the mouse PPARα was

more reactive than the human PPARα. The study also provided the C20max values for each PFAA 

(the concentration at which the PFAA produced 20% of the maximal response elicited by the 

most active PFAA). For PFOA, this was 6 μmol in mouse PPARα and 7 μmol in human PPARα.

For comparison, PFOS was 94 μmol and 262 μmol, respectively. 

Perfluorooctanoic acid (PFOA) – May 2016 3-139
3.4 Hazard Characterization

3.4.1 Synthesis and Evaluation of Major Noncancer Effects 

Serum Lipids. Because of the structural similarities between linear perfluorinated acids and the 

short- and medium-chain fatty acids, the potential for these chemicals to cause elevated serum 

lipids has been an area of considerable interest. High levels of serum lipids (TC and LDL) are 

risk factors for cardiovascular disease in humans, including IHD, a condition in which blood 

flow to the heart is decreased through the development of atherosclerotic plaque or clots in the 

cardiac arteries.

The association between PFOA and serum lipids has been examined in several studies in 

different populations. Cross-sectional and longitudinal studies in occupational settings (Costa et 

al. 2009; Olsen et al. 2000, 2003; Olsen and Zobel 2007; Sakr et al. 2007a, 2007b; Steenland et 

al. 2015) and in the high-exposure community (the C8 Health Project study population) (FitzSimon et al. 2013; Frisbee et al. 2010; Steenland et al. 2009; Winquist and Steenland 2014a) 

generally observed positive associations between serum PFOA and TC in adults and children 

(aged 1–< 18 yrs); most of these effect estimates were statistically significant. Although 

exceptions to this pattern are present (i.e., some of the analyses examining incidence of selfreported high cholesterol based on medication use in Winquist and Steenland 2014a and in 

Steenland et al. 2015), the results are relatively consistent and robust. Similar associations were 

seen in analyses of LDL, but were not seen with HDL. The range of exposure in occupational 

studies is large (means varying between 0.4 and > 12 μg/mL), and the mean serum levels in the 

C8 population studies were around 0.08 μg/mL. Positive associations between serum PFOA and 

TC (i.e., increasing lipid level with increasing PFOA) were observed in most of the general 

population studies at mean exposure levels of 0.002–0.007 μg/mL (Eriksen et al. 2013; Fisher et 

al. 2013; Geiger et al. 2014a; Nelson et al. 2010; Starling et al. 2014). The interpretation of these 

general population results is limited, however, by the moderately strong correlations (Spearman r 

> 0.6) and similarity in results seen for PFOS and PFOA. The most recent update of disease 

incidence in workers in the C8 Health Project study population identified 35 cases of 

nonhepatitis liver disease (with medical validation) (Steenland et al. 2015); no association was 

seen with cumulative exposure when analyzed without a lag (HRs by quartile 1.0, 0.58, 1.43, 

1.20; trend p = 0.86 for log cumulative exposure), but there was a possible trend in the analysis 

using a 10-year lag (HRs by quartile 1.0, 1.46, 2.13, and 2.02; trend p = 0.40). 

Cholesterol and/or triglycerides were monitored in only a few of the animal studies, which 

did not all measure concurrent serum PFOA levels. Information on serum lipids from animal 

studies has received less attention than in the human population because of the fact that 

decreases in triglycerides, cholesterol, and lipoprotein complexes are an expected consequence 

of PPARα activation in rodents. The PPARα response in animals tends to lower rather than raise 

serum cholesterol and associated lipid levels. Peroxisomes are subcellular organelles that 

increase beta oxidation of long-chain fatty acids using a beta oxidation pathway that is not linked 

to adenosine triphosphate (ATP) production and release the shortened fatty acids to the cytosol 

as an endproduct for export in VLDLs or hepatic ATP-production via mitochondrial beta 

oxidation (Garrett and Grisham 1999). PPARα activation also stimulates metabolic changes that 

lower hepatic cholesterol. The effects of human PPARα activation are much less pronounced 

than those in rats and mice.

Perfluorooctanoic acid (PFOA) – May 2016 3-140
Nakamura et al. (2009) and Minata et al. (2010) examined the lipid endpoints relative to the 

mouse strain’s PPARα status and PFOA exposure. Nakamura et al. (2009) found that mice with a 

normal PPARα receptor had significantly increased levels of cholesterol and triglycerides in liver 

but not plasma at a LOAEL of 0.3 mg/kg/day. However, there were no differences in serum or 

liver cholesterol or triglycerides between PFOA-treated mice with a humanized PPARα receptor 

or PPARα-null mice (NOAEL= 0.3 mg/kg/day) and their respective controls. The study by 

Minata et al. (2010) used higher doses than Nakamura et al. (2009) and found that TC was 

significantly decreased (LOAEL= 10.8 mg/kg/day; whole blood 47 μg/mL) and total 

triglycerides significantly increased (LOAEL= 5.4 mg/kg/day; whole blood 21 μg/mL) in wildtype mice. In the PPARα-null mice, the TC was significantly decreased for the 5.4- and 

10.8-mg/kg/day doses but significantly increased for a 21.6-mg/kg/day dose while total 

triglycerides were significantly increased at all doses; these doses corresponded to whole blood 

PFOA levels of 13, 36, and 71 μg/mL, respectively. Rosen et al. (2007) found that PFOA 

activated genes for fatty acid catabolism, cholesterol biosynthesis; bile acid biosynthesis; and 

lipoprotein, steroid, and glucose metabolism in fetal livers. When comparing the response in 

PPARα wild-type to null mice (Rosen et al. 2008b), 62 of 81 activated genes were the same for 

both strains and were ones involved with lipid metabolism. 

Martin et al. (2007) identified a 45–72% decrease in serum cholesterol after treatment of 

male Sprague-Dawley rats with 20 mg PFOA/kg/day for up to 5 days (serum PFOA 245 μg/mL 

after 3 days), and Loveless et al. (2008) reported decreased TC, HDL, and non-HDL in male CD 

rats after doses of 0.3 and 1 mg/kg/day for 28 days. Triglycerides were decreased in the rats at 

≥0.3 mg/kg/day. De Witt et al. (2009) found a dose-dependent decrease in triglyceride levels in 

female C57BL/6N mice exposed to 0, 7.5, and 15 mg PFOA/kg bw in drinking water for 

10 days. In male CD-1 mice, TC, HDL, and triglycerides were decreased at 10 and 30 mg/kg/day 

(Loveless et al. 2008). In pregnant female ICR mice, triglyceride, TC, and free fatty acids were 

significantly decreased at 10 mg/kg (Yahia et al. 2010). Elcombe et al. (2010) found a significant 

decrease in cholesterol in male Sprague-Dawley rats following a 7- or 28-day exposure to 

300 ppm PFOA in the diet with a resulting serum level of 252 μg/mL at 28 days. Accordingly, 

there is not a high degree of concordance between the lipidemic effects of PFOA as noted in 

human epidemiology studies and those seen in animals.

Filgo et al. (2015) found a trend for increased liver Ito (fat) cell atrophy and lesion severity 

across the doses in CD-1 and SV-129 mice at 18 months. PFOA exposure occurred only through 

the dam during gestation and lactation in this study. This observation suggests that liver steatosis 

could be a concern late in life for animals exposed to PFOA gestationally and during their early 

postnatal period. However, the 18-month fat accumulation could also be related to normal aging 

and/or dietary fat intakes across the animal’s lifetime (Quist et al. 2015). Tan et al. (2013) found 

that the fat content of the diet was an important variable in determining the impact of PFOA 

(5 mg/kg/day) on liver and serum lipids. Intake of an HFD plus PFOA increased liver 

triglycerides and serum free fatty acids compared to an RFD plus PFOA but had no impact on 

liver cholesterol concentrations. Serum cholesterol was not monitored. 

Hepatic Effects. Both the human and animal studies suggest effects on the liver as indicated by 

increases in liver enzymes. The results of the occupational studies provide evidence of an 

association with increases in serum AST, ALT and GGT, with the most consistent results seen 

for ALT. The associations were not large and could depend on the covariates in the models, such 

as BMI, use of lipid-lowering medications, and triglycerides (Costa et al. 2009; Olsen et al. 2000, 

2003; Olsen and Zobel 2007; Sakr et al. 2007a, 2007b). Two population-based studies of highly 

Perfluorooctanoic acid (PFOA) – May 2016 3-141
exposed residents in contaminated regions near a fluorochemical industry in West Virginia have 

evaluated associations with liver enzymes, and the larger of the two studies reported associations 

of increasing serum ln ALT and ln GGT levels with increasing serum PFOA concentrations 

(Emmett et al. 2006; Gallo et al. 2012). A cross-sectional analysis of data from NHANES, 

representative of the U.S. national population, also found associations with ln PFOA 

concentration with increasing serum ALT and ln GGT levels. Serum bilirubin was inversely 

associated with serum PFOA in the occupational studies. A U-shaped exposure-response pattern 

for serum bilirubin was observed among the participants in the C8 Health Project, which might 

explain the inverse associations reported for occupational cohorts. Overall, an association of 

serum PFOA concentration with elevations in serum levels of ALT and GGT has been 

consistently observed in occupational, highly exposed residential communities, and the U.S. 

general population. The associations are not large in magnitude, but indicate the potential to 

affect liver function.

The data from animal studies for increases in ALT and AST support the findings in human 

epidemiology studies; however, the animal studies for both aminotransferases lacked serum 

PFOA measurements for comparison with the human serum data. Concurrent with the evidence 

in animals of damage to liver cells, levels of some membrane transport proteins were altered. In 

mice, the increased expression of MRP3 and MRP4 (Maher et al. 2008) and the decreased 

expression of OATPs (Cheng and Klaassen 2008) favor excretion of PFOA into the bile. 

Competition of PFOA with bile acids for transport could alter the excretion of the cholesterol 

derivatives excreted in bile. 

In animal studies, serum levels of ALT and/or AST were significantly increased indicating 

apoptosis or necrosis of liver cells (Butenhoff et al. 2012; Minata et al. 2010; Son et al. 2008).

Increased levels of ALT were observed at a LOAEL of 2.65 mg/kg/day in ICR mice by Son et al. 

(2008). Yahia et al. (2010) reported significantly increased ALT, GGT, AST, and ALP in PFOAexposed (10 mg/kg) pregnant ICR mice. Total protein, albumin, and globulin were significantly 

decreased in the same mice. 

No evidence of liver damage has been found in the human epidemiology studies with the 

exception of the few enzyme changes discussed above. In most PFOA animal studies 

(e.g., monkeys, rats, and mice), short-term and chronic exposure caused a dose-related increase 

in liver weight as at least one of the co-occurring effects (Butenhoff et al. 2002, 2004a, 2012; 

DeWitt et al. 2009; Elcombe et al. 2010; Minata et al. 2010; Pastoor et al. 1987; Perkins et al. 

2004; Son et al. 2008; Wolf et al. 2008a). Increased liver weights were observed in mice that are 

both active and null for PPARα activation (Albrecht et al. 2013; Minata et al. 2010; Wolf et al. 

2008a). The histological characteristics of the liver differed in the mice with and without the 

PPARα receptor, but the liver weight increase was the same. Liver effects were seen in mice 

with an active PPARα receptor at doses as low as 0.3 mg/kg/day (Nakamura et al. 2009) and 

1 mg/kg/day in the null mice (Wolf et al. 2008a).

Histological examination of liver tissues from PFOA-exposed wild-type mice and PPARαnull mice were distinctly different from their respective controls (Minata et al. 2010; Wolf et al. 

2008a). In the case of the wild-type PFOA-exposed mice, there was less rough ER than in 

controls and more lipid-like vacuoles scattered throughout the cytoplasm. The PFOA-exposed 

PPARα-null mice had proliferation of smooth ER and limited rough ER and Golgi bodies 

compared to their controls. The PPARα-null control mice had the scattered lipid-like vacuoles 

seen in the wild-type PFOA exposed mice; however, their lipid-like vacuoles were considerably 

larger than those seen in the wild-type animals and occupied a considerable volume within the 

Perfluorooctanoic acid (PFOA) – May 2016 3-142
cytoplasm. The vacuoles in the PPARα-null PFOA-exposed mice were hypothesized to be filled 

with PFOA as a consequence of its uptake into the cell without dispersion or assimilation.

Similarly, Albrecht et al. (2013) observed centrilobular hepatocellular hypertrophy in mouse 

dams given 3 mg/kg on GDs 1–17, but the morphological features differed slightly between 

wild-type, PPARα-humanized, and PPARα-null mice. In wild-type mice, hypertrophy was 

characterized primarily by centrilobular hepatocytes with increased amounts of densely 

eosinophilic and coarsely granular cytoplasm consistent with increased peroxisomes. In null 

mice, hypertrophy was generally less prominent than seen in wild-type mice, and affected 

hepatocytes had pale eosinophilic, finely granular-to-amorphous cytoplasm. The morphological

features of centrilobular hepatocytes in humanized mice were intermediate between those

observed in wild-type and null mice. The lesion was graded as mild in wild-type mice, minimal 

in null mice, and minimal or mild in humanized mice. An additional finding in PFOA-treated 

null and humanized mice, but not in wild-type mice, was the presence of few clear, discrete 

vacuoles within the cytoplasm of centrilobular hepatocytes. 

Hepatocellular hypertrophy and an increased liver-to-body weight ratio are common findings 

in rodents when PPARα activation leads to peroxisome proliferation and these effects are 

considered nonadverse in wild-type strains when they occur. Hepatic necrosis, effects on bile 

ducts, and other signs of liver damage unrelated to PPARα activation observed in conjunction 

with the increased liver weight and hepatpcellular hypertrophy are sufficient to justify the liver 

weight and hypertrophy as adverse (Hall et al. 2012). Low-level necrotic cell damage was 

observed in the Perkins et al. (2004) rat study and in the Loveless et al. study (2008) in CD rats 

at 10 mg/kg/day and CD1 mice at 1 mg/kg/day. In the Perkins et al. study (2004), there was a 

slight increase in coagulative necrosis at 1.94 and 6.5 mg/kg/day when compared to the control 

and lower doses. Some hepatocellular necrosis also was observed in conjunction with 

hepatocellular hypertrophy and increased liver weight in F1 male rats from the Butenhoff et al. 

(2004a) two-generation study at 3 mg/kg/day. 

Minata et al. (2010) reported degenerative histological changes in the bile ducts of PPARαnull mice at doses ≥10.6 mg/kg/day and Loveless et al. (2008) observed bile duct hyperplasia in 

CD1 mice at doses ≥10 mg/kg/day. PPARα-null mice had an increased hepatocyte PCNA

labeling index at a dose of 10 mg/kg/day (Wolf et al. 2008a). When considering the studies in 

animals with and without the active PPARα receptor, it is clear that PFOA has some effects of 

potential toxicological significance that appear to be independent of PPARα activation.

Kidney and Other Organ Effects. Overall, studies of occupational cohorts (Costa et al. 2009), a 

highly exposed community (Steenland et al. 2010; Watkins et al. 2013), and the U.S. general 

population (Shankar et al. 2011) that evaluated uric acid levels or eGFR as measure of kidney 

function found associations with decreased function. Reverse causality as an explanation cannot 

be ruled out in studies using serum PFOA as a biomarker of exposure, as a low GFR would 

diminish the removal of PFOA from serum for excretion by the kidney. 

Some studies in animals have shown effects on the kidney, mainly increased organ weight in

male rats, but the studies lacked concurrent PFOA serum levels and histological examination of 

the kidney tissues. In general, kidney effects in rats occurred at doses similar to those resulting in 

liver effects.

Increases in absolute and relative-to-body kidney weights occurred in rats given 5 mg/kg/day 

(lowest dose tested) for 28 days (Cui et al. 2009). In a two-generation study, F0 and F1 males 

Perfluorooctanoic acid (PFOA) – May 2016 3-143
had significantly increased absolute kidney weight at 1 and 3 mg/kg/day, but significantly 

decreased organ weight at 30 mg/kg/day. Organ weight-to-terminal body weight ratios for the

kidney were statistically significantly increased at ≥1 mg/kg/day. Kidney weight-to-brain weight 

ratios were significantly increased at 1, 3, and 10 mg/kg/day, but decreased at 30 mg/kg/day, 

following the trends in absolute weights (Butenhoff et al. 2004a). In the high-dose group, 

absolute and relative kidney weight changes occurred in a pattern typically associated with 

decrements in body weight. However, in the lower dose groups kidney weight, consistently 

displayed an increase (absolute and relative to body and brain weights), suggesting an induction 

of transporters for renal removal of the foreign molecule. The differential expression of 

transporters in the kidney of rats has been shown to be under hormonal control with males 

having lower levels of export transporters compared with females (Kudo et al. 2002). 

In both the Cui et al. (2009) and Butenhoff et al. (2004a) studies, PFOA was administered by 

daily gavage. No changes in kidney weight were found with dietary administration with a 

resulting dose of 14.2 mg/kg/day to male rats for 2 years (Butenhoff et al. 2012). 

In general, effects on organs other than the liver tend to occur at doses higher than those that 

affect the liver. Lung effects, including pulmonary congestion, have been observed in male 

Sprague-Dawley rats (LOAEL = 5 mg/kg/day) (Cui et al. 2009). Increased thickness and 

prominence of the adrenal zona glomerulosa and vacuolation in the cells of the adrenal cortex 

were observed in male rats fed 10 mg/kg/day for approximately 56 days (Butenhoff et al. 2004a). 

Thyroid Effects. Three large studies provide support for an association between PFOA exposure 

and incidence or prevalence of thyroid disease in female adults or children, but not in males

(Lopez-Espinosa et al 2012; Melzer et al. 2010; Winquist and Steenland 2014b). In addition, 

associations between PFOA and TSH were also seen in pregnant females with anti-TPO 

antibodies (Webster et al. 2014). However, generally null associations were found between 

PFOA and TSH or thyroid hormones (T4 or T3) in people who have not been diagnosed with 

thyroid disease. 

Effects of PFOA on thyroid hormones in animals are generally not as well characterized as 

those of PFOS. Butenhoff et al. (2002) evaluated the toxicity of PFOA in male cynomolgus 

monkeys during 6 months of oral administration and reported that levels of total T3 and FT3 in 

circulation were reduced significantly in the 30/20 mg/kg/day treatment group. The effect seen as 

early as 5 weeks after initiation of treatment, 2 weeks after the dose was lowered to 20 

mg/kg/day. Recovery of T3 deficits was noted upon cessation of chemical treatment once the 

serum level of PFOA returned to baseline 90 days later. Serum total T4, FT4, and TSH were not 

altered throughout the study. The preferential effects of PFOA on serum T3 and a lack of a TSH 

compensatory response are similar to those observed with PFOS. 

Martin et al. (2007) showed that serum total T4 and FT4 were markedly and abruptly 

depressed (~ 80%) in adult male rats 1 day after oral gavage treatment with PFOA (20 mg/kg); 

serum T3 was also reduced (25%), although to a lesser extent. These findings were confirmed 

when both male and female rats were given PFOA (10 mg/kg) daily for 3 weeks and serum 

thyroid hormones were monitored (Lau, personal communication) (Martin et al. 2007). Serum 

total T4 and FT4 were profoundly depressed (>85%) and T3 less so (~ 25%) in male rats, but 

serum TSH levels were not altered significantly. These hormonal changes were noted when 

serum PFOA level reached about 67 μg/ml. The dose-response relationship of serum total T4 

with PFOA exposure has yet to be fully evaluated and the lowest effective dose remains 

unknown. 

Perfluorooctanoic acid (PFOA) – May 2016 3-144
None of the thyroid hormones were affected by PFOA in mature female rats, primarily 

because these animals were able to clear the chemical effectively (with half-life estimate of 2–4 

hours compared to that of 6–7 days for male rats). This suggests that the thyroid disrupting 

effects of PFOA are directly related to endogenous accumulation of the chemical and could be 

relevant to humans because of the long PFOA human half-life.

Displacement of T4 from binding to TTR has been proposed as a possible mechanism to 

account for the hypothyroxinemia in rats. However, although PFOA binds to human TTR, its 

binding affinity is only one-fifteenth of that of the natural ligand T4 (Weiss et al. 2009). Based 

on a toxicogenomic analysis of rat liver after an acute exposure to PFOA, Martin et al. (2007) 

suggested a possible role of peroxisome proliferators in the thyroid hormone imbalance, although 

this hypothesis has yet to be explored in detail. 

Hyperglycemia. Several human epidemiology studies have examined PFOA in relation to 

diabetes (incidence or prevalence) or measures of hyperglycemia These studies do not show a 

pattern of results that suggest an association between PFOA and diabetes or hyperglycermia in 

occupational settings (Costa et al. 2009; Olsen et al. 2000, 2003; Sakr et al. 2007a; Steenland et 

al. 2015), in the high-exposure community population (MacNeil et al. 2009), or in the general 

population (Lin et al. 2009; Nelson et al. 2010). 

Hines et al. (2009) found no differences in glucose tolerance tests at 15–16 weeks and at 

17 months of age in PFOA-exposed CD-1 mice, but did observe significantly increased serum 

leptin and insulin levels at 21 and 31 weeks of age, suggesting that the insulin resistance 

mechanistic pathway could be affected by PFOA. Conversely, Quist et al. (2015) found no doserelated impact on serum leptin in CD-1 pups from the Hines et al. study (2009) when examined 

on PND 91 for the mice on an RFD and on an HFD fasted for 4 hours before serum collection. In 

the animals on a HFD that did not fast before serum collection, there was a trend towards a doserelated decrease in serum leptin. Thus, the fat content of the diet and the timing of serum 

collection are important variables that can influence study results relative to leptin levels and 

indicators of insulin resistance.

Nervous System Effects. The data pertaining to neurotoxicity (including neurodevelopmental 

effects) of PFOA are limited, but do not indicate the presence of associations between PFOA and 

a variety of outcomes. Fei et al. (2008b) found no association between maternal serum PFOA 

concentrations and fine motor skills, gross motor skills, and cognitive abilities of children aged 6 

and 18 months. Fei and Olsen (2011) found no association between behavioral or coordination 

problems in children aged 7 years and prenatal PFOA exposure. Epidemiology studies of 

children derived from the NHANES and C8 populations found a weak statistical association 

between serum PFOA with parental reports of ADHD (Hoffman et al. 2010; Stein et al. 2013).

One animal study (Johansson et al. 2009) suggests a potential effect on habituation and 

activity patterns in NMRI mice treated on PND 10 with a single dose of PFOA and evaluated at 

and 2 and 4 months of age (LOAEL=0.58 mg/kg). The in vivo observations were supported by 

changes in the expression of a variety of neurologically active brain proteins in the treated pups 

(Johansson et al. 2009). The offspring of C57BL/6/Bkl dams fed 0.3 mg PFOA/kg/day 

throughout gestation had detectable levels of PFOA in their brains at birth (Onishchenko et al. 

2011). Behavioral assessments of the offspring starting at 5 weeks of age revealed gender-related 

differences in exploratory behavior patterns. In the social group setting, the PFOA-exposed 

males were more active and PFOA-exposed females were less active than their respective 

controls. The PFOA-exposed males also had increased activity counts compared to control males 

Perfluorooctanoic acid (PFOA) – May 2016 3-145
in circadian activity experiments. The results of an in vitro study of hippocampal synaptic 

transmission and neurite growth in the presence of long-chain perfluorinated compounds showed 

that 50 and 100 μmol PFOA increased spontaneous synaptic current and had an equivocal impact 

on neurite growth (Liao et al. 2009a, 2009b). These data suggest a need for additional studies of 

the effects of PFASs, including PFOA, on the brain. 

Reproductive and Developmental Effects. There have been numerous human studies 

examining PFOA exposure and reproductive and/or developmental effects in both humans and 

animals. A series of studies in the high-exposure C8 Health Project study population have 

reported associations between PFOA exposure and pregnancy-induced hypertension or 

preeclampsia (Darrow et al. 2013; Savitz et al. 2012a, 2012b; Stein et al. 2009). Each of these 

studies provides evidence of an association between PFOA exposure and risk of pregnancyinduced hypertension or preeclampsia, with the most robust findings from the methodologically 

strongest study (Darrow et al. 2013). 

The association between PFOA and birth weight has been examined in numerous human 

studies. Most studies measured PFOA using maternal blood samples taken in the second or third 

trimester or in cord blood samples. Studies on the high-exposure C8 community population 

(Darrow et al. 2013; Nolan et al. 2009; Stein et al. 2009; Savitz et al. 2012a, 2012b) have not 

observed associations between PFOA and either birth weight among term births or the risk of 

low birth weight among all (singleton) births. In contrast, several analyses of general populations 

indicate a negative association between PFOA levels and birth weight (Apelberg et al. 2007; Fei 

et al. 2007; Maisonet et al. 2012), while others did not attain statistical significance (Chen et al. 

2012; Hamm et al. 2010; Monroy et al. 2008; Washino et al. 2009). A meta-analysis of many of 

these studies found a mean birth weight reduction of 19 g (95% CI: -30, -9) per each 1-unit 

(ng/mL) increase in maternal or cord serum PFOA levels (Johnson et al. 2014). However, when 

low GFR was accounted for in PBPK simulations by Verner et al. (2015), the association 

reported between PFOA and birth weight is less than that found in their meta-analysis of the 

epidemiology data. The study authors reported that the actual association might be closer to a 7-g 

reduction (95% CI: -8, -6). Verner et al. (2015) also showed that, in individuals with low GFR, 

there are increased levels of serum PFOA and lower birth weights. This suggests that a portion of 

the association between PFOA and birth weight could be confounded by low maternal GFR

under conditions such as preeclampsia and pregnancy-induced hypertension. While there is some 

uncertainty in the interpretation of the observed association between PFOA and low GFR with

birth weight, given the available information, the association between PFOA exposure and 

reduced birth weight observed for the general population is plausible. In humans with low GFR, 

the impact on body weight is likely due to a combination of the low GFR and the serum PFOA. 

Two studies examined development of puberty in females in relation to prenatal exposure to 

PFOA as measured through maternal or cord blood samples in follow-up of pregnancy cohorts 

conducted in England (Christensen et al. 2011) and in Denmark (Kristensen et al. 2013). The 

results of these two studies are conflicting, with no association (or a possible indication of an 

earlier menarche seen with higher PFOA) in Christensen et al. (2011), and a later menarche seen 

with higher PFOA in Kristensen et al. (2013). Another study examined PFOA exposure 

measured concurrently with the assessment of pubertal status (Lopez-Espinosa et al. 2011). An 

association between later age at menarche and higher PFOA levels was observed, but the 

interpretation of this finding is complicated by the potential effect of puberty on the exposure 

biomarker levels (i.e., reverse causality).

Perfluorooctanoic acid (PFOA) – May 2016 3-146
Limited data suggest a correlation between higher PFOA levels (>0.02 μg/mL) in females

and decreases in fecundity and fertility (Fei et al. 2009; Vélez et al. 2015), but there are no clear 

effects of PFOA on male fertility endpoints (0.0035–0.005 μg/mL) (Joensen et al. 2009, 2013). 

Knox et al. (2011) found that the odds of having experienced menopause were significantly 

higher in the highest PFOA quintile group relative to the lowest PFOA group. Two studies found 

delayed puberty in females (Kristensen et al. 2013; Lopez-Espinosa et al. 2011), but reverse 

causality needs to be considered. However, Christensen et al. (2011) found no association 

between puberty and PFOA exposure in children of the Avon Longitudinal Study of Parents and 

Children in the United Kingdom. Removal of PFOA with the start of monthly menstruation and 

the cessation of this route with menopause or hysterectomy are additional factors that can 

influence serum PFOA levels that are the result of the developmental milestones rather than a 

cause (Taylor et al. 2014; Wong et al. 2014). Costa et al. (2009) found no association between 

serum PFOA concentration and estradiol or testosterone in workers at a PFOA production plant. 

Measures of postnatal development and behavior in children were not associated with PFOA 

levels in the mother (0.001–0.0057 μg/mL) (Andersen et al. 2010, 2013; Fei et al. 2008b; Fei and 

Olsen 2011; Høyer et al. 2015a, 2015b). Fei et al. (2008b) found no association between 

maternal PFOA concentration and fine motor skills, gross motor skills, and cognitive skills in 

offspring at 6 and 18 months of age. Fei and Olsen (2011) also found no association between 

prenatal PFOA exposure and behavior or coordination problems in children aged 7 years. The 

age at which children reached developmental milestones did not show any relationship to 

maternal plasma PFOA concentration. Halldorsson et al. (2012) found that low-dose 

developmental exposures to PFOA resulted in obesogenic effects in female offspring at 20 years.

Among the animal studies, there was no effect of PFOA on reproductive or fertility 

parameters in rats (Butenhoff et al. 2004a; York et al. 2010), but effects on male fertility were 

observed in male mice (Lu et al. 2015). In mouse gavage studies, decreased body weight and 

decreased neonatal survival were observed at ≥ 1 mg/kg/day, increased full litter resorptions and 

increased stillbirths were observed at ≥ 5 mg/kg/day, increased time to parturition was observed 

at ≥10 mg/kg/day, and decreased maternal weight gain was observed at ≥20 mg/kg/day for 

exposures lasting from GD1–17 (Abbott et al. 2007; Lau et al. 2006; White et al. 2007; Wolf et 

al. 2007). 

Postnatal development also has been studied extensively in rats and mice as discussed below.

A separate group of studies in mice focused on mammary gland development in dams and 

female offspring. Both species showed some indication of potential developmental toxicity. 

Doses that elicited a response were higher in rats compared with in mice. The species differences 

in dose response are likely related to half-life differences of hours for the female rat and days-toweeks for the female mouse.

Reduced postnatal growth leading to developmental delays was observed in both rats and 

mice. A two-generation diet study in rats resulted in significantly decreased body weight gain 

prior to weaning and delayed sexual maturity in the first generation males and females at 

30 mg/kg/day (Butenhoff et al. 2004a). For treatment beginning on PND 21, delayed vaginal 

opening was also observed in BALB/c mice at ≥1 mg/kg/day and in C57BL/6 mice at 

≥5 mg/kg/day, although body weight was not decreased until doses of ≥10 mg/kg/day in both 

strains (C. Yang et al. 2009). Cross-fostering studies in mice showed that gestational PFOA 

exposure maximized decreased postnatal body weight, delayed eye opening, delayed body hair 

growth, and decreased survival in the offspring (Wolf et al. 2007). Restricted exposure studies 

Perfluorooctanoic acid (PFOA) – May 2016 3-147
showed that gestational exposure to PFOA over differing gestational time periods led to differing 

offspring effects (Wolf et al. 2007). The longer the gestational exposure, the greater the body 

weight deficit in the male and female pups over PND 2–22. In males, the difference in body 

weight persisted until PND 92. Delayed eye opening and body hair growth were observed at 

5 mg/kg/day in offspring exposed GD 7–17 or 10–17, but decreased postnatal survival was 

observed at the same dose in offspring exposed GD 15–17. 

Two developmental studies compared wild-type mice with PPARα-null mice, but results are 

inconclusive. One study revealed that the litter resorptions were independent of PPARα 

expression (≥5 mg/kg), while decreased neonatal survival (0.6 mg/kg) and delayed eye opening 

(1 mg/kg) were dependent upon PPARα expression (Abbott et al. 2007). These results are only 

partially supported by Albrecht et al. (2013), who used a single dose of 3 mg/kg. They found 

decreased pup survival only in wild-type mice, but no differences in litter resorptions or eye 

opening between wild-type and null mice. Albrecht et al. (2013) did not find effects on pup 

survival in PPARα-humanized mice, suggesting that the mouse PPARα could be involved in the 

etiology of PFOA-induced neonatal mortality. 

Qualitative assessment found delayed mammary gland development of female CD1 mouse 

pups following maternal doses ≥ 0.01 mg PFOA/kg in Macon et al. (2011) and Tucker et al. 

(2015). Macon et al. (2011) also found significant differences from controls in quantitative 

measures of longitudinal and lateral growth and numbers of terminal end buds at 1 mg/kg/day. 

However, Albrecht et al. (2013) found no significant differences in the average length of 

mammary gland ducts and the average number of terminal end buds per mammary gland per 

litter in female pups of PPARα wild-type, PPARα-null, or hPPARα sv/129 mice following a 

maternal dose of 3 mg/kg, using an approach to scoring that lacked a qualitative component 

adjustment such as that used by Macon et al. (2011). 

The approach to scoring mammary gland development was not consistent across studies, and 

little information was provided on the qualitative components of the scores. This makes 

comparisons across studies difficult. Statistical significance was attained at higher dose levels for 

the quantitative portion of the Macon et al. (2011) scoring protocol than for the qualitative 

component of the score. The process used to score the qualitative developmental score by Macon 

et al. (2011) was not described. Tucker et al. (2015) found that CD-1 mice were considerably 

more sensitive to effects on mammary gland development (LOAEL = 0.01 mg/kg/day) than 

C57BL/6 mice (NOAEL 0.1 mg/kg/day). Scoring was conducted using the Macon et al (2011) 

approach. 

White et al. (2011) used doses of 0 and 1 mg PFOA/kg/day to F0 dams throughout gestation 

with and without the addition of drinking water containing 5 ppb PFOA beginning on GD 7 and 

continuing the contaminated drinking water during the production of two more generations; no 

persistent significant differences were found in the body weights of the pups in the F1 and F2 

generations for the pups receiving 1 mg/kg/day, indicating a poor correlation between mammary 

duct branching patterns and the ability to support pup growth during lactation. The 5 mg/kg/day 

dose did have an impact on body weight. Albrecht et al. (2013) also found no significant impacts 

on pup body weight in their one-generation assay at a dose of 3 mg/kg/day. Despite the 

diminished ductal network assessed in the qualitative mammary gland developmental score of 

the dams in White et al. (2011), milk production was sufficient to nourish growth in the exposed 

pups as reflected in the body weight measurements compared to controls. The MoA for PFOAinduced delayed mammary gland development is unknown and requires further investigation. 

Perfluorooctanoic acid (PFOA) – May 2016 3-148
At doses of 5 and 10 mg/kg/day, mammary gland development was delayed in BALB/c mice 

(C. Yang et al. 2009). In C57BL/6 mice, mammary gland development was accelerated at 

5 mg/kg/day, but delayed at 10 mg/kg/day, indicating strain differences in pubertal mammary 

gland development following a dose of 5 mg/kg/day. Y. Zhao et al. (2010) showed that 5 mg 

PFOA/kg/day stimulates mammary gland development in C57BL/6 mice by promoting steroid 

hormone production in the ovaries and increasing mammary gland growth factor levels. 

Immune Effects. Associations between prenatal, childhood, or adult PFOA exposure and risk of 

infectious diseases (as a marker of immune suppression) have not been consistently seen, 

although there was some indication of effect modification by gender (i.e., associations seen in 

female children but not in male children) (Fei et al. 2010a; Granum et al. 2013; Looker et al. 

2014; Okada et al. 2012).

The WHO guidelines for immunotoxicity risk assessment recommend measures of vaccine 

response as a measure of immune effects, with potentially important public health implications 

(WHO 2012). Three studies have examined associations between maternal and/or child serum 

PFOA levels and vaccine response (measured by antibody levels) in children (Grandjean et al. 

2012; Granum et al. 2013) and adults (Looker et al. 2014). The study in adults was part of the 

high-exposure community C8 Health Project; a reduced antibody response to one of the three 

influenza strains tested after receiving the flu vaccine was seen with increasing levels of serum 

PFOA; these results were not seen with PFOS. The studies in children were conducted in general 

populations in Norway and in the Faroe Islands. Decreased vaccine response in relation to PFOA 

levels was seen in these studies, but similar results also were seen with correlated PFASs 

(e.g., PFOS).

Several animal studies demonstrate effects on the spleen and thymus as well as their cellular 

products (B lymphocytes and T-helper cells) in several strains of mice. Studies by Yang et al. 

(2000, 2001, 2002b) and DeWitt et al. (2008) were conducted using relatively high PFOA doses 

(~30–40 mg/kg/day). In each study, the PFOA-treated animals exhibited significant decreases in 

spleen and thymus weights as well as in splenocyte and thymocyte populations at various stages 

of differentiation. Recovery usually occurred within several days of cessation of PFOA dosing. 

However, when the response of C57BL/6 Tac PPARα mice was compared to wild-type of the 

same strain, the KO mice showed no response of both spleen and thymus weights at 

30 mg/kg/day, whereas there was a response in the wild-type strain (DeWitt et al. 2015). Both 

strains showed an increase in IgM in response to a SRBC injection. The 30-mg/kg/day dose was 

the LOAEL for the KO mice and 7.5 mg/kg/day was the response level for the wild-type strain. 

Thus the suppression of the immune system is not totally a PPARα-related response. In a similar 

experiment (Yang et al. 2002a), no significant changes in spleen weight or cellularity were 

observed in PPARα-null mice as compared to wild-type mice, but there was a small and 

significant decrease in thymus weight and cellularity compared to controls.

DeWitt et al. (2008) used different functionality assays in their study in C57Bl/6 mice. The 

IgM response to SRBC was suppressed by 20% when mice were immunized immediately after 

exposure to the initial dose of 30 mg PFOA/kg/day ceased. However, there was no significant 

increase in the response to BSA 4 days post-PFOA exposure or in the IgG response to SRBC 

15 days post-PFOA exposure. These results are indicative of recovery once PFOA exposures 

have ceased.

Perfluorooctanoic acid (PFOA) – May 2016 3-149
DeWitt et al. (2008) followed their initial study of PFOA with one designed to examine the 

dose response for a 15-day drinking water exposure in a slightly different mouse strain, 

C57Bl/6N. The study design examined the spleen and thymus weights, splenocyte and 

thymocyte numbers, and IgM response of the immune system to the immunological challenges 

as described above. The LOAEL was 3.75 mg/kg/day based on a significant decrease in IgM 

response, and the NOAEL was 1.88 mg/kg/day. 

In one component of the Yang et al. study (2002b), the functional impact of changes in 

spleen and thymus were evaluated through the response of treated mice to HRBCs. The control 

mice responded to the HRBC exposure with an increased plaque-forming response; however, the 

PFOA-treated mice did not have an increased plaque-forming response when tested (Yang et al. 

2002b). In addition, when blood from PFOA-treated mice was evaluated posttreatment, there 

was no increase in lymphocyte proliferation in response to the addition of Con-A and LPS to the 

test media. The control mice responded with the expected lymphocyte proliferation after the 

addition of Con-A and LPS antigens. 

Loveless et al. (2008) looked at the IgM response to SRBC in male CD rats and CD-1 mice 

following a 29-day exposure to 0–30 mg PFOA/kg/day. The thymus and spleen cell counts and 

organ weights and the IgM titers were not altered by PFOA treatment in rats. In mice, however, 

thymus and spleen weights, thymus and spleen cell counts, and IgM titers were decreased at 

≥10 mg PFOA/kg/day. CORT also was increased in mice at the same doses.

The data collected from the immunotoxicity studies support a MoA through which PFOA 

interferes with splenocyte and thymocyte precursor cells in the bone marrow as well as 

maturation of the cells once they have been transported to their respective organs. Examination 

of cell populations at different stages of development reveals lower numbers of the CD4-

CD8-

cells formed in bone marrow as well as decreased populations of splenocyte and thymocyte cells 

at different stages of expressing the surface proteins that mark them as functional beta 

lymphocytes (thymus) or T-helper cells (spleen) (Son et al. 2009). Although the studies that 

measured the splenocyte and thymocyte populations were carried out at doses higher than the 

3.75 mg/kg/day LOAEL observed by DeWitt et al. (2008), the fact that the IgM response to an 

antigenic material was decreased at that dose indicates an inability to produce antibodies at 

adequate levels when exposed to a challenge.

Loveless et al. (2008) hypothesized that the observed effects on serum lymphocytes could be 

the result of adenocorticotropic steroids in a response to stress. A study by DeWitt et al. (2009) 

in which the immunological response of ADX mice treated with PFOA were compared to shamoperated controls did not support the Loveless et al. (2008) hypothesis. 

Data from PPARα-null mice suggest that rodents might be more susceptible to the 

immunosuppressive impacts of PFOA than humans. However, the fact that there were still 

effects on the thymus weight and cellularity even in the PPARα-null mouse strain indicate the 

potential for an inadequate humoral response in exposed populations. 

3.4.2 Synthesis and Evaluation of Carcinogenic Effects

Evidence of carcinogenic effects of PFOA in epidemiology studies is based primarily on 

studies of kidney and testicular cancer. These cancers have relatively high survival rates 

(e.g., 2005–2011 5-year survival rates 73% and 95%, respectively, for kidney and testicular 

cancer based on NCI Surveillance, Epidemiology and End Results data). Thus studies that 

Perfluorooctanoic acid (PFOA) – May 2016 3-150
examine cancer incidence are particularly useful for these types of cancer. The high-exposure 

community studies also have the advantage, for testicular cancer, of including the age period of 

greatest risk, as the median age at diagnosis is 33 years. The two occupational cohorts in 

Minnesota and West Virginia (most recently updated in Raleigh et al. 2014 and Steenland and 

Woskie 2012) do not support an increased risk of these cancers, but each of these is limited by a 

small number of observed cases (six kidney cancer deaths, 16 incident kidney cancer cases, and 

five inciden testicular cancer cases in Raleigh et al. 2014; and 12 kidney cancer deaths and 1 

testicular cancer death in Steenland and Woskie 2012). Two studies involving members of the 

C8 Health Project showed a positive association between PFOA levels (mean at enrollment 

0.024 μg/mL) and kidney and testicular cancers (Barry et al. 2013; Vieira et al. 2013); there is 

some overlap in the cases included in these studies. No associations were found in the general 

population between mean serum PFOA levels up to 0.0866 μg/mL and colorectal, breast, 

prostate, bladder, and liver cancer (Bonefeld-Jørgensen et al. 2014; Eriksen et al. 2009; Hardell 

et al. 2014; Innes et al. 2014); none of these studies examined kidney or testicular cancer. 

Two animal carcinogenicity studies indicate that PFOA exposure can lead to liver adenomas 

(Biegel et al. 2001), Leydig cell adenomas (Biegel et al. 2001; Butenhoff et al. 2012), and 

PACTs (Biegel et al. 2001) in male Sprague-Dawley rats. Liver adenomas were observed in the 

Biegel et al. study (2001) at an incidence of 10/76 (13%) at 20 mg/kg/day. The incidence in the 

control group was 2/80 (3%). Although no liver adenomas were observed in Butenhoff et al. 

(2012), carcinomas were identified in the male controls, males in the low-dose group 

(2 mg/kg/day), and male and female rats in the high-dose group (20 mg/kg/day). The differences 

from control were not significant in either study, but the carcinoma incidence among the 

Butenhoff et al. (2012) high-dose males (10/50) was similar to that for the adenomas in the 

Biegel et al. study (2001) (10/76). Liver lesions were identified in the males and females at the 

1- and 2-year sacrifices (Butenhoff et al. 2012). An increased incidence of diffuse 

hepatomegalocytosis and hepatocellular necrosis occurred at 20 mg/kg/day. At the 2-year 

sacrifice, hepatic cystoid degeneration (characterized by areas of multilocular microcysts in the 

liver parenchyma) was observed in 8, 14, and 56% in males of the control, 2-, and 20-mg/kg/day 

dose groups, respectively. Hyperplastic nodules in male livers were increased in the high-dose 

group (6% versus 0% in control rats).

Filgo et al. (2015) examined the livers of three strains of mice exposed only during 

gestation/lactation for tumors when they were sacrificed at 18 months. Liver tumors were found 

in each dose group, but tumor types varied and the data did not display any evidence of dose 

response. The animals were survivors from two different projects and the number per dose group 

was small. Thus, the data are not adequate for determining whether PFOA is a carcinogen in 

mice.

Testicular LCTs were identified in both the Butenhoff et al. (2012) and Biegel et al. (2001) 

studies. The tumor incidence was 0/50 (0%), 2/50 (4%), and 7/50 (14%) for the control, 2.0-, and 

20-mg/kg/day dose groups, respectively (Butenhoff et al. 2012). The Biegel et al. study (2001) 

included one dose group (20 mg/kg/day); the tumor incidence was 8/76 (11%) compared to 0/80 

(0%) in the control group. LCT incidence at 20 mg/kg/day was comparable between the two 

studies (11 and 14%). 

PACTs were only observed in the Biegel et al. study (2001). The incidence was 8/76 (11%; 

7 adenoma, 1 carcinoma) at 20 mg/kg/day while none were observed in the control animals. 

Although no PACTs were observed by Butenhoff et al. (2012), pancreatic acinar hyperplasia was 

observed at 2 and 20 mg/kg/day at incidences of 2/34 (6%) and 1/43 (2%), respectively, which 

Perfluorooctanoic acid (PFOA) – May 2016 3-151
lacked dose response. Reexamination of the pancreatic lesions in Butenhoff et al. (2012) and 

Biegel et al. (2001) resulted in the conclusion that 20 mg/kg/day increased the incidence of 

proliferative acinar cell lesions in both studies. Some lesions in the Biegel et al. study (2001) had 

progressed to adenomas. 

The initial findings from the Butenhoff et al. study (2012) were equivocal for mammary 

fibroadenomas in female rats. However, a reexamination of the tissues by a PWG found no 

statistically significant differences in the incidence of fibroadenomas or other neoplasms of the 

mammary gland between control and treated animals (Hardisty et al. 2010). The PWG used the 

diagnostic criteria and nomenclature of the Society of Toxicological Pathologists for the 

reexamination. Under those criteria, there was an increase in the number of tumors documented 

in the control group, especially fibroadenomas originally classified as lobular hyperplasia. The 

reclassification led to a loss of significance when the tumors in the treated animals were 

compared to tumors in the control animals. 

Ovarian tubular hyperplasia and adenomas also were observed in female rats (Butenhoff et 

al. 2012). Mann and Frame (2004) reexamined the ovarian lesions using an updated 

nomenclature system, which resulted in some of the hyperplastic lesions being reclassified. The 

ovarian lesions originally described as tubular hyperplasia or tubular adenomas were regarded as 

gonadal stromal hyperplasia and/or adenomas. After the reclassification, there were no 

statistically significant increases in hyperplasia (total number), adenomas, or 

hyperplasia/adenoma combined in treated groups compared to controls. 

Mutagenicity studies of PFOA using the S. typhimurium (Friere et al. 2008; Lawlor 1995, 

1996) and E. coli (Lawlor 1995, 1996) system have resulted in negative results in the presence 

and absence of activation. One mutagenicity study (Lawlor 1995, 1996) in S. typhimurium gave a 

positive result, but it was not reproducible. Clastogenicity studies in CHO by Murli (1996b, 

1996c) were positive with activation for chromosomal abnormalities and polyploidy and 

equivocal in the absence of activation. Micronucleus assays by Murli (1995, 1996d) were

negative.

A significant increase in 8-OH-dG liver levels, a biomarker for oxidative stress, was 

observed at ≥ 10 mg PFOA/kg in the liver but not the kidney of Fischer 344 male rats by Takagi 

et al. (1991). Work with HepG2 cells by Hu and Hu (2009) suggested that PFOA could induce 

apoptosis by overwhelming the homeostasis of antioxidative systems, increasing ROS, impacting 

mitochondria, and changing expression of apoptosis gene regulators. Eriksen et al. (2010) 

observed a PFOA-induced increase in ROS production in HepG2 cells, but no PFOA-induced 

oxidative DNA damage or cytotoxicity. 

3.4.3 Mode of Action and Implications in Cancer Assessment

The modes of toxicological/carcinogenic action of PFOA are not clearly understood. 

However, available data suggest that the induction of tumors is likely due to nongenotoxic 

mechanism involving membrane receptor activation, perturbations of the endocrine system, 

and/or the process of DNA replication and cell division. PFOA lacks the ability to react with and 

modify DNA, although its electrostatic properties would permit interaction with chromosomal 

histone proteins with a net positive surface charge.

Perfluorooctanoic acid (PFOA) – May 2016 3-152
Perfluorooctanoic acid (PFOA) – May 2016 3-153 

Rat Liver Tumors. PPARα agonism has been proposed as a potential MOA for the liver 

carcinomas and adenomas in rats following chronic PFOA exposure (Maloney and Waxman 

1999; Klaunig et al. 2003, 2012). In the PPARα agonism MOA, binding of PFOA to the PPARα 

receptor results in increased peroxisome proliferation and cell replication. PPARα is primarily 

expressed in the liver, but also is present in the kidney, intestines, heart, and brown adipose 

tissue (Hall et al. 2012). 

Peroxisomes are single-membrane organelles found in a number of plant and animal cells 

that have the capacity to carry out beta oxidation of long-chain fatty acids and other substrates 

through hydrogen peroxide-generating pathways and without the generation of ATP (Goodrich 

and Sul 2000). Peroxisomes metabolize the long-chain fatty acids via both beta and omega 

oxidation pathways (Fielding 2000), but are unable to metabolize fatty acid chains of eight 

carbons or less (Garrett and Grisham 1999). The shorter chain fatty acids are exported to the 

cytosol and taken up by mitochondria for further degradation via beta oxidation with resultant 

production of acetyl-CoA and ATP. 

When a chemical binds to and activates the PPARα receptor, it forms a heterodimer with the 

retinoid-X receptor and binds to the peroxisome proliferator response element found in the 

promoter region of selected genes (Spector 2000). In addition to a variety of xenobiotic 

chemicals, there are a number of endogenous substances in animals and humans that can activate 

the PPARα receptor, including unsaturated C18 fatty acids, metabolites of arachidonic acid, and 

the prostaglandin metabolite PGJ2 (Spector 2000). PPARα activation is accompanied by 

upregulation of many genes associated with catabolism of fatty acid and cholesterol biosynthesis 

and lipid transport (Hall et al. 2012; Rosen et al. 2008a).

There are four key events in the PPARα-agonist MOA for liver tumors (Klaunig et al. 2003, 

2012) (see Figure 3-1). The first key event is activation of PPARα. Increased palmitoyl-CoA 

oxidase activity is used in many studies as a biomarker for PPARα activations. Other associated 

indicators are hepatocellular hypertrophy and increased liver weight. However, these indicators 

alone are not sufficient to establish a PPARα MOA because they also are caused by chemicals 

that have no influence on PPARα.

The primary data that demonstrate PFOA activation of the PPARα receptor are those from 

Rosen et al. (2008a, 2008b) that examined the transcript profiles in the livers of wild-type and 

PPARα-null mice dosed with 1, 3, and 10 mg/kg/day PFOA for up to 7 days. The data from the 

wild-type mice were compared to those from the known PPARα gene activator Wyeth 14,643 

and PPARα-null mice. Based on the analysis of gene regulation, it was clear that PPARα 

activation was required for a majority of the transcriptional changes observed in the mouse liver 

following PFOA or Wyeth 14,643 exposure. The data from this study demonstrate the ability of 

PFOA to act as a PPARα agonist.

Multiple studies in both rats and mice provide evidence that PFOA induces peroxisome 

proliferation in the liver (Elcombe et al. 2010; Minata et al. 2010; Pastoor et al. 1987; Wolf et al. 

2008b; Yang et al. 2001). PFOA also was found to activate mouse and human PPARα using a 

transient transfection cell assay (Takacs and Abbott (2007). Maloney and Waxman (1999) also 

demonstrated that 5–10 μmol PFOA (2 to 4 mg/L) activated mouse PPARα using COS1 cells 

(kidney fibroblast-derived cells) transfected with a luciferase reporter gene.
Source: USEPA 2005c 

Figure 3-1. PPARα Agonist MoA for Liver Tumors

In rodents, hepatic physical and biochemical changes observed after activation are highly 

correlated with liver tumors leading to the hypothesis that a > 3-fold increase in peroxisomes and 

> 1.5 fold increase in liver weights in short-term studies are sufficient to cause liver cancers in 

long-term studies (Hall et al. 2012). The temporal and dose-response relationship of measures of 

peroxisome proliferation, hepatocellular hypertrophy, liver weight, and liver histopathology were 

examined in male Sprague-Dawley rats following 4, 7, and 13 weeks of administration of 

dietary PFOA at doses ranging from 0–6.5 mg/kg/day (Perkins et al. 2004). There was no 

evidence of peroxisome proliferation, hepatocellular hypertrophy, or liver weight increases at 

0.06 mg/kg/day. However, at 13 weeks, the 6.5-mg/kg/day dose had an increase in palmitoylCoA oxidase activity (an indicator for peroxisomes) that was 3.4 times greater than that for the 

pair-fed control. The absolute liver weight was 1.6 times greater than the pair-fed control. At the 

lower 1.94 mg/kg/day doses, the increases were 2.8 and 1.4 for the palmitoyl-CoA and liver 

weight, respectively

There are indications that PFOA also acts through PPARα-independent mechanisms 

associated with CAR and PXR receptors. Martin et al. (2007) examined the genomic signature 

from PFOA-treated Sprague-Dawley rats (up to 5-day exposure) using microarray expression 

profiling, pathway analysis, and quantitative PCR. The animal responses were consistent with 

PPARα agonism, but there was also evidence of PPARγ agonism (downregulation of cholesterol 

synthesis) and activation of CAR- and PXR-related genes. CAR activation can lead to hepatocyte 

proliferation and hepatocarcinogenesis in animals. However, the human CAR receptor is 

relatively resistant to mitogenic effects and less likely to induce cancers through this mechanism 

(Hall et al. 2012). In rodents, the PXR receptor can interact with PPARα in the coordination of 

Perfluorooctanoic acid (PFOA) – May 2016 3-154
hepatocyte proliferation, but there are differences in the amino acid composition of the ligand 

binding domain of the mouse receptor and the human receptor (10% homology) (Hall et al. 

2012). Accordingly, although the line of evidence is strongest for PPARα activation as the 

initiator for the downstream events in the PFOA cancer MOA, there can be involvement from 

other membrane receptors other than PPARα.

The second step in the PPARα MoA calls for evidence for increased cell proliferation and 

decreased apoptosis. Few studies examined the occurrence of these events with PFOA. Son et al. 

(2008) saw evidence of decreased apoptosis in liver and kidney cells stained for caspace3 in 

4-week-old male ICR mice treated for 21 days at a dose of about 20 mg/kg/day. However, 

Elcombe et al. (2010) failed to see a significant decrease in male Sprague-Dawley rats with a 

28-day exposure to a diet containing 300 ppm (~20 mg/kg/day) PFOA (comparable to the high 

dose in both cancer studies). In wild-type 129S4/SvlmJ mice, Minata et al. (2010) observed 

increased apoptosis in hepatocytes, arterial walls, and bile duct epithelium and in the bile duct 

epithelium of PPARα-null mice at 10.8 and 21.6 mg/kg PFOA. Thus, the apoptosis data for 

PFOA are not consistently supportive of the key step in this proposed MoA (i.e., a decrease in 

apoptosis). 

Using a BrdU labeling technique, Elcombe et al. (2010) observed significant increases in cell 

proliferation in male Sprague-Dawley rats after 1, 7, and 28 days of exposure to a 300-ppm 

PFOA dietary dose. The highest increase was observed after 7 days of treatment (a fivefold 

increase) and declined to a twofold increase after 28 days of dosing. The liver results from the 

Biegel et al. (2001) mechanistic study were negative for cell proliferation in male SpragueDawley rats exposed to the same dietary concentration (20 mg/kg/day) and sacrificed at 1, 3, 6, 

9, 12, and 15 months. However, based on the Elcombe et al. (2010) observations, the timing of 

the interim sacrifice would have missed the peak of the proliferative response. The Butenhoff et 

al. study (2012) identified hyperplastic nodules in 3/50 high-dose males and 2/50 high-dose 

females at 20 mg/kg/day; 5/50 males and 1/50 females had hepatocellular carcinomas.

The study by Wolf et al. (2008a) looked at the labeling index in 129S1/SvlmJ mice and 

PPARα-null mice and found a difference in their dose response. In the wild-type mice, the 

labeling index was increased at all doses ≥ 1 mg/kg/day; however, in PPARα-null mice, the 

labeling index was increased only at the highest dose, 10 mg/kg/day. 

There were no studies identified that focused specifically on preneoplastic foci and clonal 

expansion of altered cells after PPAR activation. Minata et al. (2010) observed a dose-dependent 

increase in eosinophilic cytoplasmic changes consistent with peroxisome proliferation in liver 

parenchyma, but found no focal necrosis at doses ≤ 21.6 mg/kg/day in wild-type 129S4/SvlmJ 

mice.

Klaunig et al. (2003, 2012) concluded that, based on the available data, PFOA-induced liver 

tumors in Sprague-Dawley rats can be attributed to a PPARα MOA since there are data available 

addressing most of the key steps in this proposed MoA. However, some data gaps exist for key 

events and other mechanisms that might be involved. Overall, the tumor response observed in the 

available studies was not strong and did not demonstrate a dose-related response in males (3/49, 

1/50, and 5/50 hepatocellular carcinomas in the control, 2-mg/kg/day, and 20-mg/kg/day dose 

groups, respectively) and a single carcinoma in females at the high dose. Biegel et al. (2001) did 

not identify any liver carcinomas (0/76) in males at their 20-mg/kg dose, but there were 10/76 

males with adenomas. This is consistent with a hyperplastic tissue response rather than the loss 

of cell cycle control characteristic of cancer. The data from the Butenhoff et al. (2012) and 

Perfluorooctanoic acid (PFOA) – May 2016 3-155
Biegel et al. (2001) studies suggest that PFOA is not a potent hepatic carcinogen based on the 

low tumor incidence and finding of hyperplastic nodules. 

Leydig Cell Tumors (LCT). LCTs were observed in both the Butenhoff et al. (2012) and Biegel 

et al. (2001) studies. The LCT incidence was 0/49, 2/50, and 7/50 at doses of 0, 2, and 

20 mg/kg/day, respectively, in Butenhoff et al. (2012) and 2/78 (pair-fed control) and 8/76 at 

20 mg/kg/day in Biegel et al. (2001). 

A large number of nongenotoxic compounds of diverse chemical structures have been 

reported to induce LCTs in rats, mice, or dogs. LCTs also occur in humans but are relatively rare 

at about 1–3% of human testicular tumors, which also are infrequent (1%) (Carpino et al. 2007). 

A workshop report (Clegg et al. 1997) on the MOA for LCT classified the chemicals that caused 

LCT in animal studies into seven MOA categories. The postulated MOAs support the following 

hormonal steps to the process:

1. A xenobiotic chemical inhibits the production of testosterone, leading to low serum 

levels. 

2. Low serum testosterone levels signal the hypothalamus to produce gonadotropin 

releasing hormone (GnRH). 

3. GnRH signals the pituitary to release LH. 

4. LH signals the Leydig cells to produce testosterone. 

5. LH causes Leydig cell proliferation. 

Several of the available PFOA studies support an impact of PFOA on decreased testosterone 

production. Studies conducted by Cook and colleagues (Biegel et al. 1995; Cook et al. 1992; Liu 

et al. 1996) found that adult male rats administered PFOA by gavage for 14 days had decreased 

serum testosterone and increased serum estradiol levels (Cook et al. 1992). These endocrine 

changes correlated with its potency to induce LCTs (Biegel et al. 2001). 

Subsequent experiments demonstrated that PFOA increased levels of estradiol by inducing 

cytochrome P450 CYP19 (aromatase). Aromatase converts androgens to estrogens, including the 

conversion of testosterone to estradiol. PFOA directly inhibits testosterone production when 

incubated with isolated Leydig cells and ex vivo studies demonstrate that this inhibition is 

reversible (Biegel et al. 1995). However, in the mechanistic bioassay by Biegel et al. (2001), 

serum testosterone and LH levels were not significantly altered at the levels of PFOA that 

resulted in LCTs (20 mg/kg/day). 

This inhibition of testosterone biosynthesis can be mediated by PPARα (Gazouli et al. 2002).

Support for PPARα-mediated inhibition of testosterone production is found in Li et al. (2011). 

Lower testosterone concentrations, reduced reproductive organ weights, and increased sperm 

abnormalities were found in PFOA-treated male PPARα wild-type and humanized PPARα mice 

but not in PPARα-null mice. Similarly, disruption of testosterone biosynthesis by lowering the 

delivery of cholesterol into the mitochondria and decreasing the conversion of cholesterol to 

pregnenolone and androstandione in the testis was noted in wild-type and humanized PPARα 

mice. These effects were not seen in PPARα-null mice. Decreased serum testosterone was noted 

after oral exposure to PFOA in studies by Biegel et al. (1995, 2001) and Cook (1992). 

The induction of LCTs by PFOA also can be attributed to a hormonal mechanism whereby 

PFOA either inhibits testosterone biosynthesis and/or lowers testosterone by increasing its 

conversion to estradiol through increased aromatase activity in the liver. Both of these 

mechanisms appear to be mediated by PPARα. However, data are not currently sufficient to 

Perfluorooctanoic acid (PFOA) – May 2016 3-156
demonstrate that the other key steps in the postulated MOA are present in PFOA-treated animals 

following exposures that lead to tumor formation. Studies are needed to demonstrate the increase 

of GnRH and LH in concert with the changes in aromatase and estradiol. There was also no 

indication of increased Leydig cell proliferation at the doses that caused adenomas in the Biegel 

et al. study (2001). Thus, additional research is needed to determine if the hormone testosteroneestradiol imbalance is a key factor in development of LCTs as a result of PFOA exposure. 

Two studies involving members of the C8 Health Project showed a positive association 

between PFOA levels (mean at enrollment 0.024 μg/mL) and kidney and testicular cancers 

(Vieira et al. 2013; Barry et al. 2013). This contributed to the EPA conclusion that PFOA can be 

classified as having suggestive evidence for carcinogenicity. 

Pancreatic Acinar Cell Tumors. The 2-year bioassay by Biegel (2001) identified PACTs in 7/6 

rats receiving a 20-mg/kg dose for 2 years compared to 1/79 in the pair-fed controls. As with 

LCTs, the MOA for PACTs is not understood. These tumors are most commonly identified in 

rats, but do occur in other animal species (e.g., mice, hamsters) and in humans (Wisnoski et al. 

2008). Males are more susceptible to pancreatic tumors than females. Two hypothetical MOAs 

have been proposed and are as follows (Klaunig et al. 2003, 2012; Obourn et al. 1997):

• There is a change in the bile acid flow or composition that leads to cholestasis, thereby 

causing an increase in CCK activating a feedback loop resulting in proliferation of the 

secretory pancreatic acinar cells. CCK is a peptide hormone that stimulates the digestion 

of fat and protein, causes the increased production of hepatic bile, and stimulates 

contraction of the gall bladder. An HFD, trypsin inhibition, and changes in bile 

composition are proposed initiators for this sequence of events. 

• Increased levels of testosterone support the growth of acinar cell preneoplastic foci, 

leading to the development of carcinomas. 

There is minimal information on the relationship of PFOA exposure to either of the proposed 

MOAs. Obourn et al. (1997) studied the impact of PFOA on CCK and trypsin using in vitro

assays and found that PFOA was not an agonist for the CCKA receptor that activates CCK 

release. PFOA also had no inhibitory action on trypsin at levels 1,000 times greater (0.31 μg/mL) 

than the positive control.

The Obourn et al. study (1997) also looked at Wyeth 14,643, a peroxisome proliferator, in 

these same assays and found results similar to those for PFOA. When they conducted an in vivo

study with 100 ppm Wyeth 14,643, they found a small but significant increase (p<0.05) in bile 

flow per unit body weight, a decrease (p<0.05) in bile flow per unit liver weight, and a small 

decrease (p<0.05) in the total bile acid concentration following a 6-month dietary exposure. 

There is the potential for PFOA to change the composition of bile because of its competition 

with bile acids for biliary transport. In mice, increased expression of MRP3 and MRP4 

transporters (Maher et al. 2008) and decreased expression of OATPs (Cheng and Klaassen 2008) 

favor excretion of PFOA into the bile. Minata et al. (2010) found the levels of PFOA in bile from 

wild-type male mice to be considerably higher than those in PPARα-null mice, suggesting a link 

to PPARα. In the same study, male wild-type and PPARα-null mice were orally dosed with ~0, 

5.4, 10.8, and 21.6 mg/kg/day of PFOA for 4 weeks. Total bile acid was significantly increased 

at the highest dose in PPARα-null mice suggesting that, in the presence of PFOA, activation of 

PPARα increases PFOA excretion, a scenario that could possibly decrease the flow of bile acids 

competing for the same transporters. In the Butenhoff et al. study (2012), there was a lack of 

Perfluorooctanoic acid (PFOA) – May 2016 3-157
PACT tumors but an increase in proliferative lesions of the acinar cells. One hypothesis offered 

for the difference in results was differences in the diets used in the two studies (Chang et al. 

2014). 

PFOA appears to suppress testosterone production through the induction of aromatase 

(Biegel et al. 1995; Cook et al. 1992; Liu et al. 1996) and to increase the estradiol. Accordingly, 

the second proposed MOA for PACTs does not appear to apply to PFOA. 

The data on a PPARα-linked MoA are strongest for the liver tumors. Some data also provide 

a link of PPARα to the Leydig cell and PACT tumors observed in the rat 2-year bioassays. They 

are not as strong and identify a need for additional research justifying the suggestive evidence 

finding. However, when integrated with the metabolic inertness of PFOA in animals and 

humans, a linear response to dose is not likely. This is consistent with the tumor data. Thus a 

nonlinear MOA is likely and the remaining challenge is to identify the critical event in each 

MOA that leads to development of the tumors. 

Other Potential Modes of Action. There are other potential MOAs that could apply to PFOA. 

They include interruption of intercellular communication, mitochondrial effects, and hormonal 

effects. None of these mechanisms are considered to be key steps in the MOAs discussed above.

GJIC, a process by which cells exchange ions, messages, and other small molecules, is 

important for normal growth, development, and differentiation as well as for maintenance of 

homeostasis in muticellular organisms. Because tumor formation requires loss of homeostasis 

and many tumor promoters inhibit GJIC, it has been hypothesized that GJIC might play a role in 

carcinogenesis (Trosko et al. 1998). PFOA has been demonstrated to inhibit GJIC in liver cells 

in vitro and in vivo (Upham et al. 1998, 2009). However, inhibition of GJIC is a widespread 

phenomenon, and the effect by PFOA was neither species- nor tissue-specific. In addition it was 

reversible. Thus, the significance of GJIC inhibition in regard to the mode of carcinogenic action 

of PFOA is unknown. 

Several chemicals structurally related to PFOA have been shown to manifest their toxicity by 

interfering with mitochondria biogenesis and bioenergetics. Walters et al. (2009) found evidence 

supporting mitochondrial proliferation in Sprague-Dawley rats receiving 30 mg/kg/day of PFOA 

for 28 days as reflected in measurements of mitochondrial DNA, transcription factors, and other 

biomarkers for mitochondrial effects. Dietary PFOA also was demonstrated to uncouple 

oxidative phosphorylation in mitochondria of the liver from rats exposed via their diet (Keller et 

al. 1992). At high concentrations, PFOA caused a small increase in resting respiration rate and 

slight decreases in the membrane potential. The observed effects were attributed to a slight 

increase in nonselective permeability of the mitochondria membranes caused by PFOA’s 

surface-active properties (Starkov and Wallace 2002). Quist et al. (2015) found evidence of 

mitochondrial proliferation in the liver of CD-1 mice pups from dams exposed to 1 mg/kg/day 

during gestation and lactation when tissues were examined using transmission electron 

microscopy at PND 21 and 91. 

3.4.4 Weight of Evidence Evaluation for Carcinogenicity 

The findings for cancer in humans provide support for an association between PFOA and 

kidney and testicular cancers; however, the number of independent studies examining each of 

these is limited. The support comes from high-exposure community studies examining cancer 

incidence and covering children and young adults (Barry et al. 2013; Vieira et al. 2013); there is 

Perfluorooctanoic acid (PFOA) – May 2016 3-158
some overlap in the cases included in these studies. The two occupational cohorts in Minnesota 

and West Virginia (most recently updated in Raleigh et al. 2014 and Steenland and Woskie 

2012) do not support an increased risk of kidney or testicular cancer, but are limited by a very 

small number of observed cases. None of the general population studies examined these cancers, 

but associations were not seen in the general population studies addressing colorectal, breast, 

prostate, bladder, and liver cancer, with mean serum PFOA levels up to 0.0866 μg/mL 

(Bonefeld-Jørgensen et al. 2014; Eriksen et al. 2009; Hardell et al. 2014; Innes et al. 2014).

The only chronic bioassays of PFOA were conducted in rats (Butenhoff et al. 2012; Biegel et 

al. 2001). The two studies support a positive finding for the ability of PFOA to be tumorigenic in 

one or more organs of male, but not female, rats. There are no carcinogenicity data from a 

second animal species. There are some data that provide support for the hypothesis that the 

PPARα agonism MOA is wholly or partially linked to each of the observed tumor types. The 

data support a PPARα MOA for the liver tumors and thus are indicative of lack of relevance to 

humans. PPARα activation also could play a role in the other tumor types observed, but more 

data to support intermediate steps in the proposed MOAs are needed.

The mutagenicity data on PFOA are largely negative, although there is some evidence for 

clastogenicity in the presence of microsomal activation and at cytotoxic concentrations. Given 

the chemical and physical properties of PFOA—including the fact that it is not metabolized, 

binds to cellular proteins, and carries a net negative electrostatic surface charge—the clastogenic 

effects are likely the result of an indirect mechanism. PFOA has the potential to interfere with 

the process of DNA replication because of its protein binding properties and the fact that histone 

proteins, spermine and spermidine, carry a net positive surface charge. Involvement of ROS in 

the MOA as a result of PFOA alone is unlikely because of its metabolic stability. Conditions 

leading to ROS would be a function of metabolic responses perturbed by PFOA rather than 

PFOA alone. A compound that is not metabolized will not be able to covalently alter the 

structure of DNA or intercalate because of electrostatic repulsion between the aromatic base pi 

bond electrons with the partial negative charges on the PFOA fluoride atoms. Due to its protein 

binding properties, PFOA could have an impact on one or more of the proteins involved in the 

process of DNA replication or cell division (cytoskeletal proteins); however, no mechanistic 

studies were identified that examined the biochemical effects of PFOA on DNA replication or 

cell division. There are no data that support the clastogenic MOA. 

Despite the limitations in the data for the LCTs and PACTs, under the U.S. EPA Guidelines 

for Carcinogen Risk Assessment (USEPA 2005a) there is suggestive evidence of carcinogenic 

potential of PFOA in humans. 

3.4.5 Potentially Sensitive Populations

Human biomonitoring studies do not suggest major differences between serum PFOA levels 

in males and females. However, the worker populations that are those most likely to demonstrate 

such differences because of their higher exposures were predominantly male.

Some animal species have gender differences that affect toxicity of PFOA. Sexually mature 

female rats excreted almost all of a 10-mg/kg dose of PFOA within 48 hours compared to only 

19% excreted by male rats. Male hamsters excrete PFOA faster than female hamsters, and 

female rabbits excrete PFOA slightly faster than male rabbits. Male and female mice excrete 

PFOA at approximately the same rate (Hundley et al. 2006). Studies of the transporters involved 

in the toxicokinetics of PFOA demonstrate that they are differentially impacted by the presence 

Perfluorooctanoic acid (PFOA) – May 2016 3-159
of male and female sex hormones influencing tissue persistence (Cheng et al. 2006; Kudo et al. 

2002). As studied in rats (Kudo et al. 2002), the male sex hormones increased half-life 

(decreased excretion) of PFOA while the female hormones were associated with shorter halflives (increased excretion). The gender differences in mice are not as pronounced as those in rats. 

Work by Cheng et al. (2006) and Cheng and Klaassen (2009) demonstrated that these hormones 

impact transporters in the liver and kidney. 

In studies in which both male and female rats were used, the males were more sensitive to 

toxicity than were the female rats (Butenhoff et al. 2004a). Mice displayed similar sensitivities 

following PFOA exposure (Kennedy 1987). In the monkey studies, the number of animals per 

gender per dose group was too small to reveal a difference related to gender.

Unfortunately, much work remains to be done to determine whether the gender difference 

seen in rats is relevant to humans. Similarities are possible because the long half-life in humans 

suggests that they might be more like the male rat than the female rat. There is a broad range of 

half-lives in human epidemiology studies, suggesting a variability in human transport and 

binding capabilities resulting from genetic variations in transporter structures and, consequently, 

in function. Genetic variation in human OATs and OATPs has been identified as described in a 

review by Zaïr et al. (2008).

Neonates, Infants, and Fetuses 

The developing fetus might be sensitive to effects of PFOA. The observed effects on birth 

weight in animals are supported by evidence of an association between PFOA and low birth 

weight in humans (Johnson et al. 2014). There is some uncertainty related to the interpretation of 

the small change in birth weight observed in humans. Specifically, it has been suggested that low 

GFR also can impact birth weight (Morken et al. 2014). Verner et al (2015) conducted a metaanalysis based on PBPK simulations and found that, in individuals with low GFR, there are 

increased levels of serum PFOA as well as lower birth weights. Thus, while there is some 

uncertainty in the interpretation of the observed association between PFOA and low GFR and 

birth weight given the available information, the data indicate that PFOA exposure does impact

birth weight in the general population. In humans with low GFR (which includes females with 

pregnancy-induced hypertension or preeclampsia) who also are exposed to PFOA, the effect on 

body weight is likely due to a combination of both. 

Several animal studies have examined potential MoAs for developmental effects following 

prenatal exposure to PFOA. PFOA exposure during development in rats and mice resulted in 

increased resorptions (mouse), increased fetal skeletal variation (rat, mouse), decreased neonatal 

survival (rat, mouse), decreased postnatal body weight (mouse), delayed eye opening and body 

hair growth (mouse), delayed vaginal opening (mouse), accelerated preputial separation (mouse), 

and delayed mammary gland development (mouse dam and offspring) (Abbott et al. 2007; 

Butenhoff et al. 2004a; Lau et al. 2006; Macon et al. 2011; Tucker et al. 2015; White et al. 2007, 

2009, 2011; Wolf et al. 2007). Other long-term effects observed in the surviving offspring 

included increased body weight gain, serum leptin, and serum insulin levels along with changes 

in adipose tissue (Hines et al. 2009). The MOAs for these developmental effects are unknown, 

but several potential MoAs have been investigated. 

Wolf et al. (2007) restricted mouse prenatal PFOA exposures to 3–11-day periods during 

gestation to determine if PFOA was affecting a certain stage of organogenesis resulting in the 

observed developmental effects. Decreased postnatal survival was observed at the highest dose 

Perfluorooctanoic acid (PFOA) – May 2016 3-160
used (20 mg/kg/day). Eye opening and body hair growth were delayed in offspring exposed for 

the longest periods of time (GD 7–17 and GD 10–17) and might have been the result of a higher 

cumulative dose or greater sensitivity during early gestation. A cross-fostering paradigm was 

used to determine if the developmental effects were the result of gestational exposure, lactational 

exposure, or a combination of both. Postnatal survival was decreased in offspring exposed 

through gestation and lactation (5 mg/kg/day). Eye opening and body hair growth were delayed 

and body weight was reduced in offspring exposed during gestation (5 mg/kg/day), and gestation 

and lactation (3 and 5 mg/kg/day). No developmental delays in eye opening and body hair 

growth were observed in offspring exposed via lactation only, indicating that, for these 

developmental endpoints, PFOA alters growth regulation in the developing fetus that persists as 

growth continues postnatally. 

Both gestational and lactational exposures contribute to the impact of PFOA on body weight 

during early life as illustrated by cross-fostering control unexposed female pups with those dosed 

with PFOA. Three cross-fostering combinations were evaluated by White et al. (2009): control 

pups nursed by exposed dams, exposed pups nursed by control dams, and exposed pups nursed 

by exposed dams. Two doses were evaluated: 3 and 5 mg/kg/day, but the body weight data was 

only provided for the 5-mg/kg/day dose group for PND 1–10. PFOA exposures significantly 

reduced pup body weights and increased liver weights. The body weight deficits compared to 

control were greatest for the gestation and lactation exposure combination and lowest for the 

lactation-only group. 

Abbott et al. (2007) examined activation of PPARα as a factor in the developmental toxicity 

of PFOA. Wild-type and PPARα-null mice experienced full litter resorptions following 

gestational (GD 1–17) PFOA exposure (≥5 mg/kg/day), indicating that the mechanism of PFOAinduced resorptions was independent of PPARα expression. These resorptions could be due to 

insufficient trophoblast cell type differentiation and/or increased trophoblast cell necrosis (Suh et 

al. 2011). Postnatal survival was significantly decreased in wild-type offspring but not in 

PPARα-null offspring, indicating that PPARα expression was required for postnatal lethality

(Abbott et al. 2007). Eye opening was significantly delayed in wild-type offspring, but not in 

PPARα-null offspring, although a trend was observed in those offspring for later eye opening. 

The results indicated that PPARα expression was important for eye opening, but other PPARαindependent factors also might play a role in its mechanism. Takacs and Abbott (2007) showed 

that PFOA can activate mouse PPARβ/δ, which is expressed in developing tissue, and suggested 

that inappropriate activation of PPARβ/δ could cause adverse effects. Further research needs to 

be conducted to fully elucidate the mechanism. 

Mouse mammary gland development was another endpoint examined in prenatally PFOAexposed offspring. White et al. (2007) found that dams dosed with 5 mg PFOA/kg/day on GD 

1–17 and GD 8–17 had significantly delayed mammary gland development (full of alveoli, 

visible adipose tissue, not well differentiated) at PND 10, which is at the peak of lactation in 

rodents. The delayed dam mammary gland development could play a role in the observed 

reduced offspring body weight gain if the quantity or quality of the milk is altered by PFOA 

(Abbott et al. 2007; Lau et al. 2006; White et al. 2007; Wolf et al. 2007). 

Restricted gestational exposure and cross-fostering studies showed that delayed offspring 

mammary gland development observed PND 1–63 occurred regardless of exposure duration or 

timing (gestation versus lactation exposure; maternal dose of 1 mg/kg/day). The developmental 

delays persisted even as the internal PFOA dose decreased (Macon et al. 2011; White et al. 2007, 

2009, 2011). More studies need to be conducted to elucidate the MOA for dam and offspring 

Perfluorooctanoic acid (PFOA) – May 2016 3-161
mammary gland effects and its potential functional consequences for lactating humans. White et 

al. (2011) conducted a multigeneration study of the effects of PFOA on mammary gland 

development and found no dose-related effects on the pup body weights nourished by dams with 

lower mammary gland scores than the controls. Tucker et al. (2015) demonstrated that a doseresponse for developmental mammary gland effects varies by more than an order of magnitude, 

depending on the strain of mouse studied. CD-1 mice are more sensitive than C57BL/6 mice 

(Tucker et al. 2015). 

Mammary gland development also was affected by peripubertal exposure to PFOA (C. Yang 

et al. 2009, Y. Zhao et al. 2010). Low doses (5 mg/kg/day) of PFOA from 3 to 7 weeks of age 

caused accelerated mammary gland development in C57BL/6 mice, but delayed mammary gland 

development in BALB/c mice, suggesting strain-related differences. 

Experiments examining the mechanism for accelerated mammary gland development showed 

that PFOA promotes steroid hormone production in the ovaries and increases the levels of 

several mammary gland growth factors in C57BL/6 wild-type and PPARα-null mice. The 

mechanism for delayed mammary gland development following a peripubertal PFOA exposure 

needs to be examined. 

Hines et al. (2009) found that low doses of PFOA given during gestation to CD-1 mice 

resulted in significant weight gain and increased serum insulin and leptin levels of the offspring 

in mid-life. The increased leptin levels, as well other hormone perturbations, might place PFOA 

into the environmental endocrine disruptor obesogen category similar to diethylstilbestrol 

(Newbold et al. 2007). However, in a study by Quist et al. (2015) using the mature animals from 

the Hines et al. study (2009), there was no dose-related impact on serum leptin in CD-1 pups 

gestationally exposed across a dose range of 0–1 mg/kg/day when examined on PND 91, except

in the group given an HFD and not fasted before serum collection. For those animals, there was a 

dose-related decrease in leptin. Other mice on an HFD that were fasted for 4 hours before serum 

collection in the same study lacked a dose-related leptin response. In humans, increased leptin 

levels are associated with increased body fat and suggestive of a leptin-resistance MOA for being 

overweight (Considine et al. 1996). A similar relationship might occur in prenatally PFOAexposed mice; however, the Quist et al. study (2015) suggests that the fat content of the diet and 

the time of serum collection are important variables that need to be considered. Studies 

determining MOAs need to be conducted to determine relevance of the mammary gland effects 

to animal and human health. 

Diet might influence the risk associated with PFOA exposures. Animal studies demonstrate 

an increased risk for liver steatosis in animals on an HFD (Quist et al. 2015; Tan et al. 2013) and 

possibly for insulin resistance (Hines et al. 2009). The epidemiology data are not supportive of a 

correlation with insulin resistance, but the observations of elevated serum triglycerides, 

especially among a highly exposed population, could be viewed as a risk factor for steatosis. 

Most of the epidemiology studies did not evaluate dietary factors as part of the study design for 

either birth weight or serum lipids (e.g., cholesterol, triglycerides, LDL).

Perfluorooctanoic acid (PFOA) – May 2016 3-162
4 DOSE-RESPONSE ASSESSMENT

4.1 Dose-Response for Noncancer Effects

An RfD or reference concentration (RfC) is used as a benchmark for the prevention of longterm toxic effects other than carcinogenicity. RfD/RfC determination assumes that thresholds 

exist for toxic effects, such as cellular necrosis, significant body or organ weight changes, blood 

disorders, and so forth. The RfD is expressed in terms of mg/kg/day and the RfC is expressed in 

milligrams per cubic meter (mg/m3). The RfD and RfC are estimates (with uncertainties 

spanning perhaps an order of magnitude) of the daily exposure to the human population 

(including sensitive subgroups) that is likely to be without an appreciable risk of deleterious 

effects during a lifetime.

4.1.1 RfD Determination

The derivation of the RfD for PFOA presented a number of challenges due to the 

toxicokinetic complexity of PFOA, variability in half-life between species, and metabolic 

inertness of PFOA in living organisms. The toxicokinetic features of PFOA lead to differences in 

half-lives across species and in the case of rats, and possibly humans, differences between 

genders. Toxicokinetics also influence intraindividual and lifestage variability in response to 

dose. Additionally there were inconsistencies across the epidemiology studies and the effects 

observed in animal studies, and a number of animal studies lacked a NOAEL. Each of these 

factors highlights the importance of having measures of internal dose for quantification of an 

RfD and supports the utilization of a PK model as a component of the dose-response assessment.

Human Data. Key studies examined occupational and residential populations at or near largescale PFOA production plants in the United States in an attempt to determine the relationship 

between serum PFOA concentration and various health outcomes suggested by the standard 

animal toxicological database. Health outcomes assessed include blood lipid and clinical 

chemistry profiles, reproductive parameters, thyroid effects, diabetes, immune function, birth and 

fetal and developmental growth measures, and cancer. 

Epidemiology studies examined workers at PFOA production plants, a high-exposure 

community population near a production plant in the United States (i.e., the C8 cohort), and 

members of the general population in the United States, Europe, and Asia. These studies 

examined the relationship between serum PFOA concentration (or other measures of PFOA 

exposure) and various health outcomes. Exposures in the highly-exposed C8 community are 

based on the concentrations in contaminated drinking water and serum measures. Exposures 

among the general population typically included multiple PFAS as indicated by serum 

measurements. The correlation among these compounds is often moderately strong 

(e.g., Spearman r > 0.6 for PFOA and PFOS in the general population). Mean serum levels 

among the occupational cohorts ranged approximately 1–4 μg/mL and in the C8 cohort they 

ranged 0.01–0.10 μg/mL. Geometric mean serum values for the NHANES general population 

(≥ age 12; 2003–2008) were 0.0045 μg/mL for males and 0.0036 μg/mL for females (Jain 2014).

These studies have generally found positive associations between serum PFOA concentration 

and TC (i.e., increasing lipid level with increasing PFOA) in the PFOA-exposed workers at mean 

serum levels 0.4 to >12 μg/mL and the high-exposure community at mean serum about 

0.08 μg/mL; similar patterns are seen with LDLs but not with HDLs. These associations also 

Perfluorooctanoic acid (PFOA) – May 2016 4-1 
were seen in most of the general population studies (mean serum 0.002–0.007 μg/mL), but 

similar results were seen with PFOS and the studies did not adjust for these correlations. 

Associations between PFOA exposure and elevations in serum levels of ALT and GGT, were 

consistently observed in occupational cohorts, the high-exposure community, and the U.S. 

general population at serum PFOA concentrations also associated with increased TC. The 

associations are not large in magnitude, but they indicate the potential to affect liver function. 

Thyroid disease incidence was associated with PFOA in women and girls in the highexposure C8 study population and in women with background exposure at mean serum 

concentrations of 0.026–0.123 μg/mL. Changes in thyroid hormones were not consistently 

associated with PFOA concentration.

Associations between PFOA exposure and risk of infectious diseases (as a marker of immune 

suppression) have not been found, but a decreased response to vaccines in relation to PFOA 

exposure was reported in studies in adults in the high-exposure community population (median 

0.032 μg/mL) and in studies in children in the general population (mean 0.004 μg/mL). In the 

latter studies, it is difficult to distinguish associations with PFOA from those of other correlated 

PFAAs. Increased risk of ulcerative colitis was reported in the high-exposure community study 

and in a study limited to workers in that population. 

Studies in the high-exposure community reported an association between serum PFOA at 

approximately 0.01–0.02 μg/mL and risk of pregnancy-related hypertension or preeclampsia. 

This outcome has not been examined in other populations. An inverse association between 

maternal PFOA (measured during the second or third trimester) or cord blood PFOA 

concentrations and birth weight was seen in several studies, but the magnitude was small. It has 

also been suggested that low GFR can impact birth weight (Morken et al. 2014). Verner et al. 

(2015) conducted a meta-analysis based on PBPK simulations and found that some of the 

association reported between PFOA and birth weight could be partially attributable to low GFR. 

However, the study authors demonstrated that in individuals with low GFR there also are 

increased levels of serum PFOA. Thus, while there is some uncertainty in the interpretation of 

the observed association between PFOA and low GFR and birth weight given the available 

information, the data indicate that PFOA exposure does impact birth weight in the general 

population. 

The epidemiology studies have not found associations between PFOA and diabetes, 

neurodevelopmental effects, or preterm birth and other complications of pregnancy. 

Developmental outcomes including delayed puberty onset in girls has been reported; however, in 

the two studies examining prenatal PFOA exposure in relation to menarche, conflicting results 

were found (i.e., no association or a possible indication of an earlier menarche seen with higher 

maternal PFOA levels in one study and a later menarche seen with higher maternal PFOA levels 

in the other study). 

Animal Data—Long Term Studies. Some of the effects in animal studies are associated with 

activation of the PPARα receptor leading to peroxisome proliferation. These include increased 

liver weight; decreases in serum triglycerides, cholesterol, and lipoproteins; and increases in 

ALT, AST, or both. However, although the mechanisms for other effects, such as decreased 

body weight, immunological effects, and developmental delays are unknown, they might be 

relevant to human health risk assessment.

Perfluorooctanoic acid (PFOA) – May 2016 4-2 
As an initial step in the dose-response assessment, EPA identified a suite of animal studies 

with NOAELs, LOAELs, or both that identified the studies as candidates for development of a 

chronic RfD. These studies are listed in Table 4-1. The candidate studies were selected based on 

their NOAEL, LOAEL, or both; a duration of ≥ 7 weeks; use of a control; and two or more 

doses. Table 4-1 does not include the data from human epidemiology studies because, although 

they include information on serum levels, they do not identify exposure sources or external 

doses.

Table 4-1. NOAEL/LOAEL Data for Oral Subchronic and Chronic Studies of PFOA

Species

Study 

Duration

NOAEL

mg/kg/day

LOAEL

mg/kg/day Critical Effects (s) Reference

Monkey

Male

90 days none 3 ↑ relative pituitary weight Goldenthal 1978

Monkey

Female

90 days 3 10 ↓ absolute and relative heart weight Goldenthal 1978

Monkey

Male

26 weeks none 3 ↑ absolute liver weight 

(hepatocellular hypertrophy) and 

mean liver-to-body weight 

percentages

Butenhoff et al. 

2002

Rat

Male

13 weeks 0.64 1.94 ↑ absolute and relative liver weight 

with hepatocellular hypertrophy 

accompanied by a slight, but not 

significant, increase in hepatic 

coagulative necrosis

Perkins et al. 

2004

Rat Male

F0

generation

84 days none 1 ↑ absolute and relative liver and 

kidney weight accompanied by ↓ 

body weight

Butenhoff et al. 

2004a; York et 

al. 2010

Rat Male

F1 generation

16 weeks none 1 ↓ body weights and weight gains; ↑ 

absolute and relative liver weights, 

liver hypertrophy; ↑ absolute and 

relative kidney weights

Butenhoff et al. 

2004a; York et 

al. 2010

Rat Female

F0 generation

127 days 30 none No significant effects observed Butenhoff et al. 

2004a; York et 

al. 2010

Rat Female

F1 generation

10 weeks 10 30 Delay in sexual maturity, ↓ body 

weight and weight gain

Butenhoff et al. 

2004a; York et 

al. 2010

Rat

Male and 

Female

2 years 1.3 (m)

1.6 (f)

14.2 (m)

16.1 (f)

M: ↓ body weight gain; 

histopathology lesions in liver, testes, 

and lungs.

F: ↓ body weight gain 

Butenhoff et al. 

2012

When examining the effects associated with the LOAELs summarized in Table 4-1, changes 

in relative liver weight, absolute liver weight, or both appear to be a common denominator for 

monkeys and rats (Butenhoff et al. 2002, 2004a; Perkins et al. 2004) with or without other 

hepatic indicators of adversity. Serum PFOA levels, where available, associated with increased 

liver weight were 81 and 41 μg/mL for the male monkey and rat, respectively. However, the 

increases in liver weight and hypertrophy are effects associated with activation of cellular 

PPARα receptors, making it difficult to determine whether this change is totally a reflection of 

the PPARα activation or PFOA toxicity and meet the Hall et al. (2012) criteria for establishing 

adversity for a PPARα-activating chemical. Studies in PPARα null mice and animals with a 

Perfluorooctanoic acid (PFOA) – May 2016 4-3 
human PPARα receptor (Li et al. 2011; Minata et al. 2010; Nakamura et al. 2009; Wolf et al. 

2008b), along with studies of hepatic gene activation by PFOA (Albrecht et al. 2013; Bjork and 

Wallace 2009; Nakamura et al. 2009; Rosen et al. 2008a, 2008b), suggest that the increase in 

liver weight is at least partially due to cellular impacts that are not controlled by PPARα 

receptors. However, it remains difficult to separate the impact of PPARα activation from the 

direct effects of PFOA in the candidate studies.

According to Hall et al. (2012), increases in liver weight can be considered adverse when 

accompanied by cellular necrosis, inflammation, fibrosis of the liver, and/or macrovesicular 

steatosis. There was some evidence of hepatic necrosis in the studies of Perkins et al. (2004) and 

in the male F1 generation adult rats from the Butenhoff et al. study (2004a), but the incidences 

were not statistically significant or described in detail. To the extent that adverse lesions reflect 

sensitivity in the animals impacted, they are used in the assessment to reflect that the liver 

hypertrophy and increased liver weight are adverse in individual animals where they are 

accompanied by necrosis.

Body weight effects were seen in several studies (Butenhoff et al. 2004a, 2012) and are a 

more toxicologically-relevant endpoint, especially in the cases where they were not accompanied 

by decreased food intake and when found in neonates (Butenhoff et al. 2004a). There were 

developmental delays for males and females in the two-generation study published by Butenhoff 

et al. (2004a). Testicular effects were observed by Butenhoff et al. (2012) and in the chronic onedose study by Biegel et al. (2001). There was evidence of increased kidney weight in male F1 

Sprague-Dawley rats (Butenhoff et al. 2004a) confounded by decreases in body weight at higher 

doses, but at lower doses the kidney weight effect is likely a reflection of tissue adaptation as a 

result of the requirement for upregulation of tubular transporters to facilitate urinary excretion 

using transporters developed for excretion of endogenous and dietary substrates rather than 

PFOA.

Four of the longer term studies in Table 4-1 lack a NOAEL and have LOAELs that range 

1–3 mg/kg/day. The NOAELs for the remaining 7 studies range from 0.64 (male rats) to 

30 mg/kg/day (female rats). Male monkeys and rats appear to respond at doses that are lower 

than their female counterparts. No long-term studies in mice were identified. Since NOAELs and 

LOAELs are to some extent the product of concentration or dose level selection, examination of 

the dose information in Table 4-1 suggests that several of the data sets that have serum data to 

inform modeling of internal doses have the potential to be co-critical in the dose-response 

evaluation. 

Animal Data—Short Term Studies. A number of studies identified adverse effects following 

low dose exposures over durations of 7 to 38 days. The studies fall into two clusters, those 

evaluating developmental or reproductive effects and those with a focus on immunological 

effects. The critical shorter-term studies in rats and mice are summarized in Table 4-2. Although 

the exposure duration is shorter in developmental studies than the two-generation study, the 

developmental studies are important in quantification of dose-response because the exposures 

occur during critical windows of development and predicate effects that can occur later in life. 

Perfluorooctanoic acid (PFOA) – May 2016 4-4 
Table 4-2. Shorter-term and Developmental Oral Exposure Studies

Species

Study 

Duration

NOAEL

(mg/kg/day)

LOAEL

(mg/kg/day) Critical Effect(s) Reference

Rat

Male 29 days 1 10 Increased absolute and relative 

liver weight, focal liver necrosis

Loveless et al. 

2008

Male 14 days 0.2 2 ↑ liver weight, ↑ serum estradiol 

and hepatic aromatase

Liu et al. 1996

Mouse

Female 

offspring

17 days none 1 Delayed mammary gland 

development in dams during 

lactation

White et al. 

2011

Male &

Female

38 days 1 5 ↑ Ito cell hypertrophy at 18 

months. Dosing occurred during 

gestation and lactation only.

Filgo et al. 

2015

Female 38 days 0.01 0.03 ↑ TC at PND 91 for fasted and 

nonfasted animals receiving a HFD 

but not those receiving the 

standard fat content control diet. 

Exposure occurred only during 

gestation and lactation.

Quist et al. 

2015

Male 29 days 0.3 1 ↑ absolute and relative liver 

weight, ↓ relative spleen weight, 

moderate-severe liver hypertrophy 

with single cell and focal necrosis

Loveless et al. 

2008

Male 28 days none 5 Significantly ↓ fertility based on 

pregnant females per male mouse, 

and ↓ pup birth weight. 

Lu et al. 2015

Female 17 (pups) 

/18 (dams) 

days

none 1 ↑ absolute maternal liver weight, ↓ 

ossification (calvarin, enlarged 

fontanel), accelerated onset of 

puberty in male offspring.

Lau et al. 2006

Female 17 days

GDs 1–17

none 3 ↑ absolute and relative maternal 

liver weight, delayed offspring eye 

opening and body hair growth, ↑ 

offspring relative liver weight, ↓ 

offspring body weight, delayed 

mammary gland development 

(female offspring).

White et al. 

2009; Wolf et 

al. 2007

Female

CD1

17 days

GDs 1–17

none 0.3 Delayed mammary gland 

development

Macon et al. 

2011

Female CD-1 17 days none 0.01 Delayed mammary gland 

development at PND 56. Exposure 

occurred only during gestation.

Tucker et al. 

2015

Female

C57BL6 

17 days 0.1 0.3 Delayed mammary gland 

development at PND 61. Exposure 

occurred only during gestation.

Tucker et al. 

2015

Female 15 days 1.88 3.75 ↓ IgM (1 day post-dose), increased 

IgG (15 days post-dose), ↓absolute 

and relative spleen weight (1 day 

post-dose)

DeWitt et al. 

2008

Perfluorooctanoic acid (PFOA) – May 2016 4-5 
Species

Study 

Duration

NOAEL

(mg/kg/day)

LOAEL

(mg/kg/day) Critical Effect(s) Reference

Male 15 days 7.5 30 ↓ sheep red blood cell IgM 

response in PPAR null mice 

indicate the response not 

completely PPAR dependent

DeWitt et al. 

2015

Male 14 days 2.5 5 ↓ sperm count, changes in 

testicular morphology, evidence of 

↑ free radical oxidation

Liu et al. 2015

Male 14 days 0.2 2 ↑ liver weight, serum estradiol and 

hepatic aromatase activity

Liu et al. 2015

Female 11 days

GDs 7–17

none 5 ↑ maternal and pup relative liver 

weight, delayed offspring eye 

opening and hair growth, ↓ male 

offspring body weight, delayed 

mammary gland development 

(female offspring)

White et al. 

2009; Wolf et 

al. 2007

Female CD 1 8 days

GDs 10–17

none 0.01 Delayed mammary gland 

development on PND 21 (female 

offspring)

Macon et al. 

2011

Female CD 1 17 days

GDs 1–17

none 0.3 Delayed mammary gland 

development on PND 14 (female 

offspring)

Macon et al. 

2011

All but two of the short term studies used mice as the target species. Mice differ from rats in 

that the toxicokinetics of the males and females are similar. The half-life of PFOA in male rats is 

much longer than that in females, favoring higher serum levels in males after equivalent 

exposures. The difference in the excretion kinetics is a consequence of differences in renal 

transporters between male and female rats that appear to be under hormonal control. Several of 

the short term studies include serum data to support PK modeling of internal dose-response 

(DeWitt et al. 2008; Lau et al. 2006; Macon et al. 2011). 

As was the case with the longer-term studies, increased liver weight was a common effect 

among the shorter-term studies. Increases in absolute or relative liver weights were reported in 

six of the studies that provided dose-response data from short term exposures (Lau et al. 2006; 

Liu et al. 1996, 2005; Loveless et al. 2008; White et al. 2009; Wolf et al. 2007) (Table 4-2). In 

some of the remaining studies, liver weight was not monitored as a variable. The Loveless et al. 

study (2008) identified significant focal liver necrosis in rats at the 10 mg/kg/day LOAEL, and 

both single cell and focal liver necrosis in mice at a LOAEL of 1 mg/kg/day. This might indicate 

that mice are more susceptible to necrosis than rats. Hepatic necrosis was reported in the longer 

duration Perkins et al. study (2004) of male rats and in the male F1 generation adult rats from the 

Butenhoff et al. study (2004a), but hepatic necrosis was present in few animals and not evaluated 

for statistical significance.

The co-occurring effects at the LOAEL were effects on spleen, thymus, liver, and/or 

developmental endpoints. Four of the studies involved exposures that occurred only during 

gestation and lactation and resulted in effects that were observed in the mature offspring. The 

hepatic and serum cholesterol effects in Quist et al. (2015) at PND 91 at a LOAEL of 

0.03 mg/kg/day were present only in animals with elevated intakes of dietary fat. In adult 

animals with the same gestation/lactation only exposures, Filgo et al. (2015) identified a LOAEL 

of 5 mg/kg/day for accumulation of fat deposits in the liver Ito cells (steatosis). The study did not 

Perfluorooctanoic acid (PFOA) – May 2016 4-6 
provide information on the intakes of dietary lipid that could be compared with the data from 

Quist et al. (2015) to determine whether the effects were correlated with postnatal dietary fats or 

from the exposure during gestation and lactation. 

A NOAEL of 2.5 mg/kg/day and a LOAEL of 5 mg/kg/day was reported for sperm counts 

and testicular morphology after a14-day exposure by Liu et al. (2015). No impact on male 

fertility was observed in Sprague-Dawley rats in the two-generation Butenhoff et al. study 

(2004a), in contrast to the Lu et al. study (2015) where male fertility was decreased in mice at a 

dose of 5 mg/kg/day for 28 days. A 14-day exposure to 2 mg/kg/day (Liu et al. 2015) led to 

significantly increased serum estradiol and increased hepatic aromatase activity.

The developmental impacts of PFOA exposure ranged from delayed mammary gland 

development in pups (Albrecht et al. 2013; Macon et al. 2011; Tucker et al. 2015; White et al. 

2009, 2011; Wolf et al. 2007) to delays in attaining developmental milestones such as 

ossification, eye opening, and hair growth (Lau et al. 2006; Wolf et al. 2007). The LOAEL for 

developmental effects on mammary glands in female offspring from dams given 0.01 mg/kg/day 

for 8 days from Macon et al. (2011) is of unknown biological significance. In the same study, no 

effects on offspring body weight were found at maternal doses up to 3 mg/kg/day for 17 days 

(Macon et al. 2011). Data from White et al. (2011) showed no significant effects on body weight 

gain in pups nursing from dams treated with 1 mg/kg/day despite these dams having less fullydeveloped mammary glands compared to controls. Similarly, no differences in response to 

lactational challenge were seen in PFOA-exposed dams with morphologically delayed mammary 

gland development (White et al. 2011). 

The studies that looked at the delays in other developmental milestones including eye 

opening, hair growth, and bone ossification all lacked a NOAEL. In Lau et al. (2006), the 

LOAEL was 1 mg/kg/day for reduced ossification of the proximal phalanges (forelimb and 

hindlimb). In the Wolf et al. (2007) study, delays in eye opening and hair growth occurred at a 

LOAEL of 5 mg/kg/day for gestational exposures of both 1–17 days and 7–17 days. Attainment 

of sexual maturity in males from the Lau et al. study (2006), rather than being delayed, was 

accelerated, at the LOAEL of 1 mg/kg/day. 

The LOAEL for the mammary gland developmental effects in female offspring from dams 

given 0.01 mg/kg/day for 8 days from Macon et al. (2011). In the same study, no effects on 

offspring body weight were found at maternal doses up to 3 mg/kg/day for 17 days (Macon et al. 

2011). Data from White et al. (2011) showed no significant effects on body weight gain in pups 

nursing from dams treated with 1 mg/kg/day despite these dams having less fully-developed 

mammary glands compared to controls. Similarly, no differences in response to lactational 

challenge were seen in PFOA-exposed dams with morphologically-delayed mammary gland 

development (White et al. 2011). Given that milk production was sufficient to nourish growth in 

exposed pups, there is uncertainty related to the functional impact of this endpoint and thus it 

was not considered quantitatively.

A NOAEL of 2.5 mg/kg/day and a LOAEL of 5 mg/kg/day were reported for sperm counts 

and testicular morphology after a 14-day exposure by Liu et al. (2015). No impact on male 

fertility was observed in Sprague-Dawley rats in the two-generation Butenhoff et al. study 

(2004a) in contrast to the Lu et al. study (2015) where male fertility was decreased in mice at a 

dose of 5 mg/kg/day for 28 days. A 14-day exposure to 2 mg/kg/day (Liu et al. 2015) lead to 

significantly-increased serum estradiol and increased hepatic aromatase activity.

Perfluorooctanoic acid (PFOA) – May 2016 4-7 
The studies by DeWitt et al. (2008, 2015) demonstrate effects of PFOA on spleen and 

thymus weights and the immunoglobulin response to SRBC or dinitrophenyl ficol in wild-type 

and PPARα null mice. The DeWitt et al. (2015) data indicate that some but not all of the 

response is related to PPARα activation. As supported by the epidemiology data, suppression of 

the immune system response to a challenge because of PFOA exposure is an area of concern for 

humans as well as animal species.

Six of the twelve studies lacked a NOAEL. For those studies with a NOAEL, the value 

ranged 0.01–7.5 mg/kg/day, while the LOAELs ranged 0.01–30 mg/kg/day. The range of values 

across studies is narrow, with overlap between the NOAELs and LOAELs. In all instances, the 

durations of exposure in shorter-term studies were less than 39 days, suggesting that 

physiological responses to PFOA occur early in the exposure continuum and at doses, but not 

necessarily average serum levels, comparable to those observed in the long term studies.

4.1.1.1 PK Model approach

In linking chemical exposure to toxic endpoints, careful consideration of PKs is crucial. This 

is especially true for PFOA, where inter-species and gender variation in CL half-life can vary by 

several orders of magnitude. If the toxicological endpoints are assumed to be driven by internal 

concentrations, the internal exposure needs to be calculated and considered across species. 

Differences in PKs (e.g., male rats excrete PFOA more slowly than females) and differences 

across species produce differences in the external dose needed to achieve the same internal dose. 

The use of the animal data and the available PK model allows for the incorporation of species 

differences in saturable renal resorption, dosing duration, and serum measurements for doses 

administered to determine HEDs based on average serum concentration and CL. 

Candidate studies for RfD development were restricted to those of adequate duration 

(preferably > 7 weeks), multiple dose groups, use of a concurrent control, and with serum data 

amenable for modeling that showed the most sensitive effects following exposure to PFOA. 

Those studies included the subchronic study by Perkins et al. (2004), the two-generation study by 

Butenhoff et al. (2004a), both conducted in rats, and the Butenhoff et al. study (2002) in 

monkeys. Also included are the developmental studies of Lau et al. (2006) and Wolf et al. 

(2007)/White et al. (2009), and the DeWitt et al. study (2008) of immunotoxicity in mice that 

showed effects of lifetime concern despite their briefer exposure durations. Together these 

studies encompassed the range of doses evaluated and the LOAELs observed in other studies that 

lacked serum data.

The Butenhoff et al. study (2002) was included as it is the only longer-term study in a 

nonhuman primate and had serum PFOA data available. The dose of 3 mg/kg/day was a LOAEL 

for increased liver weight in the absence of clear adverse effects. The small number of animals 

per group (2–4) made evaluation of accompanying liver effects difficult to evaluate against the 

Hall et al. (2012) criteria. Thus, while included in the model for comparison of serum levels, data 

from Butenhoff et al. (2002) will not be considered further in the quantitative RfD assessment.

PFOA has dose-dependent kinetics. Although repeated doses rapidly result in quasiequilibrium blood concentrations, a single dose results in a much longer half-life than is 

consistent with a rapid approach to quasi-equilibrium (Andersen et al. 2006; Lou et al. 2009). 

Using a simple, linear PK model (e.g., a one- or two-compartment model) to predict internal 

dose resulted in estimated HEDs from repeated exposures that were greater than the external 

doses supported by the experimental data (Butenhoff et al. 2004a; Lou et al. 2009). Application 

Perfluorooctanoic acid (PFOA) – May 2016 4-8 
of a saturable renal resorption model (Andersen et al. 2006) predicted average serum values at 

the time of sacrifice and the duration necessary to reach steady state.

PK data (serial blood concentrations following treatment with known quantities of PFOA) 

were collected for three species: cynomolgus monkey (Butenhoff et al. 2004b), Sprague-Dawley

rat (Kemper 2003), and mice. Data were available for two strains of mice and these were 

analyzed separately: CD1 (Lou et al. 2009) and C57BL6/N (DeWitt et al. unpublished, cited in 

DeWitt et al [2008]). Due to the pronounced difference in the PKs of male and female rats, the 

two genders were fit separately. For mice, only female data were used. For monkeys a limited 

amount of female data was used jointly with male data, assuming the only difference between the 

genders for monkey was bodyweight. 

For each study with a toxicological endpoint and LOAEL, the AUC and final serum 

concentrations were determined for the exposure duration investigated in that study. These 

values are summarized in Table 4-3 for rats, Table 4-4 for mice, and Table 4-5 for monkeys. In 

order to make a rough assessment of the validity of the model predictions, a final serum 

concentration was predicted for each treatment so that it could be compared to measured serum 

values. The predicted final serum concentration is the estimate for serum concentration at the 

time of sacrifice. They differed by a factor of four when strains were different and closer to a 

factor of two when predicted using parameters from the same strain. Because these predictions 

do not perfectly match the measured serum concentrations, there remains uncertainty about the 

exposure estimates, and this uncertainty has not been fully characterized.

Table 4-3. Predicted Final Serum Concentration and Time-Integrated Serum 

Concentration (AUC) for Studies in Rats

Study

Species/ 

Strain

Exposure 

Duration

Oral 

Doses

mg/kg/day

Measured 

Final Serum 

value

μg/ml

Species/ 

Strain Used 

for 

Prediction

Predicted 

Final Serum 

Value

μg/ml

Predicted AUC 

mg/L*h

Perkins et 

al. 2004

Rat (M)

ChR-CD

13 weeks

Diet

0.06

0.64

1.94

6.50

7.1 (1.2)

41 (13)

70 (16)

138 (34)

Rat (M)

SpragueDawley

3.8 (0.0955)

34.8 (0.865)

79.5 (3.84)

139 (13.1)

7230 (181)

69100 (1540)

168000 (6520)

326000 (27100)

Butenhoff 

et al. 2004a 

Rat (M)

SpragueDawley

F0M: 10 

wk pre 

matingmating

Gavage

1

3 

10

30

NT

NT

51.5 s

45.3

Rat (M)

SpragueDawley

49.9 (1.53)

102 (6.5)

153 (17.3)

169 (27.7)

92500 (2600)

204000 (10900)

345000 (34200)

412000 (70500)

Notes: Numbers in parentheses indicate SD

M = male; s = serum; NT = not tested

Since the Kemper (2003) data were not tied to toxicological endpoints and were only used in model development, they are not 

included in this table.

Perfluorooctanoic acid (PFOA) – May 2016 4-9 
Table 4-4. Predicted Final Serum Concentration and Time-Integrated Serum 

Concentration (AUC) for Studies in Mice

Study

Species/ 

Strain

Exposure 

Duration

days

Oral 

Doses

mg/kg/d

Measured 

Serum 

Value 

μg/ml

Species/ 

Strain Used 

for 

Prediction

Predicted 

Final Serum 

Value μg/ml

Predicted 

AUC mg/L*h

White et al. 

2009; Wolf 

et al. 2007

Mouse (F)

CD-1 

GDs 1–17a

Gavage

3

5

NT

NT

Mouse (F)

CD-1

25 (2.22)

25.6 (2.26)

33,700 (1,860) 

40,700 (2,170)

GDs 7–17 

Gavage

5 24.8 Mouse (F)

CD-1

29 (2.55) 25,400 (1,320)

DeWitt et 

al. 2008

Mouse (F)

C57BL/6N 

15 days

Drinking 

water

0.94

1.88

3.75

7.5

15

30

NTb

NTb

35.3

42.8

50.0

162.6

Mouse (F)

C57BL/6N

29.7 (1.58)

51.9 (1.89)

70.2 (2.57)

81.4 (3.91)

94.7 (11.8)

117 (29.3)

7,300 (541)

13,800 (951)

22,400 (1,290)

30,500 (1,540)

40,100 (4,720)

56,000 (12,300)

Lau et al. 

2006

Mouse (F)

CD-1 

GDs 1–17

Gavage

1

3 

5 

10

20

40

21.9c

40.5 c

71.9 c

116 c

181 c

271 c

Mouse (F)

CD-1 

57.6 (3.82)

87.2 (7.93)

95.2 (7.41)

106 (5.84)

121 (11)

148 (30.2)

16,400 (606)

33,600 (1,930)

40,700 (2,180)

49,600 (1,980)

61,400 (5,050)

80,100 (12,700)

Notes: Numbers in parentheses indicate SD

NA = not applicable; could not be determined

F = female; GD = gestation day; NT = not tested a Sacrificed on PND 22.

b DeWitt et al. (2008) had 0.94 and 1.88 mg/kg/day dose groups in a second experiment. c The Lau et al. (2006) serum data were provided by the author for animals treated GDs 1–17. 

Table 4-5. Predicted Final Serum Concentration and Time-Integrated Serum (AUC) 

in Studies of Monkeys

Study

Species/

Strain

Exposure 

Duration

Oral Doses

mg/kg/day

Measured Serum

value μg/ml

Species/ 

Strain Used

For 

Prediction

Predicted 

Final 

Serum 

Value 

μg/ml

Predicted 

Exposure (AUC) 

mg/L*h

Butenhoff et 

al. 2002, 

2004b

Monkey 

(M)

Cynomolgus

26 weeks

Oral capsule

3 (n = 3)

10 (n = 4)

30/20 (n = 3)

117.9 (87.6-141)

77.35 (55.4- 96.5)

283.2 (61.7-489)

Monkey

Cynomolgus

89.1 (12.4)

121 (14)

149 (31)

380,000 (50,100)

553,000 (62,800)

710,000 (144,000)

Notes: Numbers in parentheses indicate SD

M = male

The average serum concentration for the LOAEL or NOAEL was determined through 

numeric simulation. The average or mean value has the advantage of normalizing the serum 

concentration across the exposure durations to generate a uniform metric for internal dose in 

situations where the dosing durations varied and serum measurements were taken immediately 

prior to sacrifice. The averaged serum concentration is a hybrid of the AUC and the maximum 

serum concentration. Compared across studies, PFOA average serum concentration appears to be 

a stable reflection of internal dosimetry. 

Table 4-6 provides the AUC from the model, the dosing duration from each of the modeled 

studies, and the resultant average serum concentration. The data from the monkey study 

(Butenhoff et al. 2002, 2004b) were not used because of the small number of animals evaluated 

Perfluorooctanoic acid (PFOA) – May 2016 4-10
and the wide variability in the responses. The internal doses associated with the effect levels 

(LOAELs) differ by less than an order of magnitude (13.1–96.2 mg/L), while the AUC values 

differ by over two orders of magnitude (5,360–380,000 mg/L*h). Given the differences in 

external doses, the projected serum levels are proportionally quite similar.

Table 4-6. Average Serum Concentrations Derived from the AUC and the Duration of 

Dosing

Study

Dosing 

duration

days

NOAEL

mg/kg/day 

(AUC

mg/L*h)

NOAEL

(Av serum

mg/L)

LOAEL 

mg/kg/day

(AUC

mg/L*h)

LOAEL

(Av serum

mg/L)

DeWitt et al. 2008: mice; ↓ IgM 

response to SRBC

15 1.88

(13,800)

38.2 (2.63) 3.75

(22,400)

61.9 (3.58)

Lau et al. 2006: mice

reduced pup ossification (m, f), 

accelerated male puberty

17 None None 1

(16,400)

38.0 (1.4)

Perkins et al. 2004: rats; ↑liver 

weight/necrosis

91 0.64

(69,100)

31.6 (0.073) 1.94

(168,000)

77.4 (2.98)

Wolf et al. 2007: mice; GDs 1–17

↓Pup body weighta 

17 None None 3

(33,700)

77.9 (4.3)

Wolf et al. 2007: mice; GDs 7–17

↓Pup body weighta

11 None None 5

(25,400)

87.9 (4.57)

Butenhoff et al. 2004a:

↓relative body weight/↑ relative 

kidney weight and ↑kidney:brain 

weight ratio in F0 and F1 at sacrifice

84 None None 1

(92,500)

45.9 (1.29)

Notes: Significance p < 0.05 or < 0.01

m = male; f = female; SRBC = Sheep Red Blood Cell

a serum from pups on PND 22

Table 4-6 identifies serum values of 38, 45.9, and 61.9 mg/L as the lowest concentrations 

associated with adverse effects in the Lau et al. (2006), Butenhoff et al. (2004a), and DeWitt et 

al. (2008) studies, respectively. Thus, it appears that the LOAELs are roughly consistent across 

gender, species, and treatment with respect to average serum concentration. Assuming that MoA, 

susceptibility to toxicity, or both do not vary and that PKs alone explains variation, it is 

reasonable to expect similar concentrations to cause similar effects in humans.

The Andersen et al. (2006) model, used to make the predictions in Tables 4-3, 4-4, and 4-5 

can be solved analytically to predict the steady-state concentration (Css) resulting from a fixed

infusion DR (in units of μmol/h): 

CCssss = DDDD

ffffffff ∗ QQffffff �1 +

TTmmmmmm

QQffffff ∗ kkTT + DR�

Although the assumption of a constant infusion exposure simplifies the actual dose regimen 

used, this assumption permits rapid calculation and analysis of Css. Using this equation, one can 

calculate a range of Css values for each DR. The range of Css values are derived from the speciesspecific combinations of parameters from the Bayesian analysis of the available PK data. This 

result for Css depends nonlinearly on DR. The PFOA toxicity studies used discrete, daily doses; 

these doses were converted into rates by dividing daily dose by 24 hours. In Table 4-7, the Css is 

compared with the average serum concentration predicted. The fraction of Css is calculated, 

indicating that the studies range 36%–91% of Css. 

Perfluorooctanoic acid (PFOA) – May 2016 4-11
For human exposure to PFOA, one needs to rely on steady-state calculations since there is a 

lack of both the sufficient PK and exposure knowledge to make more complicated estimates. The 

average serum concentrations of the LOAEL in Table 4-7 are all within roughly one order of 

magnitude (12.4–87.9 mg/L).

Table 4-7. Comparison of Average Serum Concentration and Steady-State Concentration

Study

Dosing 

duration 

days

LOAEL 

mg/kg/day

Css (mg/L) for 

constant infusion 

of LOAEL

Average Serum 

Conc. for Study

(mg/L)

Fraction of Css

(Average / Css)

Perkins et al. 2004 91 1.94 84.4 (3.81) 77.4 (2.98) 0.913 (0.00746)

Butenhoff et al. 2004a 84 1 52.5 (1.72) 45.9 (1.29) 0.874 (0.00776)

Wolf et al. 2007; GDs 7–17 11 5 107 (6.8) 87.9 (4.57) 0.82 (0.0117)

Wolf et al. 2007; GDs 1–17 17 3 95.9 (6.73) 77.9 (4.3) 0.813 (0.0148)

DeWitt et al. 2008 15 3.75 84.1 (4.5) 61.9 (3.58) 0.736 (0.0233)

Lau et al. 2006 17 1 67.8 (4.39) 38 (1.4) 0.561 (0.0277)

Notes: Average serum concentrations from PK simulations of toxicity study treatment regimens and Css were both predicted 

using species-specific parameter distributions (i.e., draws from the Markov Chain determined by analyzing the available PK data 

for the appropriate species). The number in parentheses is the SD.

The predicted average serum concentrations can be converted into an oral equivalent dose by 

recognizing that, at steady state, CL from the body should equal dose to the body. CL can be 

calculated if the rate of elimination (derived from the half-life) and the Vd are both known. In 

making the calculation based on CL, it is important also to consider whether the exposure to 

PFOA has lasted long enough to reach steady state. Four of the endpoints modeled represent 

serum values that are greater than 80% of steady state, but none represent steady-state 

concentrations. Those endpoints representing lower percentages of steady state require 

consideration of the uncertainty resulting from use of a projection that is not representative of a 

steady-state concentration (UFS) when establishing an RfD for a chronic lifetime exposure 

endpoint. 

Measures of half-life in humans have been determined for both workers and the general 

population (section 2.6.2). Olsen et al. (2007) gives the human half-life as 3.8 years for PFOA in 

an occupationally-exposed U.S. cohort. Bartell et al. (2010) determined a value of 2.3 years 

based on the decline in serum levels among members of the general population exposed via 

drinking water in the area near the DuPont Works plant in Washington, West Virginia after the 

drinking water concentrations decreased. EPA chose to use the Bartell et al. (2010) half-life 

value because it is the one most relevant to scenarios where exposures result from ingestion of 

contaminated drinking water by members of the general population. 

Thompson et al. (2010) gives a Vd of 0.17 L/kgbw (section 2.6.3). The Vd is defined as the total 

amount of PFOA in the body divided by the blood or serum concentration. The Vd was calibrated 

using human serum concentrations and exposure data from NHANES and assumes that most 

PFOA intake came from contaminated drinking water. The value for Vd was calibrated so that the 

model prediction of elevated blood levels of PFOA was consistent with the values from NHANES. 

The Vd value determined by Thompson et al. (2010) did not consider PFOA contributions from 

sources other than drinking water. However this estimate is not radically different from the 0.198 

L/kgbw determined for the monkey in the study by Butenhoff et al. (2004b).

The half-life (t1⁄2) and Vd are utilized to calculate the CL for PFOA according to the following 

equation assuming first order kinetics for CL (Medinsky and Klaassen 1996): 

Perfluorooctanoic acid (PFOA) – May 2016 4-12
CL = Vd x (ln 2 ÷ t1⁄2) = 0.17 L/kgbw x (0.693 ÷ 839.5 days) = 0.00014 L/kg bw/day

Where:

Vd = 0.17 L/kg 

ln 2 = 0.693 

t1⁄2 = 839.5 days (2.3 years x 365 days/year = 839.5 days) 

These values combined give a CL of 1.4 x 10-4 L/kg bw/day. 

Scaling the derived average serum concentrations (in mg/L) for the NOAELs and LOAELs 

in Table 4-6 yields the predicted oral HED in mg/kg/day for each corresponding serum 

measurement. The HED values are the predicted human oral exposures necessary to achieve 

serum concentrations equivalent to the LOAEL (and NOAEL where available) in the animal 

toxicity studies. Note that this scaling assumes linear first-order human kinetics in contrast to the 

nonlinear phenomena observed at high doses in animals. It is justifiable at the lower dose 

NOAEL and LOAEL concentrations from the animal studies that that demonstrate the firstorder, linear response to dose necessary to calculate CL.

Thus, HED = average serum concentration (in mg/L) x CL 

Where:

Average serum is from model output in Table 4-8 

CL = 0.00014 L/kg bw/day 

Table 4-8. HEDs Derived from the Modeled Animal Average Serum Values

Study

Dosing 

duration 

days

NOAEL 

mg/kg/d

NOAEL 

Av serum 

mg/L

HED 

mg/kg/d

LOAEL 

mg/kg/d

LOAEL

(Av serum)

mg/L

HED 

mg/kg/d

DeWitt et al. 2008: mice; ↓ 

IgM response to SRBC

15 1.88 38.2 0.0053 3.75 61.9 0.0087

Lau et al. 2006: mice 

reduced pup ossification 

(m,f), accelerated male 

puberty

17 None - - 1 38.0 0.0053

Perkins et al. 2004: rats; 

↑liver weight/necrosis

91 0.64 31.6 0.0044 1.94 77.4 0.0108

Wolf et al. 2007: mice; GDs 

1–17

↓pup body weight

17 None - - 3 77.9 0.0109

Wolf, et al. 2007: mice; 

GDs 7–17

↓pup body weight1

11 None - - 5 87.9 0.0123

Butenhoff et al. 2004a: ↓F0 

body weight/↑ absolute and 

relative kidney weight

84 None - - 1 45.9 0.0064

Macon et al. (2011)

GDs 1–17 ↓mammary gland 

development2

17 - - - 0.3 12.4 0.0017

Notes: Significance p < 0.05 or < 0.01

m = male; f = female; SRBC = Sheep Red Blood Cell

1 serum from pups on PND 22

2 serum from pups on PND 7

Perfluorooctanoic acid (PFOA) – May 2016 4-13
4.1.1.2 RfD Quantification

The subset of studies amenable for use in derivation of HED based on average serum 

measurements from the PK model is based solely on results from studies that have dose and 

species-specific serum values for model input, as well as exposure durations of sufficient length 

to achieve values near to steady-state projections or applicable to developmental endpoints with 

lifetime consequences following short term exposures.

As explained previously, human data identified significant relationships between serum 

levels and specific indicators of adverse health effects but lacked the exposure information for 

dose-response modeling. For this reason none of the human studies provided an appropriate POD 

for RfD derivation. The pharmacokinetically-modeled average serum values from the animal 

studies are restricted to the animal species selected for their low dose response to oral PFOA 

intakes. Extrapolation to humans adds a layer of uncertainty that needs to be accommodated in 

deriving the RfD.

HED PODs. The HEDs derived from Perkins et al. (2004), DeWitt et al. (2008), Lau et al. 

(2006), and Butenhoff et al. (2004a) were each examined as the potential basis for the RfD. Only 

Perkins et al. (2004) and DeWitt et al. (2008) identified a NOAEL from which the HED could be 

derived. The Lau et al. (2006) and Wolf et al. (2007)/White et al. (2009) LOAEL HEDs included 

developmental effects in the offspring as accompanied by the increased liver weight that is an 

accepted biomarker for PFOA exposure. These are developmental exposure studies that carry 

lifetime consequences for a less-than-lifetime exposure. The Butenhoff et al. study (2004a) 

included significant decreased body weight (not confounded by reduced food intake) in F0 males 

accompanied by increased kidney weight (consistent with the need for renal tubular transport) as 

co-critical at the LOAEL. The DeWitt et al. study (2008) has a LOAEL for decreased IgM, 

increased IgG, and increased absolute and relative spleen weight after a 15-day exposure. 

Table 4-9 provides the calculations for potential RfDs using the HEDs derived from PK 

modeling based on the serum values collected at animal sacrifice. The table applies UFs to each 

POD and illustrates the array of potential RfD outcomes. Each POD is impacted by the doses 

utilized in the subject study, the endpoints monitored, and the animal species/gender studied. 

Thus, the array of outcomes, combined with knowledge of the individual study characteristics, 

helps to inform selection of an RfD that will be protective for humans.

The potential lifetime RfD values in Table 4-9 differ by about an order of magnitude 

(0.00002–0.00015 mg/kg/day) but so do the UFs applied to the POD. These results demonstrate 

the ability of the model to normalize the animal data across species, strain, gender, and exposure 

duration. The UFs applied in the derivation of the potential RfDs alter the first-order, direct 

relationship between the animal serum measurements associated with the animal studies and the 

resultant RfD. Accordingly, the resultant RfD cannot be extrapolated to a corresponding human 

serum value equivalent to the RfD using the CL value applied when calculating the HED from 

the animal serum.

Perfluorooctanoic acid (PFOA) – May 2016 4-14
Table 4-9. The Impact of Quantification Approach on the RfD Outcomes for the HEDs 

from the PK Model Average Serum Values 

POD

Value

mg/kg/day UFH UFA UFL UFS UFD UFtotal

Candidate 

RfD

mg/kg/day

PK-HEDNOAEL Perkins 

rats; ↑liver weight/necrosis

0.0044 10 3 - - - 30 0.00015

PK-HEDLOAEL Wolf GD 1–17 

mice; ↓pup body weight

0.0109 10 3 10 - - 300 0.00004

PK-HEDLOAEL Wolf GD 7–17 

mice; ↓pup body weighta

0.0123 10 3 10 - - 300 0.00004

PK-HEDNOAEL DeWitt 

mice; ↓ IgM response to 

SRBC

0.0053 10 3 - 10 - 300 0.00002

PK-HEDLOAEL Lau 

mice reduced pup 

ossification (m,f), 

accelerated male puberty

0.0053 10 3 10 - - 300 0.00002

PK-HEDLOAEL Butenhoff 

↓F0 body weight/↑ absolute 

and relative kidney weight 

0.0064 10 3 10 - - 300 0.00002

Notes: m = male; f = female; SRBC = Sheep Red Blood Cell

a serum from pups on PND 22

The Perkins et al. (2004) and Butenhoff et al. (2004a) studies were conducted in male 

Sprague-Dawley rats with durations of 91 days via diet and 84 days via gavage, respectively. 

Both were associated with increased relative liver weight accompanied by some hepatic necrosis 

in a small number of animals. The Butenhoff et al. study (2004a) also observed a significant 

decrease in body weight compared to controls for the F0 male rats at the end of the 84-day 

exposure. The studies by Lau et al. (2006) and Wolf et al. (2007)/White et al. (2009) were 

conducted in pregnant female mice with a 17-day average exposure via gavage, resulting in 

increased liver weights in the dams and low body weights and developmental delays in offspring. 

Uncertainty Value Application

A UF for intraspecies variability (UFH) of 10 is assigned to account for variability in the 

responses within the human populations because of both intrinsic (toxicokinetic genetic, life 

stage, health status) and extrinsic (life style) factors that can influence the response to dose. No 

information to support a UFH other than 10 was available to characterize interindividual and agerelated variability in the toxicokinetics or toxicodynamics among humans other than variability 

in serum levels measured among populations residing in common geographical locations with 

presumably fairly similar exposures.

A UF for interspecies variability (UFA) of three was applied to account for uncertainty in 

extrapolating from laboratory animals to humans (i.e., interspecies variability). The 3-fold factor 

is applied to account for toxicodynamic differences between the animals and humans. The HEDs 

were derived using average serum values from a model to account for PK differences between 

animals and humans.

A UF for LOAEL to NOAEL extrapolation (UFL) of 10 was applied to all PODs other than 

the Perkins et al. study (2004) to account for use of a LOAEL for the POD. The POD for the 

Perkins et al. study (2004) is a NOAEL. 

Perfluorooctanoic acid (PFOA) – May 2016 4-15
A UF for extrapolation from a subchronic to a chronic exposure duration (UFS) is one, 

because the PODs are based on average serum concentrations and determined to represent > 80% 

of steady state for each study (81%–91%) except for the Lau et al. (2006) developmental study 

(56%). The Lau et al. (2006) developmental HED was not adjusted for lifetime exposures, 

because the average serum values associated with the developmental studies are more protective 

than those for the longer-term studies of systemic toxicity. A UFS of 10 was applied to the 

DeWitt et al. study (2008) serum-derived HED reflecting 73% of steady state because the data 

suggest that longer-term exposures to the same dose have the potential to increase serum values 

beyond the levels indicated by the 15-day exposure to mice. In addition, the NOAEL for 

immunological effects (0.94 mg/kg/day) was a LOAEL for effects on liver weight in the absence 

of histological evaluation on both days 16 and 31 following a 15-day exposure. Thus, there is a 

potential that lifetime exposures at serum steady state could impact the liver, increasing the risk 

for tissue damage.

A database UF (UFD) of one was applied to account for deficiencies in the animal study 

database for PFOA. There are extensive human data from epidemiology studies in the general 

population, as well as in worker cohorts. The epidemiology data provide strong support for the 

identification of hazards observed following exposure to PFOA in the laboratory animal studies 

and the human relevance of the critical effects. However, uncertainties in the use of the available 

epidemiology data precluded their use at this time in the quantification of an RfD. There are 

extensive data from short term, subchronic, chronic, reproductive, developmental, and 

mechanistic studies in laboratory animals that support the endpoints used in calculating the 

potential RfD. The potential RfD is the one applicable to those most at risk from exposure to 

PFOA present in drinking water, the fetus and young infants. The alternative identical RfDs are 

values that could be more appropriate for other exposure scenarios and endpoint concerns. 

4.1.2 RfD Selection

The candidate RfDs in Table 4-9 range 0.00002–0.00015 mg/kg/day. The RfD of 

0.00002 mg/kg/day calculated from HED average serum values from Lau et al. (2006) was 

selected. The RfD based on Lau et al. (2006) is derived from reduced ossification of the proximal 

phalanges (forelimb and hindlimb) and accelerated puberty in male pups (4 days earlier than 

controls) as the critical effects. The selected RfD from the Lau et al. study (2006) is supported by 

the RfD for effects on the response of the immune system (DeWitt et al. 2008) to external 

challenges as observed following the short-term 15-day exposures to mature mice and effects on 

organ and body weights in F1 adult males observed following chronic exposure.

Decreased pup body weights were also observed in studies conducted by Wolf et al. 

(2007)/White et al. (2009) and Lu et al. (2015) using mice receiving external doses within the 

same order of magnitude (1, 3, and 5 mg/kg/day respectively) as those in the chosen study for the 

RfD. The selected RfD from the reproductive and developmental study is supported by the 

longer-term RfD for effects on the response of the immune system (DeWitt et al. 2008) to 

external challenges as observed following the short-term exposures to mature mice and the 

effects on kidney weight observed at the time of sacrifice in the F0 parental males in the 

Butenhoff et al. study (2004a) that provided the modeled serum data. 

Support for the selected RfD is provided by other key studies with NOAELs and LOAELs 

similar to those used for quantification, yet lacked serum data that could be used for modeling. 

There were effects on liver weight and hepatic hypertrophy in the Perkins et al. (2004) and 

Perfluorooctanoic acid (PFOA) – May 2016 4-16
DeWitt et al. (2008) studies that were not considered in the identification of the study LOAEL 

because of a lack of data to demonstrate adversity as determined by the Hall criteria (Hall 2012). 

The LOAEL for evidence of hepatic necrosis and other signs of tissue damage in the F1 male 

rats from the Butenhoff et al. study (2004a) was 3 mg/kg/day, and the NOAEL was 1 mg/kg/day. 

In the Loveless et al. study (2008), for male rats 1 mg/kg/day was a NOAEL for increased 

relative liver weight and focal liver necrosis was seen at a LOAEL of 10 mg/kg/day, while in 

male mice the NOAEL was 0.3 mg/kg/day for the increased liver weight and focal liver necrosis 

was at a LOAEL of 1 mg/kg/day following a 29-day exposure. In the study by Tan et al. (2013) 

the degree of damage to the liver at 5 mg/kg/day became more severe with increased necrosis, 

inflammation, and steatosis when animals were given an HFD. The HED modeled from the 

average serum value in mice for the LOAEL of 3 mg/kg/day from Wolf et al. (2007)/White et al. 

(2009) was 0.0109 mg/kg/day, about twice that of the 0.0053 mg/kg/day for the rats in the Lau et 

al. study (2006) at a LOAEL of 1 mg/kg/day. Both studies lacked a NOAEL. 

Using the PK model of Wambaugh et al. (2013), average serum PFOA concentrations were 

derived from area under the curve (AUC) considering the number of days of exposure before 

sacrifice. The predicted serum concentrations were converted as described above to oral HEDs in 

mg/kg/day for each corresponding serum measurement. The POD for the derivation of the RfD 

for PFOA is the HED of 0.0053 mg/kg/day that corresponds to a LOAEL that represents 

approximately 60% of steady-state concentration. An UF of 300 (10 UFH, 3 UFA, and 10 UFL) 

was applied to the HED LOAEL to derive an RfD of 0.00002 mg/kg/day. 

There are extensive human data from epidemiology studies on the general population, as well 

as worker cohorts. The epidemiology data provide support for the human relevance of the 

hazards identified in the laboratory animals. However, they lack the quantitative information on 

the human exposures (doses and durations) responsible for the human serum levels. Although 

some associations show a relationship between effects and serum measures, the serum measures 

are lower than the PODs from the animal studies and some associations are confounded by 

reverse causality. Data supporting a first-order kinetic relationship between dose/duration and 

serum concentrations are needed before the human data can be used in a manner comparable to 

the process utilized in the RfD derivation.

4.1.3 RfC Determination

Limited data from human epidemiology and animal toxicity studies were available with 

which to evaluate the potential health effects resulting from continuous inhalation exposure to 

PFOA. The available data base, summarized below for human and animal data, does not provide 

the minimum data needed for derivation of the RfC as discussed in USEPA (1994b). Thus, the 

RfC for PFOA is not recommended or derived. 

Human Data. Studies have examined occupational and residential populations at or near largescale PFOA production plants in the United States in an attempt to determine the relationship 

between serum PFOA concentration and various health outcomes suggested by the standard 

animal toxicological database. While inhalation is an important route of exposure to workers, 

drinking water was identified as a contributor to exposure in the general population. In all of the 

epidemiology studies, wide ranges of serum levels were reported that are likely a reflection, not 

only of intra-human toxicokinetic variability, but also of diversity in external exposure sources 

and routes of exposure. Thus, the data cannot be clearly applied to quantification of doseresponse via inhalation. 

Perfluorooctanoic acid (PFOA) – May 2016 4-17
Animal Data. Inhalation toxicity data in laboratory animals were limited to acute, single, and 

repeated exposures for PK studies, and a developmental toxicity study in rats. No subchronic or 

chronic inhalation toxicity studies in animals were available for assessment. Generally, adverse 

effects observed following inhalation exposure to PFOA were similar to effects following 

exposure to an irritating dust. For male rats exposed to PFOA as a dust in air, the 4-hour LC50

was 980 mg/m3, with adverse clinical signs of body weight loss, irregular breathing, red 

discharge around the nose and eyes, and corneal opacity and corrosion (Kennedy et al. 1986, 

2004). 

Distinct toxicokinetic differences between male and female rats were found following single 

and repeated inhalation exposures. Sprague-Dawley rats were exposed nose-only to PFOA 

aerosols of 0, 1, 10, or 25 mg/m3 for 6 hours or for 6 hours/day, 5 days/week, for 3 weeks 

(Hinderliter 2003). Absorption was indicated in both males and females after single and repeated 

exposures with plasma PFOA concentrations proportional to exposure concentration. The Cmax

values were approximately 2–3 times higher in males than in females and persisted for up to 

6 hours in males compared to just 1 hour in females. Similarly, the elimination of PFOA was 

rapid by females at all exposure levels, and by 12 hours after exposure the plasma levels had 

dropped below the analytical LOQ (0.1 μg/ml). In males, the plasma concentration remained 

approximately 90% of the peak concentration at all exposure levels at 24 hours after exposure, 

and steady state was reached following repeated exposures. While these results clearly show 

toxicokinetic differences between male and female rats, toxicity data were not included, limiting 

use of the information in a quantitative risk assessment.

In a developmental toxicity study, pregnant Sprague-Dawley rats were exposed whole-body 

to PFOA dust concentrations of 0, 0.1, 1, 10, or 25 mg/m3

 for 6 hours/day on GDs 6–15 (Staples 

et al. 1984). Dams were either sacrificed on GD 21 or allowed to litter and rear their offspring 

until PND 35. Maternal toxicity at 10 and 25 mg/m3

 consisted of wet abdomens, 

chromodacryorrhea, chromorhinorrhea, a general unkempt appearance, lethargy (highconcentration group only), and decreased body weight and food consumption. Five out of 

24 dams died during treatment at 25 mg/m3

. Significantly increased mean liver weight (p<0.05) 

was seen at 25 mg/m3. No effects were observed on the maintenance of pregnancy or fetal and 

pup survival. At 25 mg/m3, mean offspring body weight was lower than that of controls on GD 

21 and throughout lactation.

4.2 Dose-Response for Cancer Effects

As discussed in section 3.4.5, there is equivocal evidence that PFOA exposure might be 

associated with an increased risk for cancer from the human epidemiology database and animal 

studies. Only one study in highly exposed worker populations showed a positive association 

between death from cancer and PFOA exposure. In that study, a significant increase in mortality 

due to kidney cancer was found for workers in the highest quartile of cumulative PFOA 

exposure; the estimated average mean serum level was 0.35 μg/mL (Steenland and Woskie 

2012). Mortality from cancer in PFOA workers was not elevated in several other studies 

(Leonard et al. 2008; Lundin et al. 2009; Raleigh et al. 2014). Serum levels were not available in 

studies reporting only mortality. No association was found between PFOA level and cancer 

incidence rate in workers with mean serum of 0.113 μg/mL (Steenland et al. 2015). 

No associations were found in the general population between mean serum PFOA levels up 

to 0.0866 μg/mL and colorectal, breast, prostate, bladder, and liver cancer (Bonefeld-Jørgensen

Perfluorooctanoic acid (PFOA) – May 2016 4-18
et al. 2014; Eriksen et al. 2009; Hardell et al. 2014; Innes et al. 2014). In contrast, two studies 

involving members of the C8 Health Project showed a positive association between PFOA levels 

(mean at enrolment 0.024 μg/mL) and kidney and testicular cancers (Barry et al. 2013; Vieira et 

al. 2013).

The only chronic bioassays of PFOA were conducted in rats (Biegel et al. 2001; Butenhoff et 

al. 2012). The two studies support a positive finding for the ability of PFOA to be weakly 

tumorigenic in one or more organs of male but not female rats. There are no carcinogenicity data 

from a second animal species. The study by Butenhoff et al. (2012) examined male and female 

rats; the Biegel et al. study only evaluated males. The tumor types observed were: 

• Liver (Butenhoff et al. 2012). 

• Leydig Cell (Biegel et al. 2001; Butenhoff et al. 2012).

• Pancreatic Ascinar Cell (Biegel et al. 2001).

The dose response information and tumors incidence data from the Butenhoff et al. (2012) 

and Biegel et al. (2001) studies are summarized in Table 4-10. The data are limited in that only 

Butenhoff et al. tested more than one dose. Only one tumor-type (Leydig cell adenoma) 

demonstrated a dose-response relationship. 

Table 4-10. Summary of Tumor Data from Animal Studies

Tissue

Concentration in Diet (ppm)

0 Tumor Type Reference a 30 300

Liver Male 7/50 2/50 10/50 Hepatocellular carcinoma Butenhoff et al. 2012

Liver Male 0/80 NT 0/76 Hepatocellular carcinoma Biegel et al. 2001

Liver Male 2/80 NT 10/76 Hepatocellular adenoma Biegel et al. 2001

Liver Female 0/50 0/50 2/50 Hepatocellular carcinoma Butenhoff et al. 2012

Testes Male 0/50 2/50 7/50 Leydig Cell adenomas Butenhoff et al. 2012

Testes Male 0/80 NT 8/76 Leydig Cell adenomas Biegel et al. 2001

Pancreas Male 1/80 NT 0/76 Acinar Cell carcinoma Biegel et al. 2001

Pancreas Male 0/80 NT 7/76 Acinar Cell adenoma Biegel et al. 2001

Notes: a The value reported is for the ad libitum control

NT = Not tested

There are some data that provide support for the hypothesis that the PPARα agonism is the 

MOA for the observed liver tumors in rats. PPARα is found in human livers and, when activated, 

is linked through activation to a number of metabolic responses but not to the large-scale 

peroxisome proliferation associated with tumors in rats and other rodent species. The data 

support a PPARα MOA for the liver tumors and, thus, are indicative of lack of relevance to 

humans. 

PPARα activation might also play a role in the other tumor types observed. However for the 

Leydig tumors the PPARα involvement is indirect. The favored hypothesis for the DNA 

replication errors responsible for induction of Leydig tumors are postulated to be a consequence 

of the following sequence of events: 

• Decreased testosterone synthesis.

• Increased GnRH and increased levels of LH leading to chronic stimulation of Leydig 

cells by growth-stimulating mediators including IGF-1, TGF-α, leukotrienes and various 

free radicals (Clegg et al. 1997; Li et al. 2011). 

Perfluorooctanoic acid (PFOA) – May 2016 4-19 
There are some experimental data that demonstrate systemic effects of PFOA leading to 

decreased testosterone and increased estradiol as a result of increased activity of aromatase, the

cellular enzyme responsible for the metabolic conversion of testosterone to estradiol (Biegel et 

al. 1995). However, more data to support the relationship of PFOA to intermediate steps in the 

proposed MOAs are needed. 

Current MOA theories for the PACT tumors are linked to the impact of either the mitogenic 

effects of elevated testosterone levels or intestinal tissue hormones (CCK, gastrin, or both) in 

promoting proliferation of acinar cell preneoplastic foci (Klaunig et al. 2003; Obourn et al. 

1997). PACT tumors are most commonly found in rats but also occur in humans. Because PFOA 

is associated with decreased rather than increased levels of testosterone, the mechanistic link 

between PFOA exposure and PACT is more likely associated with gastric hormone changes, 

possibly associated with alterations in bile composition. Some of the membrane transporters that 

are impacted by PFOA function in transport of bile components from the liver to the gallbladder 

and thereby to the intestines. Cholecystokinen and gastrin stimulate contraction of the 

gallbladder and release of bile into the intestines. Data to support this hypothesis are not 

available for PFOA. Obourn et al. (1997) studied the impact of PFOA on CCK using in vitro

assays and found that it was not an agonist for the CCKA receptor that activates CCK release.

The increase in hepatocellular tumors did not show a direct relationship to dose in male rats 

and was not significantly elevated in either males or females at the high dose when compared to 

controls.

There was a dose-related significant increase in LCTs in male rats in the Butenhoff et al. 

study (2012), which was confirmed by the high dose in the single-dose mechanistic study by 

Biegel et al. (2001). At the high dose (300 ppm in the diet; 14.2 mg/kg/day), tumors were found 

in 14% of the male rats at the end of 2 years in the Butenhoff et al. study (2012) and 4% at the 

low dose (1.3 mg/kg/day). In the Biegel et al. study (2001), 11% were affected at a dose of 300 

ppm in the diet (13.6 mg/kg/day). In each case, there were no LCTs in the controls. The PACT 

tumors, only detected in the single dose Biegel et al. study (2001), do not support quantification. 

Under the EPA 2005 cancer guidelines, the evidence for the carcinogenicity of PFOA is 

considered suggestive because only one species has been evaluated for lifetime exposures and 

the tumor responses occurred primarily in males. Dose-response data are only available for the 

LCTs in one study. However, two studies involving members of the C8 Health Project showed a 

positive association between PFOA levels (mean at enrolment of 0.024 μg/mL) and kidney and 

testicular cancers (Barry et al. 2013; Vieira et al. 2013). Therefore, the data on LCTs from 

Butenhoff et al. (2012) were modeled to provide a perspective on the magnitude of the potential 

cancer risk as it compares with the level of protection provided by the RfD. 

The dose-response for the LCTs from Butenhoff et al. (2012) was modeled using EPA’s 

Benchmark Dose Software (BMDS) Version 2.3.1. The multistage cancer model predicted the 

dose at which a 4% increase in tumor incidence would occur. The 4% was chosen as the low-end 

of the observed response range within the Butenhoff et al. (2012) results. Both the first and 

second degree polynomials gave identical goodness-of-fit criteria (p value and Akaike’s 

Information Criterion [AIC]). Results are shown in Table 4-11 and Figure 4-1 and details of the 

model run are given in Appendix A. 

Perfluorooctanoic acid (PFOA) – May 2016 4-20
Table 4-11. Multistage Cancer Model Dose Prediction Results for a 4% Increase

in LCT Incidence

BMD (mg/kg/day) BMDL (mg/kg/day)

First Degree Polynomial Fit 3.51 1.99

Second Degree Polynomial Fit 3.51 1.99

AIC = 62.6936 P = 0.2245

Source: Butenhoff et al. (2012)

Figure 4-1. BMD Model Results for LCTs (Butenhoff et al. 2012) 

The CSF for PFOS is derived from the BMDL04 of 1.99 mg/kg/day after converting the 

animal BMDL to a HED using body weights to the 3⁄4 power. The HED is calculated as follows2: 

HED = Animal BMDL x (animal body weight)1/4 ÷ (human body weight)1/4

HED = 1.99 mg/kg/day x [(0.523 kg)1/4 ÷ (70 kg)1/4] = 1.99 mg/kg/day x 0.29 = 

0.58 mg/kg/day 

Where:

1.99 mg/kg/day = BMDL04 for LCTs

0.29 = The dosimetric adjustment factor

The CSF is calculated from the BMDL04 HED as follows 

CSF = response ÷ BMDL04 HED

CSF = 0.04 ÷ 0.58 mg/kg/day = 0.07 (mg/kg/day)-1

2 Body weight for male Sprague-Dawley rats (chronic Exposures) USEPA 1988

 0

 0.05

 0.1

 0.15

 0.2

 0.25

 0.3

 0 2 4 6 8 10 12 14

Fraction Affected

dose

Multistage Cancer Model with 0.95 Confidence Level

11:59 05/09 2013

BMDL BMD

 

Multistage Cancer

Linear extrapolation

Perfluorooctanoic acid (PFOA) – May 2016 4-21

 
The CSF should not be used at doses > 0.58 mg/kg/day, the HED corresponding to the POD 

for the 4% incidence of LCTs following lifetime exposure to PFOA. The observed dose-response 

relationships do not continue linearly above this level, and the fitted dose-response models better 

characterize the dose-response for the higher exposures. The calculated concentration in drinking 

water with one-in-a-million risk for an increase in testicular tumors at levels greater than 

background is 0.0005 mg/L. 

10-6 Cancer Risk Concentration = 0.000001 x 80 kg ÷ [0.07 (mg/kg/day)−1 x 2.5 L/day] = 

0.000457 mg/L (rounded to 0.0005 mg/L).

The equivalent concentration derived from the RfD for noncancer effects using default adult 

body weight of 80 kg and a default DWI rate of 2.5 L/day (USEPA 2011) and a 20% relative 

source contribution (RSC) to account for the contribution of drinking water to the total exposure 

is 0.0001 mg/L. The 0.0001 mg/L concentration derived for an adult based on the RfD for 

noncancer effects with a 20% RSC for drinking water is lower than the concentration of 

0.0005 mg/L associated with a one-in-a-million risk for testicular cancer also derived with adult 

default values indicating that a guideline derived from the developmental endpoint will be 

protective for the cancer endpoint. 

Perfluorooctanoic acid (PFOA) – May 2016 4-22
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Appendix A: Literature Search Strategy Developing the Search

The literature search strategy was planned with input from EPA library services staff. Chemical 

Abstracts Service (CAS) numbers served as the basis for identification of relevant search terms. 

Trial searches were conducted, and results were evaluated to refine the search strategy (e.g., to 

pevent retrieval of citations unrelated to health and occurrence). The search string was refined to 

improve the relevancy of the results. All searches were conducted in the PubMed database, 

which contains peer-reviewed journal abstracts and articles on various biological, medical, 

public health, and chemical topics. The first search string (as well as future iterations) is 

presented below. 

Every 2 weeks, a search was run in PubMed and a bibliography of the search results was 

compiled. 

In 2012, the State of New Jersey Department of Environmental Protection (NJDEP) initiated a 

monthly search in PubMed for emerging literature on perfluorinated chemicals, primarily from 

the carboxylic acid and sulfonate families. These searches were provided to the EPA on a 

monthly basis. There was a high degree of overlap with the results from the EPA search, thus 

increasing the confidence in the search strategy.

In 2013, the EPA search strategy was expanded to cover other members of the 

perfluorocarboxylic acids (C-4 to C-12) and sulfonate families (C-4, C-6, C-8). The search string 

was altered in June of 2013 to rely more on the search features offered by PubMed. 

A change in the PubMed database structure in 2015 required additional modifications to the 

search strategy.

The NJDEP search terms did not change from 2012 to 2015. All search iterations are noted 

below. However, the reports shared with EPA were streamlined to remove information on 

analytical methods and other nonhealth related citations that were not consistent with the criteria 

presented previously in the backgorund section of this document. 

Search Strategy Examples: (Arranged from most recent to oldest).

2015 

Search: perfluorooctanoate OR “perfluorooctanoic acid” OR “perfluoroctanoic acid” OR pfoa 

OR “perfluorinated chemicals” OR “perfluorinated compounds” OR “perfluorinated homologue 

groups” OR “perfluorinated contaminants” OR “perfluorinated surfactants” OR perfluoroalkyl 

acids OR “perfluorinated alkylated substances” OR “perfluoroalkylated substances” OR pfba OR 

“perfluorobutanoic acid” OR perfluorochemicals OR “telomer alcohol” OR “telomer alcohols” 

OR “fluorotelomer alcohols” OR “polyfluoroalkyl compounds” OR “perfluorooctane sulfonate” 

OR pfos OR “perfluorooctanesulfonic acid” OR “perfluorooctane sulfonic acid” OR 

“perfluorooctane sulphonate” OR perfluorooctane sulfonate OR “perfluorooctanyl sulfonate” OR 

“Heptadecafluorooctane-1-sulphonic” OR “Heptadecafluoro-1-octanesulfonic acid” OR 

perfluorononanoate OR pfhxa OR “perfluorohexanoic acid” OR “fluorinated surfactants” 

Perfluorooctanoic acid (PFOA) – May 2016 A-1 
Filters: English. 

Frequency: Every 2 weeks 

September 2013 

Search: perfluorooctanoate OR “perfluorooctanoic acid” OR “perfluoroctanoic acid” OR pfoa 

OR “perfluorinated chemicals” OR “perfluorinated compounds” OR “perfluorinated homologue 

groups” OR “perfluorinated contaminants” OR “perfluorinated surfactants” OR perfluoroalkyl 

acids OR “perfluorinated alkylated substances” OR “perfluoroalkylated substances” OR pfba OR 

“perfluorobutanoic acid” OR perfluorochemicals OR “telomer alcohol” OR “telomer alcohols” 

OR “fluorotelomer alcohols” OR “polyfluoroalkyl compounds” OR “perfluorooctane sulfonate” 

OR pfos OR “perfluorooctanesulfonic acid” OR “perfluorooctane sulfonic acid” OR 

“perfluorooctane sulphonate” OR perfluorooctane sulfonate OR “perfluorooctanyl sulfonate” OR 

“Heptadecafluorooctane-1-sulphonic acid” OR “Heptadecafluoro-1-octanesulfonic acid” OR 

perfluorononanoate OR pfhxa OR “perfluorohexanoic acid” OR “fluorinated surfactants” 

Filters: English. 

Frequency: Every 2 weeks 

June 2013 

Search: (PFOA[tw] OR perfluorooctanoic acid[tw] OR 335-67-1[tw] OR PFBA[tw] OR 

perfluorobutanoate[tw] OR 3794-64-7[tw] OR PFDA[tw] OR perflurordecanoic acid[tw] OR 

335-76-2[tw] OR PFHpA[tw] OR perfluoroheptanoic acid[tw] OR 375-85-9[tw] OR PFHxA[tw] 

OR perfluorohexanoic acid[tw] OR 307-24-4[tw] OR PFNA[tw] OR perfluorononanoic acid[tw] 

OR 375-95-1[tw] OR PFPtA[tw] OR perfluoropentanoic acid[tw] OR 2706-90-3[tw] OR 

PFPA[tw] OR pentafluoropropionic acid[tw] OR 422-64-0[tw]) AND (human* [tw] OR 

mammal*[tw]) NOT (environment* OR ecolog*) 

Filters: English. 

Frequency: Every 2 weeks 

February, 2013 

Search: perfluorooctanoate OR “perfluorooctanoic acid” OR “perfluoroctanoic acid” OR pfoa 

OR “perfluorinated chemicals” OR “perfluorinated compounds” OR “perfluorinated homologue 

groups” OR “perfluorinated contaminants” OR “perfluorinated surfactants” OR perfluoroalkyl 

acids OR “perfluorinated alkylated substances” OR “perfluoroalkylated substances” OR pfba OR 

“perfluorobutanoic acid” OR perfluorochemicals OR “telomer alcohol” OR “telomer alcohols” 

OR “fluorotelomer alcohols” OR “polyfluoroalkyl compounds” OR “perfluorooctane sulfonate” 

OR pfos OR “perfluorooctanesulfonic acid” OR “perfluorooctane sulfonic acid” OR 

“perfluorooctane sulphonate” OR perfluorooctane sulfonate OR “perfluorooctanyl sulfonate” OR 

“Heptadecafluorooctane-1-sulphonic acid” OR “Heptadecafluoro-1-octanesulfonic acid” OR 

perfluorononanoate OR pfhxa OR “perfluorohexanoic acid” OR “fluorinated surfactants” 

Perfluorooctanoic acid (PFOA) – May 2016 A-2 
Filters: English. 

Frequency: Every 2 weeks 

June 2011 

Search (perfluorooctanoate OR “perfluorooctanoic acid” OR “perfluoroctanoic acid” OR pfoa 

OR “perfluorinated chemicals” OR “perfluorinated compounds” OR “perfluorinated homologue 

groups” OR “perfluorinated contaminants” OR “perfluorinated surfactants” OR 

perfluoroalkylacids OR “perfluorinated alkylated substances” OR “perfluoroalkylated 

substances” OR pfba OR “perfluorobutanoic acid” OR perfluorochemicals OR “telomer alcohol” 

OR “telomer alcohols” OR “fluorotelomer alcohols” OR “polyfluoroalkyl compounds” OR 

“perfluorooctane sulfonate” OR pfos OR “perfluorooctanesulfonic acid” OR “perfluorooctane 

sulfonic acid” OR “perfluorooctane sulphonate” OR perfluorooctanesulfonate OR 

“perfluorooctanyl sulfonate” OR “Heptadecafluorooctane-1-sulphonic acid” OR 

“Heptadecafluoro-1-octanesulfonic acid” OR perfluorononanoate OR pfhxa OR 

“perfluorohexanoic acid” OR “fluorinated surfactants” OR 335-67-1 [rn])

Limits: Publication Date—Dates will change for each search, English Language only

June 2009 

Search (perfluorooctanoate OR “perfluorooctanoic acid” OR “perfluoroctanoic acid” OR pfoa 

OR “perfluorinated chemicals” OR “perfluorinated compounds” OR “perfluorinated homologue 

groups” OR “perfluorinated contaminants” OR “perfluorinated surfactants” OR 

perfluoroalkylacids OR “perfluorinated alkylated substances” OR “perfluoroalkylated 

substances” OR pfba OR “perfluorobutanoic acid” OR perfluorochemicals OR “telomer alcohol” 

OR “telomer alcohols” OR “fluorotelomer alcohols” OR “polyfluoroalkyl compounds” OR 

“perfluorooctane sulfonate” OR pfos OR “perfluorooctanesulfonic acid” OR “perfluorooctane 

sulfonic acid” OR “perfluorooctane sulphonate” OR perfluorooctanesulfonate OR 

perfluorononanoate OR pfhxa OR “perfluorohexanoic acid” OR “fluorinated surfactants” OR 

335-67-1 [rn] OR 1763-23-1 [rn])

Limits: Entrez Date from 2009/04/07 to 2009/04/12 

New Jersey Search Terms

Search: perfluorinated OR perfluorooctanoate OR perfluorononanoate OR 

perfluorooctanesulfonate OR perfluorooctanesulphonate OR perfluoroalkylated OR 

perfluoroalkyl OR polyfluoroalkyl OR polyfluorinated OR PFBA OR PFBS OR PFDA OR 

PFHA OR PFHPA OR PFHXA OR PFHXS OR PFNA OR PFOA OR PFOAs OR PFOS OR 

PFUNDA OR “perfluorooctanoic acid” OR “perfluoro octanoic acid” OR “perfluorooctane 

sulfonate” OR “perfluorooctane sulfonic acid” OR “perfluorooctanesulfonic acid” OR 

“perfluorooctane sulphonate” OR “perfluorooctanyl sulfonate” OR “perfluorobutanoic acid” OR 

“perfluoroalkyl acids” OR “perfluorononanoic acid” OR “perfluorohexanoic acid” OR 

“perfluorohexane sulfonate” OR “perfluorohexane sulphonate” OR perfluorobutanoate OR 

“perfluoro butanoate” OR perfluorohexanoate OR “perfluoro hexanoate” 

Filters: 1

Perfluorooctanoic acid (PFOA) – May 2016 A-3 
Appendix B: Studies Evaluated Since August 2014 

The tables that follow identify the papers that were retrieved and reviewed for inclusion 

following the August 2014 peer review for the draft PFOS Health Effects Support Document. 

The papers listed include those recommended by the peer reviewers or public commenters, as 

well as those identified from the literature searches between the completion of the peer review 

draft and December 2015. The review of papers recommended by the commenters and their 

potential impact on the updates to the draft assessments was facilitated by publications such as 

the critical review of the recent literature by Post et al. (2012). Post et al. (2012) provides an indepth analysis of the available health effects literature for PFOA. Papers included in the final 

HESD are noted and reasons provided for those that were not included in the final document. 

The tables for document retrieval and review are followed by updated versions of the 

summaries of the epidemiology summary tables from the peer reviewed draft as recommended 

by the peer reviewers. They are a useful tool to facilitate a high level comparison of the study 

outcomes for each of the epidemiological study groupings. 

The criteria utilized in determining the papers that were included in the HESD after the peer 

review and presented in the Background were the following: 

1. The study examines a toxicity endpoint or population that had not been examined by 

studies already present in the draft assessment.

2. Aspects of the study design, such as the size of the population exposed or quantification 

approach, make it superior to key studies already included in the draft document. 

3. The data contribute substantially to the weight of evidence for any of the toxicity 

endpoints covered by the draft document. 

4. There are elements of the study design that merit its inclusion in the draft assessment 

based on its contribution to the mode of action or the quantification approach. 

5. The study elucidates the mode of action for any toxicity endpoint or toxicokinetic 

property associated with PFOA exposure. 

6. The effects observed differ from those in other studies with comparable protocols.

Table B-1. PFOS Epi Papers—Post Peer Review (Retrieved and Reviewed)

Authors and Year Topic—Keywords Status/Notes

Andersen et al. 2013 Postnatal growth Added PFOA/PFOS

Back et al. 2015 Time to pregnancy Added PFOA

Barrett et al. 2015 Ovarian hormone Not Added—No association observed for 

PFOA; PFOS was not included in the 

assessment

Berg et al. 2015 Thyroid Added PFOA/PFOS

Bonefeld-Jørgenson et al. 2014 Breast cancer Added PFOA/PFOS

Bonefeld-Jørgenson et al. 2011 Breast cancer Added PFOA/PFOS

Brieger et al. 2011 Immune effects Already presented in PFOS/PFOA

Buck Louis et al. 2015 Semen quality Added PFOA/PFOS

Chang et al. 2014 Analysis of human cancer 

studies

Added PFOA in the cancer weight of 

evidence section

Chen et al. 2015 Birth weight Added PFOS

Dankers et al. 2013 Blood-testis barrier Reviewed,—not added; Study of an assay 

that used PFOA as one chemical in the test 

battery

Perfluorooctane sulfonate (PFOS) – May 2016 B-1 
Authors and Year Topic—Keywords Status/Notes

Darrow et al. 2013 Reproductive outcome Added PFOA/PFOS

Darrow et al. 2014 Miscarriage Added PFOA/PFOS

Donauer et al. 2015 Infant Neurobehavior Not added—negative for PFOS; No 

statistical differences in PFOA levels 

during pregnancy and any neuro endpoint. 

Better studies.

Eriksen et al. 2013 Total cholesterol—Danish Added PFOA/PFOS

Fitz-Simon et al. 2013 Serum lipids Added PFOA/PFOS

Fisher et al. 2013 Plasma lipids Added PFOA/PFOS

Fletcher et al. 2013 Cholesterol–genes Added PFOA/PFOS

Fu et al. 2014 Serum lipids in Chinese 

subjects

Not added: Chinese population, dataset 

available on U.S. population. More 

branched chain isomers found among the 

people in China.

Geiger et al. 2014a Lipids/children Added PFOA/PFOS

Geiger et al. 2014b Hypertension/children Added PFOA/PFOS

Ghisari et al. 2014 Breast cancer—Inuit Not added; same population as BonefeldJørgensen et al. 2014; this study focuses on 

gene polymorphisms

Governini et al. 2015 DNA effects in sperm Added PFOA/PFOS

Grandjean and Clapp 2015 Health Risks Not added; the primary studies are already 

included in the documents.

Granum et al. 2013 Immune children Added PFOA/PFOS

Hardell et al. 2014 Prostate cancer Added PFOA/PFOS

Høyer et al. 2015a Human weight Added PFOA/PFOS

Høyer et al. 2015b Behavior motor development Added PFOA/PFOS

Humblet et al. 2014 Asthma Added PFOA/PFOS

Jain 2014 NHANES Added PFOA/PFOS

Innes et al. 2014 Colorectal cancer Added PFOA/PFOS

Joensen et al. 2013 Sperm Added PFOA/PFOS

Kerger et al. 2011 Cholesterol C8 Added; demographics for cholesterol and 

PFOS in summary section of epi studies

Kjeldsen and Bonefeld-Jørgensen 

2013

Sex hormones Covered multiple PFAS in vitro no impact 

on weight of evidence

Kristensen et al. 2013 Prenatal female repro Added PFOA/PFOS

Liew et al. 2014 Cerebral palsy children Added PFOA/PFOS

Looker et al. 2014 Immune Added PFOA/PFOS

López-Doval et al. 2014 Male repro Added PFOS

Maisonet et al. 2015 Gestational diabetes Added PFOA/PFOS

Maisonet et al. 2012 Birth weight Added PFOA/PFOS

Mørck et al. 2015 PFAS levels in children Not added; No significant impact

Okada et al. 2014 Allergy children Added PFOS

Osuna et al. 2014 Antibodies PFOS PFOA Not added; focus more on methylHg and 

PCB than PFAS; only n = 38 as 

preliminary study

Roth and Wilks 2014 Neurodevelopmental . Not added; no significant impact

Shrestha et al. 2015 Thyroid Added PFOA/PFOS

Perfluorooctane sulfonate (PFOS) – May 2016 B-2 
Authors and Year Topic—Keywords Status/Notes

Starling et al. 2014 Plasma lipids Added PFOA/PFOS

Steenland et al. 2015 Workers Added PFOA

Stein et al. 2009 Pregnancy Added PFOA

Taylor et al. 2014 Menopause Added PFOA/PFOS

Vanden Heuvel 2013 Serum lipids Not added; is a rebuttal of Fletcher et al. 

2013 conclusions. No significant impact

Vassiliadou et al. 2010 PFOS in cancer vs non-cancer 

patients

Added PFOA/PFOS

Vélez et al. 2015 Fertility Added PFOA/PFOS

Verner et al. 2015 Fetal growth GFR Added PFOA/PFOS

Verner and Longnecker 2015 Menstruation/excretion Added PFOS

Vested et al. 2013 Semen quality and hormones Added PFOS/PFOS

Vesterinen et al. 2014 Fetal Growth GFR Added PFOA/PFOS

Wang et al. 2013 Thyroid Added PFOA/PFOS

Watkins et al. 2013 Kidney function Added PFOA/PFOS

Webster et al. 2014 Maternal thyroid Added PFOA/PFOS

Webster et al. 2015 Thyroid—iodine statue Added PFOS

Wen et al. 2013 Thyroid Added PFOA/PFOS 

Yeung et al. 2013 Liver cancer Added PFOA/PFOS

Zhang et al. 2015 Gestational diabetes Added PFOA/PFOS

Table B-2. PFOA Post Peer Review Animal Toxicity Studies

Authors and Year Topic Action Notes

Bjork et al. 2011 Nuclear receptor activation In vitro, mechanistic findings comparable to 

studies already included

Corsini et al. 2014 Immune data review Not added; no significant impact

Corsini et al. 2012 Immune in vitro data review Not added; no significant impact

Dewitt et al. 2015 Immunotoxicity Added PFOA

Fenton 2015 Repro editorial Not added 

Filgo et al. 2015 Liver tumors in females 

developmentally exposed

Added PFOA

Hall et al. 2012 PPARα and cancer Cited in synthesis. Paper on adversity of liver 

hypertrophy PFOA/ PFOS

Koustas et al. 2014 Fetal growth (animal studies) 

navigation guide

Added PFOA

Liu et al. 2015 Testes Added PFOA

Long et al. 2013 Neurotoxicity adult PFOS Added PFOS

Lu et al. 2015 Testes Added PFOA

Ngo et al. 2014 Tumors mice Min/+ PFOS Added PFOS 

Post et al. 2012 Review paper Not added. Key studies included in the document; 

no significant impact

Quist et al. 2015 Liver histopathology/high fat 

diet post weaning exposure

Added PFOA

Rigden et al. 2015 Acute liver effects Added PFOA

Shabalina et al. 2015 Brown fat uncoupling protein 1 Not added. Mechanistic; no significant impact

Perfluorooctane sulfonate (PFOS) – May 2016 B-3 
Authors and Year Topic Action Notes

Sheng et al. 2016 Binding to liver fatty acid 

binding protein

Not added; no significant impact, topic covered by 

other papers

Tan et al. 2012 Gene activation Added PFOA/PFOS

Tan et al. 2013 Gene activation dietary fat Added PFOA

Tucker et al. 2015 Mammary gland Added PFOA

Wallace et al. 2013 Mitochodrial respiration Not added. No significant impact, topic covered 

by other papers

Wan et al. 2014b Glucose metabolism Added PFOS

Wan et al. 2012 Hepatic steatosis Added PFOS

Wan et al. 2014a Sertoli cells Added PFOS

F. Wang et al. 2015 MiRNA liver PFOS early life Not added; no significant impact

S. Wang et al. 2014 Lysine decarboxylase Added PFOA/PFOS

L. Wang et al. 2014 Inhibition of LDL Added PFOS

Y. Wang et al. 2015 Special learning and memory Added PFOS

Yan et al. 2015 Glucose homeostasis Not added. Dose-response in Wan (2014b)

presented (more robust). Single dose for whole 

animal

Yu et al. 2015 Thyroid PFOS isomers Added PFOS

Zeng et al. 2014 Mitochondrial mediated 

apoptosis of the heart

Added PFOS

L. Zhang et al. 2013 Fatty acid binding protein Added PFOA/PFOS

Y. Zhang et al. 2013 Biological half-life Added PFOA/PFOS

W. Zhang et al. 2014 Breast cancer cell invasion—

mechanistic 

Not added; in vitro, no significant impact

Zhao et al. 2014 Testosterone reduction in 

Leydig cells PFOS

Added PFOS

Perfluorooctane sulfonate (PFOS) – May 2016 B-4 
Table B-3. Toxicokinetics: Post Peer Review 

Authors and Year Topic Action Notes

D’Alessandro et al. 2013 Serum albumin Added PFOS

Augustine et al. 2005 Transporter expression testes Not added background paper on testes transporters 

–no relationship to PFOA PFOA or any PFAS

Beesoon et al. 2011 Isomer profile Added PFOA

Beesoon and Martin 2015 Albumin binding Added PFOA

Cui et al. 2010 Excretion subchronic Added PFOA/PFOS

Fàbrega et al. 2014 PK model Added PFOA/PFOS

Kerstner-Wood et al. 2003 Plasma protein binding Added—PFOA/PFOS

Klaassen and Aleksunes 

2010

Transporter paper—Provided 

diagram of kidney transporters

Added PFOA

Loccisano et al. 2013 PK model—Human Added PFOA/PFOS

Mondal et al. 2014 Breast milk Added PFOS/PFOA 

Ospinal-Jimenez and Pozzo 

2012

Protein denaturation Added PFOS

Pérez et al. 2013 Human tissue levels New PFOA/PFOS

Ren et al. 2015 Thyroid hormone receptor 

binding (in vitro)

Added PFOA/PFOS

Rigden et al. 2015 Liver and excretion Added PFOA

Shabalina et al. 2015 Brown fat Not added; No information on MOA for body 

weight effects in the animal or human studies

Slitt et al. 2007 Transporter expression PFOA Not added. Reported on transporters during 

extrahepatic cholestasis. No data on PFOA and 

PFOS. No significant impact.

Tucker et al. 2015 Menstruation-excretory route Added PFOA

Verner and Longnecker 

2015

Excretion PFOS Added PFOS

Wambaugh et al. 2013 PK model Added PFOA/PFOS

Wong et al. 2014 Menstrual blood as excretory 

route

Added PFOA/PFOS

T. Zhang et al. 2014 Excretion general population 

and pregnancy

Added PFOA/PFOS

L. Zhang et al. 2014 PPAR gamma Added PFOS

Y. Zhang et al. 2013 Excretion, half-life Added PFOA/PFOS 

T. Zhang et al. 2013 Maternal transfer Added PFOA/PFOS

Perfluorooctane sulfonate (PFOS) – May 2016 B-5 
Tables B-4 through B-8 provide updated versions of the epidemiology summary tables from the peer-reviewed draft, as 

recommended by the reviewers. They are a useful tool to facilitate a high-level comparison of the study outcomes for each of the 

epidemiology study groupings. 

Table B-4. Association between Serum PFOA and Serum Lipids and Uric Acid

Reference Study Type n

Mean Serum 

PFOA TC VLDL LDL HDL Non-HDL TG UA

Occupational Populations

Olsen et al. 

2000

Cross-sectional 111 (1993)

80 (1995)

74 (1997)

0–80 μg/mL

0–114 μg/mL

0.1–81 μg/mL

↔

↔

↔

NM ↔

↔

↔

↔

↔

↔

NM ↔

↔

↔

NM

Olsen et al. 

2001b, 2003

Cross-sectional 206 (Antwerp)

215 (Decatur)

1.03 μg/mL

1.90 μg/mL

↑ NM NM ↔ NM ↑ NM

Olsen et al. 

2001c, 2003

Longitudinal; ~5 

years

175

(Decatur and 

Antwerp 

combined for 

analysis)

1.36–1.41 

μg/mL (1995 

baseline)

1.49–1.77 

μg/mL (2000 

follow-up)

↑ NM NM ↔ NM ↑ NM

Sakr et al. 

2007a

Cross-sectional 1,025 0.428 μg/mL ↑ ↑ ↑ ↔ NM ↔ ↑

Sakr et al. 

2007b

Longitudinal 454 (23-yr 

follow-up)

1.04 μg/mL 

(first)

1.16 μg/mL 

(last)

↑ NM ↔ ↔ NM ↔ NM

Olsen and 

Zobel 2007

Cross-sectional 506 (Antwerp, 

Cottage Grove, 

Decatur 

combined)

2.21 μg/mL ↔ NM ↔ ↓ NM ↑ NM

Costa et al. 

2009

Cross-sectional 34 workers

107 controls

4.02 μg/mL ↑ NM NM ↔ NM ↔ ↑

General Populations

Emmett et al. 

2006

Cross-sectional 371 0.354 μg/mL ↔ NM NM NM NM NM NM

Steenland et al. 

2009

Cross-sectional 

(C8)

46,294 0.08 μg/mL ↑ NM ↑ ↔ ↑ ↑ NM

Steenland et al. 

2010

Cross-sectional 

(C8)

53,458 0.086 μg/mL NM NM NM NM NM NM ↑

Perfluorooctanoic acid (PFOA) – May 2016 B-6 
Reference Study Type n

Mean Serum 

PFOA TC VLDL LDL HDL Non-HDL TG UA

Winquist and 

Steenland 

2014a

Cross-sectional 

(C8)

32,254 0.0261 μg/mL ↑ NM NM NM NM NM NM

Frisbee et al. 

2010

Cross-sectional 

(C8, children and 

adolescents)

6,536 children

5,934 

adolescents

0.0777 μg/mL

0.0618 μg/mL

↑ NM ↑ ↔ NM ↑ NM

Fitz-Simon et 

al. 2013

Longitudinal; 4.4 

years (C8)

521 0.140 μg/mL 

(baseline)

0.068 μg/mL 

(follow-up)

↔ NM ↔ ↔ NM ↔ NM

Nelson et al. 

2010

Cross-sectional 

(NHANES)

1,445 0.0046 μg/mL ↑ NM ↔ ↔ ↑ NM NM

Eriksen et al. 

2013

Cross-sectional 753 0.0071 μg/mL ↑ NM NM NM NM NM NM

Starling et al. 

2014

Cross-sectional 

(maternal at 14–

26 weeks 

gestation)

891 0.00225 

μg/mL

↔ NM ↔ ↑ NM ↔ NM

Fisher et al. 

2013

Cross-sectional 2,700 0.0025 μg/mL ↔ NM ↔ ↔ NM NM NM

Notes: ↑ = positive association; ↓ = negative association; ↔ = no association; TC = total cholesterol; VLDL= very low density lipoprotein; LDL= low-density lipoprotein; nonHDL= TC(VLDL,IDL, LDL)-HDL; HDL= high-density lipoprotein; TG = triglycerides; UA = uric acid; NM = not measured

Perfluorooctanoic acid (PFOA) – May 2016 B-7 
Table B-5. Association of Serum PFOA and Biochemical and Hematological Measures

Reference Study Type n

Mean Serum 

PFOA

Liver 

enzymes Bilirubin

Renal 

Enzymes/Function Glucose Hematology

Occupational Populations

Olsen et al. 

2000

Cross-sectional 111 (1993)

80 (1995)

74 (1997)

0–80 μg/mL

0–114 μg/mL

0.1–81 μg/mL

↔

↔

↔

↔

↔

↔

↔

↔

↔

↔

↔

↔

↔

↔

↔

Olsen et al. 

2001b, 2003

Cross-sectional 206 

(Antwerp)

215 (Decatur)

1.03 μg/mL

1.90 μg/mL

↔

↔

↔

↔

↔

↔

↔

↔

↔

↔

Olsen et al. 

2001c, 2003

Longitudinal; 

~5 years

175

(Decatur and 

Antwerp 

combined for 

analysis)

1.36–1.41 μg/mL 

(1995 baseline)

1.49–1.77 μg/mL 

(2000 follow-up)

↔

↔

↔

↔

↔

↔

↔

↔

↔

↔

Sakr et al. 

2007a

Cross-sectional 1025 0.428 μg/mL ↑ (GGT only) ↔ NM ↔ ↔

Sakr et al. 

2007b

Longitudinal 454 1.04 μg/mL (first)

1.16 μg/mL (last)

↑ (AST only) ↓ NM NM NM

Olsen and 

Zobel 2007

Cross-sectional 506 

(Antwerp, 

Cottage 

Grove, 

Decatur 

combined)

2.21 μg/mL ↑ (ALP, 

ALT, GGT 

Decatur only)

↓ NM NM NM

Costa et al. 

2009

Cross-sectional 56 workers 4.02 μg/mL ↑ (GGT, 

ALP, ALT)

↓ ↔ ↔ ↔

General Populations

Emmett et al. 

2006

Cross-sectional 371 0.354 μg/mL ↔ NM ↔ NM ↔

Lin et al. 

2010

Cross-sectional 

(NHANES)

1076 men

1140 women

0.00505 μg/mL

0.00406 μg/mL

↑ (ALT, 

GGT)

↔ NM NM NM

Gallo et al. 

2012

Cross-sectional 

(C8)

47,092 0.028 μg/mL ↑ (ALT) ↔ NM NM NM

Shankar et 

al. 2011

Cross-sectional 

(NHANES)

4587 0.0059 μg/mL NM NM ↑ (chronic kidney 

disease)

NM NM

Watkins et 

al. 2013

Cross-sectional 

(C8)

9,660 

(children)

0.0283 μg/mL NM NM ↑ (decreased eGFR) NM NM

Notes: ↑ = positive association; ↓ = negative association; ↔ = no association; ALP = alkaline phosphatase; eGFR = estimated glumerular filtration rate; GGT = gamma-glutamyl 

transpeptidase; AST = aspartate aminotransferase; ALT = alanine transaminase; NM = not measured

Perfluorooctanoic acid (PFOA) – May 2016 B-8 
Table B-6. Association between PFOA level and prevalence of thyroid disease and thyroid hormone levels

Study Study Type Population (n)

Mean Serum 

PFOA (μg/mL) Thyroid Disease TSH T3 T4

Occupational Populations

Olsen et al. 1998 Cross-sectional 111 and 80 Adult 

workers

10–30

>30

NM

NM

↑

↔

NM NM

Olsen et al. 2001b, 2003 Cross-sectional Adult workers

215 (Decatur)

206 (Antwerp)

1.9

1.03

NM ↔ ↔ ↔

Sakr et al. 2007a Cross-sectional 1,025 Adult workers 0.428 NM ↔ ↔ ↔

Costa et al. 2009 Cross-sectional 56 Adult workers 4.02 NM ↔ ↔ ↔

Olsen and Zobel 2007 Cross-sectional 506 Adult workers 2.21 NM ↔ ↑ ↔ serum

↓ free

General Populations

Emmett et al. 2006 Cross-sectional 40 (thyroid disease)

331 (no thyroid 

disease)

0.387

0.451

↔ NM NM NM

Pirali et al. 2009 Cross-sectional 28 Adults 2.0 ng/g thyroid 

tissue

↔ NM NM NM

Bloom et al. 2010 Cross-sectional 31 Adults 0.00133 NM ↔ NM ↔

Shrestha et al. 2015 Cross-sectional 51 men

36 women

0.0104 ↔ ↔ ↑ ↑

Winquist and Steenland 

2014b

Cross-sectional 32,254 (C8) 0.0261 ↔ (men)

↑ (women)

NM NM NM

Lopez-Espinosa et al. 2012 Cross-sectional 10,725 children (C8) 0.0293 ↑ ↔ NM ↔

Melzer et al. 2010 Cross-sectional 3,966 Adults 

(NHANES)

0.025 (men)

0.019 (women)

↔ (men)

↑ (women)

NM NM NM

Wen et al. 2013 Cross-sectional 1,181 (NHANES) 0.00415 NM ↔ ↔ ↔

de Cock et al. 2014 Cross-sectional 83 newborns 0.000943 (cord) NM NM NM ↔ boys

↑ girls

Lin et al. 2013 Cross-sectional 545 0.00267 NM ↔ NM ↔

Chan et al. 2011 Cross-sectional 271 Pregnant 

women

0.00135 NM ↔ NM ↔

Wang et al. 2013 Cross-sectional 903 women at 

gestation week 18

0.0022 NM ↔ NM NM

Perfluorooctanoic acid (PFOA) – May 2016 B-9 
Study Study Type Population (n)

Mean Serum 

PFOA (μg/mL) Thyroid Disease TSH T3 T4

Berg et al. 2015 Cross-sectional 375 women at 

gestation week 18, 

day 3 and week 6 

after delivery 

(Norwegian 

Mother/Child 

Cohort)

0.00153 NM ↔ ↔ ↔

Webster et al. 2014 Cross-sectional 152 women at 

gestation week 15–

18

0.0017 NM ↔ ↔ ↔

Notes: ↑ = positive association; ↓ = negative association; ↔ = no association; NM = Not Measured

Perfluorooctanoic acid (PFOA) – May 2016 B-10
Table B-7. Association between Serum PFOA and Markers of Immunotoxicity

Study Study Type Population (n)

Mean or Median 

Serum PFOA 

(μg/mL)

Disease 

Prevalence

Vaccine 

Response 

Steenland et al. 

2015

Cross-sectional Workers (6,027) 0.113 ↑ ulcerative 

colitis

↑ rheumatoid 

arthritis

↓ asthma

NM

Okada et al. 2012 Prospective 

cohort

Maternal, third 

trimester (343)

0.0014 ↔ up to 18 

months old

NM

Fei et al. 2010b Cross-sectional Maternal, first 

trimester 

(1,400)

0.0056 ↔ early 

childhood

NM

Grandjean et al. 

2012

Prospective 

cohort

Maternal at 

gestation week 

32 (587)

0.0032 NM ↓ (antibody titer)

Grandjean et al. 

2012

Prospective 

cohort

Children age 5 

years (587)

0.00406 NM ↓ (antibody titer)

Granum et al. 

2013

Prospective 

cohort

Women at 

delivery (56)

0.0011 ↔ ↔

Dong et al. 2013 Cross-sectional Children age 

10–15 years 

(231 asthmatics 

and 225 

controls)

0.0015 (asthmatics)

0.0010 

(nonasthmatics)

↑ for asthma NM

Humblet et al. 

2014

Cross-sectional Children age 

12–19 years 

(1,877)

0.0043 (asthmatics)

0.0040 

(nonasthmatics)

↑ for asthma NM

Looker et al. 2014 Cross-sectional Adults (411) 0.0337 NM ↓ (antibody titer)

Notes: ↑ = positive association; ↓ = negative association; ↔ = no association; NM = Not Measured

Perfluorooctanoic acid (PFOA) – May 2016 B-11
Table B-8. Association between Serum PFOA and Reproductive and Developmental Outcomes

Study Study Type n Mean Serum PFOA Outcome

Measures at 

Birth

Growth/ 

Development

Fecundity/ 

Fertility

Reproductive Outcome, Anthropometric Measures at Birth

Fei et al. 2007,

2008a, 2009, 

2010a

Cross-sectional 1,400 0.0056 μg/mL ↔ (gestation length)

↓ (length of 

breastfeeding)

↓ (weight)

↓ (size)

↔ (Apgar score)

NM ↑ (TTP)

↑ (infertility)

Vélez et al. 2015 Cross-sectional 1,743 0.00166 μg/mL NM NM NM ↑ (TTP)

↑ (infertility)

Nolan et al. 2009, 

2010

Cross-sectional 1,555 0.00678 μg/mL ↔ (preterm birth, 

congenital anomalies, 

labor/delivery 

complications, 

maternal risk)

↔ (weight) NM NM

Stein et al. 2009 Cross-sectional 

(C8)

1,505 0.0488 μg/mL ↔ (miscarriage) ↔ (low weight) NM NM

Darrow et al. 

2013, 2014

Cross-sectional 

(C8)

1,330 and 

1,129

0.031–0.0337 μg/mL ↔ (preterm, 

miscarriage)

↔ (low weight, 

birth weight)

NM ↑ (hypertension)

Apelberg et al. 

2007

Cross-sectional 293 0.0016 μg/mL (cord 

blood)

↔ (gestational age) ↓ (weight, head 

circumference, 

ponderal index)

NM NM

Monroy et al. 

2008

Cross-sectional 101 0.00254 μg/mL 

(maternal at 24-28 

weeks)

0.00224 μg/mL 

(maternal at delivery)

0.0019 μg/mL 

(umbilical cord 

blood)

NM ↔ (weight) NM NM

Washino et al. 

2009

Cross-sectional 428 0.0014 μg/mL NM ↔ (weight and 

size)

NM NM

Hamm et al. 2010 Cross-sectional 252 0.0021 μg/mL ↔ (gestation length) ↔ (weight) NM NM

Whitworth et al. 

2012

Cross-sectional 849 0.0021 μg/mL NM ↔ (birth weight) NM NM

Maisonet et al. 

2012

Cross-sectional 395 0.0037 μg/mL NM ↓ (birth weight) NM NM

Chen et al. 2012 Cross-sectional 429 0.0018 μg/mL NM ↔ (birth weight) NM NM

Perfluorooctanoic acid (PFOA) – May 2016 B-12
Study Study Type n Mean Serum PFOA Outcome

Measures at 

Birth

Growth/ 

Development

Fecundity/ 

Fertility

Male Fertility

Joensen et al. 

2009

(PFOA/PFOS 

combined)

Cross-sectional 105 0.0049 μg/mL NM NM NM ↓ (normal sperm)

↔ (testosterone)

Joensen et al. 

2013

Cross-sectional 247 0.0035 μg/mL NM NM NM ↔ (semen 

parameters)

↔ (testosterone, 

hormones)

Buck Louis et al. 

2015 

Cross-sectional 462 0.00429–0.00509 

μg/mL

NM NM NM ↑ (lower % sperm 

with coiled tail) 

(total of six PFAS 

associated with 

changes in sperm 

quality)

Neurodevelopmental Endpoints

Fei et al. 2008b Cross-sectional 1,400 0.0056 μg/mL NM NM ↔ (motor skills 

and mental 

develop. at 6 and 

18 months)

NM

Lieu et al. 2014 Cross-sectional 156 cases

550 controls

0.00456 μg/mL NM NM ↑ (cerebral palsy in 

boys)

NM

Fei and Olsen 

2011

Cross-sectional 787 

(behavior)

537 

(coordination)

0.0057 μg/mL NM NM ↔ (behavior and 

coordination at 7 

years)

NM

Høyer et al. 2015a Cross-sectional 1,106 0.0014 μg/mL NM NM ↔ (motor skills)

↑ (hyperactivity)

NM

Stein et al. 2013 Cross-sectional 

(C8)

321 0.0351 μg/mL (child) NM NM ↔ (behavioural 

problems)

↑ (executive 

function; ADHD 

from mother, not 

teacher)

NM

Hoffman et al. 

2010

Cross-sectional 

(NHANES)

571 children 0.0044 μg/mL NM NM ↑ (ADHD) NM

Perfluorooctanoic acid (PFOA) – May 2016 B-13
Study Study Type n Mean Serum PFOA Outcome

Measures at 

Birth

Growth/ 

Development

Fecundity/ 

Fertility

Postnatal Development

Andersen et al. 

2010

Cross-sectional 1,010 0.0052 μg/mL NM NM ↓ (weight and BMI 

in boys at 5 and 12 

months)

NM

Andersen et al. 

2013

Cross-sectional 811 (children 

at age 7 

years)

0.0053 μg/mL NM NM ↔ (height, weight, 

waist

measurement, risk 

of overweight)

NM

Høyer et al. 2015b Cross-sectional 1,022 0.001–0.0018 μg/mL NM NM ↔ (overweight)

↑ (waist-to-height 

ratio)

NM

Lopez-Espinosa et 

al. 2011

Cross-sectional 

(C8)

3,076 boys

2,931 girls

0.02–0.026 μg/mL NM NM ↑ (delayed puberty 

in girls)

NM

Christensen et al. 

2011

Cross-sectional 448 girls 0.0036–0.0039 

μg/mL (maternal)

NM NM ↔ (age at 

menarche)

NM

Kristensen et al. 

2013

Cross-sectional 343 women 0.0036 μg/mL 

(maternal)

NM NM ↑ (delayed 

puberty)

NM

Vested et al. 2013 Cross-sectional 169 men 0.0038 μg/mL 

(maternal)

NM NM ↑ (lower sperm 

conc and total 

count)

NM

Halldorsson et al. 

2012

Cross-sectional 665 0.0037 μg/mL NM NM ↑ (overweight in 

females at 20 

years)

NM

Notes: ↑ = positive association; ↓ = negative association; ↔ = no association; NM = Not Measured

Perfluorooctanoic acid (PFOA) – May 2016 B-14
Appendix C: Multistage Model for Leydig Cell Tumors

 ==================================================================== 

 Multistage Cancer Model. (Version: 1.9; Date: 05/26/2010) 

 Input Data File: C:/1Data/MyFiles/PFOA-PFOS/PFOA Docs/msc_Leydig_Opt.(d) 

 Gnuplot Plotting File: C:/1Data/MyFiles/PFOA-PFOS/PFOA Docs/msc_Leydig_Opt.plt 

 Thu May 09 11:59:27 2013 

 ==================================================================== 

 BMDS_Model_Run 

~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~ 

 The form of the probability function is: 

 P[response] = background + (1-background)*[1-EXP(-beta1*dose^1-beta2*dose^2)] 

 The parameter betas are restricted to be positive 

 Dependent variable = Col2 

 Independent variable = Col1 

 Total number of observations = 3 

 Total number of records with missing values = 0 

 Total number of parameters in model = 3 

 Total number of specified parameters = 0 

Degree of polynomial = 2 

 Maximum number of iterations = 250 

 Relative Function Convergence has been set to: 1e-008 

 Parameter Convergence has been set to: 1e-008 

Default Initial Parameter Values 

 Background = 0.0132945 

 Beta(1) = 0.0097738 

 Beta(2) = 0 

Asymptotic Correlation Matrix of Parameter Estimates 

 ( *** The model parameter(s) -Beta(2) have been estimated at a boundary point, or 

have been specified by the user, and do not appear in the correlation matrix) 

Background Beta(1)

Background 1 -0.64

Beta(1) -0.64 1

Parameter Estimates 

Variable Estimate

95.0% Wald Confidence Interval

Std. Err. Lower Conf. Limit Upper Conf. Limit

Background 0.00409839 * * *

Beta(1) 0.0116288 * * *

Beta(2) 0 * * *

* - Indicates that this value is not calculated. 

Perfluorooctanoic acid (PFOA) – May 2016 C-1 
Analysis of Deviance Table 

Model Log(likelihood) # Param's Deviance Test d.f. P-value

Full model -28.6454 3

Fitted model -29.3468 2 1.40286 1 0.2362

Reduced model -34.0451 1 10.7995 2 0.004518

 AIC: 62.6936 

Goodness of Fit 

Dose Est._Prob. Expected Observed Size

Scaled

Residual

0.0000 0.0041 0.205 0.000 50 -0.454

1.3000 0.0190 0.952 2.000 50 1.084

14.2000 0.1557 7.784 7.000 50 -0.306

 Chi^2 = 1.48 d.f. = 1 P-value = 0.2245 

 Benchmark Dose Computation 

 Specified effect = 0.04 

 Risk Type = Extra risk 

 Confidence level = 0.95 

BMD = 3.51044 

BMDL = 1.99346 

BMDU = 10.7788 

Taken together, (1.99346, 10.7788) is a 90 % two-sided confidence interval for the BMD 

Multistage Cancer Slope Factor = 0.0200656 

Perfluorooctanoic acid (PFOA) – May 2016 C-2 
 ==================================================================== 

Multistage Cancer Model. (Version: 1.9; Date: 05/26/2010) 

Input Data File: C:/1Data/MyFiles/PFOA-PFOS/PFOA Docs/msc_Leydig_Opt.(d) 

Gnuplot Plotting File: C:/1Data/MyFiles/PFOA-PFOS/PFOA Docs/msc_Leydig_Opt.plt 

 Thu May 09 12:05:42 2013 

 ==================================================================== 

 BMDS_Model_Run 

~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~ 

 The form of the probability function is: 

 P[response] = background + (1-background)*[1-EXP(-beta1*dose^1)] 

 The parameter betas are restricted to be positive 

 Dependent variable = Col2 

 Independent variable = Col1 

 Total number of observations = 3 

 Total number of records with missing values = 0 

 Total number of parameters in model = 2 

 Total number of specified parameters = 0 

Degree of polynomial = 1 

 Maximum number of iterations = 250 

 Relative Function Convergence has been set to: 1e-008 

 Parameter Convergence has been set to: 1e-008 

 0

 0.05

 0.1

 0.15

 0.2

 0.25

 0.3

 0 2 4 6 8 10 12 14

Fraction Affected

dose

Multistage Cancer Model with 0.95 Confidence Level

11:59 05/09 2013

BMDL BMD

 

Multistage Cancer

Linear extrapolation

Perfluorooctanoic acid (PFOA) – May 2016 C-3 
Default Initial Parameter Values 

 Background = 0.0132945 

 Beta(1) = 0.0097738 

Asymptotic Correlation Matrix of Parameter Estimates 

Background Beta(1)

Background 1 -0.64

Beta(1) -0.64 1

Parameter Estimates 

Variable Estimate

95.0% Wald Confidence Interval

Std. Err. Lower Conf. Limit Upper Conf. Limit

Background 0.00409839 * * *

Beta(1) 0.0116288 * * *

* - Indicates that this value is not calculated. 

Analysis of Deviance Table 

Model Log(likelihood) # Param's Deviance Test d.f. P-value

Full model -28.6454 3

Fitted model -29.3468 2 1.40286 1 0.2362

Reduced model -34.0451 1 10.7995 2 0.004518

 AIC: 62.6936 

Goodness of Fit 

Dose Est._Prob. Expected Observed Size

Scaled

Residual

0.0000 0.0041 0.205 0.000 50 -0.454

1.3000 0.0190 0.952 2.000 50 1.084

14.2000 0.1557 7.784 7.000 50 -0.306

 Chi^2 = 1.48 d.f. = 1 P-value = 0.2245 

 Benchmark Dose Computation 

 Specified effect = 0.04 

 Risk Type = Extra risk 

 Confidence level = 0.95 

BMD = 3.51044 

BMDL = 1.99346 

BMDU = 8.7003 

Taken together, (1.99346, 8.7003) is a 90% two-sided confidence interval for the BMD 

Multistage Cancer Slope Factor = 0.0200657

Perfluorooctanoic acid (PFOA) – May 2016 C-4