Perfluorooctanoic Acid – February 2014 i 

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Health Effects Document 

for Perfluorooctanoic 

Acid (PFOA) 
 

 

Health Effects 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: 822R14001 

Date: February 2014 

Perfluorooctanoic Acid – February 2014 

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ii 
Perfluorooctanoic Acid – February 2014 iii

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ACKNOWLEDGMENT

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. 
Perfluorooctanoic Acid – February 2014 iv

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Authors, Contributors, and Reviewers

Chemical manager

Joyce Morrissey Donohue, Ph.D. 

Office of Water, Office of Science and Technology 

Health and Ecological Criteria Division

U.S. Environmental Protection Agency, Washington D.C.

Authors (EPA) 

Joyce Morrissey Donohue, Ph.D. 

Office of Water, Office of Science and Technology 

Health and Ecological Criteria Division

U.S. Environmental Protection Agency, Washington D.C. 

Tina Moore Duke, M.S. 

Office of Water, Office of Science and Technology 

Health and Ecological Criteria Division

U.S. Environmental Protection Agency, Washington D.C. 

Authors (Oak Ridge National Laboratory) 

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 

Peer Reviewers

Internal

Christopher Lau, Ph.D.

National Health and Environmental Effects Research Laboratory, Office of Research and 

Development 

Reproductive Toxicology Division 

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

John Wambaugh, Ph.D. 

National Center for Computational Toxicology, Office of Research and Development 
Perfluorooctanoic Acid – February 2014 v 

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Systems Models for Chemical Toxicity and Exposure 

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

National Center for Environmental Assessment

Office of Research and Development

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

External 
Perfluorooctanoic Acid – February 2014 vi 

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TABLE OF CONTENTS

ACKNOWLEDGMENT..............................................................................................................................iii

Authors, Contributors, and Reviewers.........................................................................................................iv

TABLE OF CONTENTS............................................................................................................................. vi

LIST OF TABLES.....................................................................................................................................viii

LIST OF FIGURES ...................................................................................................................................... x

ABBREVIATIONS AND ACRONYMS .................................................................................................... xi

1.0 EXECUTIVE SUMMARY............................................................................................................. 1-1

2.0 IDENTITY: CHEMICAL AND PHYSICAL PROPERTIES......................................................... 2-1

3.0 TOXICOKINETICS ........................................................................................................................ 3-1

3.1 Absorption ................................................................................................................................ 3-1

3.2 Distribution............................................................................................................................... 3-3

3.3 Metabolism............................................................................................................................. 3-22

3.4 Excretion................................................................................................................................. 3-22

3.4.1 Mechanistic Studies of Renal Excretion................................................................................. 3-26

3.5 Toxicokinetic Considerations................................................................................................. 3-33

3.5.1 PK Models.............................................................................................................................. 3-33

3.5.2 Half-Life Data ........................................................................................................................ 3-41

3.5.3 Volume of Distribution Data .................................................................................................. 3-45

4.0 HAZARD IDENTIFICATION........................................................................................................ 4-1

4.1 Human Effects.......................................................................................................................... 4-1

4.1.1 Long-Term and Epidemiological Studies................................................................................. 4-1

4.1.1.1 Non Cancer Systemic Toxicity Studies ........................................................................... 4-2

4.1.1.2 Thyroid Effects.............................................................................................................. 4-14

4.1.1.3 Steroid Hormones.......................................................................................................... 4-17

4.1.1.4 Reproductive & Developmental Endpoints................................................................... 4-20

4.1.2 Cancer..................................................................................................................................... 4-27

4.2 Animal Studies ....................................................................................................................... 4-30

4.2.1 Acute Toxicity ........................................................................................................................ 4-30

4.2.2 Short-Term Studies................................................................................................................. 4-30

4.2.3 Subchronic Studies................................................................................................................. 4-41

4.2.4 Neurotoxicity.......................................................................................................................... 4-44

4.2.5 Developmental/Reproductive Toxicity................................................................................... 4-46

4.2.6 Chronic Toxicity..................................................................................................................... 4-65

4.2.7 Carcinogenicity....................................................................................................................... 4-68

4.3 Other Key Data....................................................................................................................... 4-71

4.3.1 Mutagenicity and Genotoxicity .............................................................................................. 4-71

4.3.2 Immunotoxicity ...................................................................................................................... 4-73

4.3.3 Hormone Disruption............................................................................................................... 4-83

4.3.4 Physiological or Mechanistic Studies..................................................................................... 4-89

4.3.5 Structure-Activity Relationship............................................................................................ 4-100

4.4 Hazard Characterization ....................................................................................................... 4-101

4.4.1 Synthesis and Evaluation of Major Noncancer Effects........................................................ 4-101

4.4.2 Synthesis and Evaluation of Carcinogenic Effects............................................................... 4-111

4.4.3 Mode of Action and Implications in Cancer Assessment..................................................... 4-113

4.4.4 Weight of Evidence Evaluation for Carcinogenicity............................................................ 4-119

4.4.5 Potentially Sensitive Populations ......................................................................................... 4-119

5.0 DOSE-RESPONSE ASSESSMENT............................................................................................... 5-1

5.1 Dose-Response for Noncancer Effects..................................................................................... 5-1

5.1.1 RfD Determination ................................................................................................................... 5-1
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5.1.1.1 Benchmark Dose Approach ............................................................................................. 5-6

5.1.1.2 Pharmacokinetic Model approach.................................................................................... 5-9

5.1.1.3 RfD Quantification ........................................................................................................ 5-18

5.1.2 RfC Determination ................................................................................................................. 5-23

5.2 Dose-Response for Cancer Effects......................................................................................... 5-24

6.0 REFERENCES................................................................................................................................ 6-1

APPENDIX A............................................................................................................................................... 1

APPENDIX B............................................................................................................................................... 1
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LIST OF TABLES

TABLE 2- 1 Chemical and Physical Properties of PFOA and APFO ....................................................... 2-3

TABLE 3-1. Protein Binding in Rat, Human, and Monkey Plasma .......................................................... 3-3

TABLE 3-2. Dissociation Constants (Kd) of Binding Between PFOA and Albumin............................... 3-4

TABLE 3-3.Tissue Distribution of PFOA in Wistar Rats After 28 Days of Treatment............................ 3-7

TABLE 3-4. Distribution of PFOA in Male Sprague-Dawley Rats After Oral Exposure......................... 3-9

TABLE 3-5. Distribution of PFOA in Female Sprague-Dawley Rats After Oral Exposure ................... 3-10

TABLE 3-6. PFOA Concentration in Wild-type and PPARα-null Mice................................................. 3-12

TABLE 3-7. Plasma PFOA Concentrations (μg/ml) in Postweaning Sprague-Dawley Rats.................. 3-13

TABLE 3-8. Plasma PFOA Concentrations in Male Rats....................................................................... 3-13

TABLE 3-9. Plasma PFOA Concentrations in Female Rats.................................................................... 3-14

TABLE 3-10. Maternal Plasma PFOA Levels (μg/ml) During Gestation and Lactation........................ 3-15

TABLE 3-11. Placenta, Amniotic Fluid, and Embryo/Fetus PFOA Concentrations (μg/ml) ................. 3-16

TABLE 3-12. Fetus/Pup PFOA Concentration (μg/ml) During Gestation and Lactation...................... 3-16

TABLE 3-13. PFOA Levels (μg/ml) in Sprague-Dawley Maternal Milk During Lactation................... 3-16

TABLE 3-14. PFOA Levels (ng/ml) During Gestation and Lactation in Selected Fluids and Tissues... 3-18

TABLE 3-15. Female Offspring PFOA Levels (ng/ml) After GD1-17 Exposure.................................. 3-19

TABLE 3-16. Female Offspring Serum PFOA Levels (ng/ml) After GD10-17 Exposure .................... 3-20

TABLE 3-17. Serum PFOA Levels (ng/ml) Over Three Generations.................................................... 3-20

TABLE 3-18. Urine PFOA Concentrations in Male and Female Rats.................................................... 3-24

TABLE 3-19. Cumulative Percent 14C-PFOA Excreted in Urine and Feces by Rats............................. 3-24

TABLE 3-20. Cumulative Percent 14C-PFOA Excreted in Urine and Feces.......................................... 3-25

TABLE 3-21. Kinetic Parameters of PFC Transport by OAT1, OAT3 and OATp1a1........................... 3-31

TABLE 3-22. Plasma and Urine PFOA Concentration 24 hr After Treatment with 30 mg/kg PFOA... 3-32

TABLE 3-23. Model Parameters for 1 and 10 mg/kg Single Doses of PFOA to CD1 Mice .................. 3-35

TABLE 3-24. Pharmacokinetic Parameters in Male Rats Following Administration of PFOA ............ 3-44

TABLE 3-25. Pharmacokinetic Parameters in Female Rats Following Administration of PFOA.......... 3-45

TABLE 4-1. Association of Serum PFOA with Serum Lipids and Uric Acid in Studies of Occupational 

Populations........................................................................................................................................ 4-7

TABLE 4-2. Association of Serum PFOA with Serum Lipids and Uric Acid in Studies of General 

Populations........................................................................................................................................ 4-9

TABLE 4-3. Associations of Serum PFOA with Serum Clinical Biochemistry and Hematology Measures

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

TABLE 4-4. Association of serum PFOA with the prevalence of thyroid disease and thyroid hormone 

levels in studies of general and worker populations........................................................................ 4-17

TABLE 4-5. Hepatic Effects of Rats Exposed to PFOA ........................................................................ 4-33

TABLE 4-6. Hepatic Effects in PFOA-treated Mice .............................................................................. 4-36

TABLE 4-7. Mouse Hepatocyte Ultrastructure After PFOA or Wythe 14,643 Treatment .................... 4-37

TABLE 4-8. Relative Response of hPPARα, mPPARα, and PPARα-null mice to PFOA..................... 4-38

TABLE 4-9. Liver Effects In Male Rats................................................................................................. 4-44

TABLE 4-10. Organ weight data from F0 males.................................................................................... 4-47

TABLE 4-11. Organ weight data from F1 males.................................................................................... 4-49
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TABLE 4-12. Overview of studies in which pregnant CD-1 mice were administered PFOA ............... 4-56

TABLE 4-13. Liver weight data in monkeys administered PFOA for 6 months.................................... 4-66

TABLE 4-14. Incidence of Ovarian Stromal Hyperplasia and Adenoma in Rats .................................. 4-69

TABLE 4-15. Mammary Gland Tumor Incidence Comparison ............................................................. 4-70

TABLE 4-16. Genotoxicity of PFOA In Vitro ........................................................................................ 4-72

TABLE 4-17. Selected Clinical Chemistry Parameters in Mice Treated with PFOA for 5 Days .......... 4-78

TABLE 4-18. Selected Clinical Chemistry Parameters in Mice Treated with PFOA for 10 Days ........ 4-79

TABLE 4-19. Impact of PFOA on Splenic and Thymic Lymphocyte Populations................................ 4-80

TABLE 4-20. Estimated EC50 Values................................................................................................... 4-86

TABLE 4-21. Data Collection for Female Mice Gestationally-exposed to PFOA................................. 4-87

TABLE 4-22. mRNA Expression of Hepatic PPARα and Related Genes ............................................. 4-92

TABLE 4-23. Activation of Mouse and Human PPAR by PFOA.......................................................... 4-95

TABLE 4-24. Impact of PFOA exposure on PC12 cells........................................................................ 4-97

TABLE 5-1. NOAEL/LOAEL data for Oral Subchronic and Chronic Studies of PFOA ........................ 5-3

TABLE 5-2. Shorter term and Developmental Oral Exposure Studies .................................................... 5-5

TABLE 5-3. Benchmark Dose Modeling for a 10% Increase in Liver Weight........................................ 5-7

TABLE 5-4. Description of prior distributions used to analyze all but the C57BL/6N mice................. 5-11

TABLE 5-5. Estimated and assumed pharmacokinetic parameters used in model development........... 5-12

TABLE 5-6. Predicted Final Serum Concentration and Time-integrated Serum Concentration (AUC) for 

Studies in Rats................................................................................................................................. 5-13

TABLE 5-7. Predicted Final Serum Concentration and Time-integrated Serum Concentration (AUC) for 

Studies in Mice................................................................................................................................ 5-14

TABLE 5-8. Predicted Final Serum Concentration and Time-integrated Serum (AUC) in Studies of 

Monkeys.......................................................................................................................................... 5-15

TABLE 5-9. Average Serum concentrations Derived from the AUC and the duration of Dosing......... 5-16

TABLE 5-10. Human Equivalent Doses Derived from the Modeled Animal Average Serum Values.. 5-18

TABLE 5-11. RfD Point of Departure Options (mg/kg/day) from the PFOA Animal Studies.............. 5-19

TABLE 5-12. The Impact of Quantification Approach on the RfD outcome for the PODs from the 

Palazzolo et al (1993) Study............................................................................................................ 5-20

TABLE 5-13. The Impact of Quantification Approach on the RfD Outcome for PODs from the BMDLs

......................................................................................................................................................... 5-21

TABLE 5-14. The Impact of Quantification Approach on the RfD Outcomes for the HEDs from the 

Pharmacokinetic Model Average Serum Values............................................................................. 5-22

TABLE 5-15. Summary of Tumor Data from Animal Studies............................................................... 5-25

TABLE 5-16. Model Results for Leydig Cell Tumors (Buttenhoff et al. (2012) ................................... 5-27
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LIST OF FIGURES

FIGURE 2-1 Chemical Structures of PFOA and APFO............................................................................ 2-1

FIGURE 2-2 PFOA Anti-Conformer ......................................................................................................... 2-1

FIGURE 2-3 PFOA Lowest Energy Conformer........................................................................................ 2-2

FIGURE 3-1. PFOA binding sites on human serum albumin.................................................................... 3-5

FIGURE 3-2. Renal Organic Acid Transporters...................................................................................... 3-27

FIGURE 3-3. Schematic for a physiologically-motivated renal resorptions pharmacokinetic model..... 3-33

FIGURE 3-4. Physiologically-motivated pharmacokinetic model schematic for PFOA-exposed rats... 3-34

FIGURE 3-5. Schematic for one compartment model............................................................................. 3-35

FIGURE 3-6. Pharmacokinetic model of gestation and lactation in mice............................................... 3-37

FIGURE 3-7. Structure of the PFOA PBPK model in monkeys and humans........................................ 3-38

FIGURE 3-8. Structure of the PBPK Model for PFOA in the Adult Sprague Dawley Rat.................... 3-40

FIGURE 4-1. PPARα-Agonism Mode of Action for Liver Tumors..................................................... 4-114

FIGURE 5-1. BMDS graphic output from selected model runs............................................................... 5-8

FIGURE 5-2. BMD Model Results for Leydig Cell Tumors (Buttenhoff et al., 2012).......................... 5-27
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ABBREVIATIONS AND ACRONYMS

ADAF age-Dependent Agjustment Factor

ADHD attention deficit hyperactivity disorder

ADX adrenalectomized

AIC Akaike's Information Criterion

ALP alkaline phosphatase

ALT alanine transaminase

ANOVA analysis of variance

Areg amphiregulin 

AP alkaline phosphatase

AST aspartate aninotransferase

ATSDR Agency of Toxic Substances and Disease Registry

AUC area under the plasma concentration time curve

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

AWWA American Water Works Association

AwwaRF American Water Works Association Research Foundation 

BMD benchmark dose 

BMDL benchmark dose – lower 95th percentile confidence bound 

BMDS benchmark dose software 

BMI body mass index 

BMR benchmark response 

BQL below quantifiable limit

BSA bovine serum albumin 

BSEP bile salt export pump

BUN blood urea nitrogen 

bw body weight 

C Celsius

Cal-EPA California Environmental Protection Agency 

CaMKII calcium/calmodulin-dependent protein kinase II 

CAR constitutive androstane receptor

CAS Chemical Abstracts Service 

CAT carnitine aceyltransferase

CCK cholecystokinin 

CCL Contaminant Candidate List 

CCL 3 Contaminant Candidate List 3 

CDC Centers for Disease Control and Prevention

CFSE 6-carboxyfluorescein succinimidyl ester 

ChAT choline acetyltransferase

CHO Chinese hamster ovary

CI confidence interval 

Clp plasma clearance

CL clearance

CLR renal clearance

Cmax peak plasma concentration at the first intestinal absorption loci 

CoA coenzyme A
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ConA concanavalinA 

CORT corticosterone

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

CWS community water system 

d day 

DCDQ Developmental Coordination Disorder Questionnaire

DCFH-DA 2’,7’-dichlorofluorescein diacetate

DES diethylstilbestrol

DHT 5α-dihydroxy-testosterone

DMSO dimethyl sulfoxide 

DNA Deoxyribonucleic acid 

DTH delayed-type hypersensitivity 

DWI drinking water intake 

E2 17-β estradiol

E3S estrone-3-sulfate

EC50 half maximal effective concentration

ECF Electro-Chemical Fluorination

EFSA European Food Safety Authority 

EGFR epidermal growth factor receptor

ELISA enzyme-linked immunosorbent assay 

ER endpplasmic reticulum 

ERα estrogen receptor α 

Errα estrogen-related receptor α

FID flame ionization detector

FSH follicle-stimulating hormone 

FXR farnesoid receptor 

g gram 

GAP-43 growth-associated protein-43 

GD gestation day 

GGT gamma-glutamyl transpeptidase 

GJIC gap junction intercellular communication 

GlyT glycogen tropholblast cell 

GnRH gonadotropin releasing hormone 

GSD geometric standard deviation 

hCG human chorionic gonadotropin 

HDL high density lipoprotein 

HED human equivalent dose 

HEK human embryonic kidney 

HET heterozygous 

HGFα hepatocyte growth factor 

HL-60 human promyelocytic leukemia cell line 

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

HRBC horse red blood cells 
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HRL health reference level 

HSA human Serum Albumin 

HSD17β1 hydroxysteroid 17β dehydrogenase 1 

HSD3β1 hydroxysteroid 3β dehydrogenase 1 

HSDB Hazardous Substances Database 

IC50 half-maximal Inhibiting Concentration

ICR imprinting control region

IGF-I insulin like growth factor I

IHD ischemic heart disease

IL-6 interleukin 6

IRR incidence rate ratio

IRIS Integrated Risk Information System 

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 

LC50 lethal concentration for 50% of animals 

LC-MS liquid chromatography – mass spectrometry

LCT Leydig cell tumors 

LD lactation day

LD50 lethal dose for 50% of animals 

LDH lactic dehydrogenase

LDL low density lipoprotein

L-FABP liver fatty acid binding protein

LH luteinizing hormone 

LHWA Little Hocking Water Association

LLOQ lower limit of quantification

LOAEL lowest observed adverse effect level 

LOQ Limit of Quantitation

LPS lipopolysaccharide 

m meter 

MCAD medium chain acyl-CoA dehydrogenase 

MCLG Maximum Contaminant Level Goal 

Mdr2 mulidrug resistance protein 2

μg microgram

mg milligram

min minute 

mL milliliter

μm micrometer

MMAD mass median aerodynamic diameter

MOA mechanism of action
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mol mole 

mPL mouse placental lactogen

mPLP mouse prolactin like protein

MRL minimum reporting level 

MRP multidrug restistance-associated protein

MTBE methyl tertiary-butyl ether

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

NAWQA National Water Quality Assessment 

NDWAC National Drinking Water Advisory Council 

ng nanogram 

NAS National Academy of Sciences

ND not detected, no differences, or no data 

Nd2 NADH dehydrogenase 2 

Ndufs8 NADH dehydrogenase iron-sulfur protein 8 

NHANES The National Health and Nutrition Examination Survey 

NIRS National Inorganics and Radionuclides Survey 

NK natural killer

NM not monitored 

NMR Nuclear Magnetic Resonance

NMRI Naval Medical Research Institute

NOAEL no observed adverse effect level 

NPDWR National Primary Drinking Water Regulation 

NRC National Research Council

Nrf1 nuclear respiratory factor 1

Nrf2 nuclear respiratory factor 2

NTCP sodium-taurocholate contransporting polypeptide

OAT organic anion transporter

OATp organic anion transporting peptide 

OGWDW Office of Ground Water and Drinking Water 

OPP-RED Office of Pesticide Programs – Reregistration Eligibility Decision

OR odds ratio

OVA ovalbumin 

OVX ovariectomized

OW Office of Water

P progesterone 

PACT pancreatic acinar cell tumors

PAH polycyclic aromatic hydrocarbon 

PB phenobarbital 

PBMC peripheral blood mononuclear cells 

PBPK pharmacologically based pharmacokinetic 

PCNA proliferating cell nuclear antigen

PenH enhanced pause airway respiration

PFAA perfluoroalkyl acid

PFC perfluorinated carboxylate acids 

PFDA perfluorodecanoic acid 

PFHxS perfluorohexanesulfonic acid 

PFOA perfluorooctanoic acid 
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PFOS perfluoroocatane sulfonate 

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

PH peroxisomal bifunctional protein 

PHA phytohemagglutinin 

PK pharmacokinetic

PND postnatal day 

POD point of departure 

pKa acid dissociation constant

PPAR peroxisome proliferator activated receptor

Ppb parts per billion

ppm parts per million 

PT peroxisomal thiolase 

PWG Pathology Working Group

PWS public water system

PXR pregnane X receptor 

Q flow in and out of tissues 

Qfilc median fraction of blood flow to the filtrate

Reg Det 2 Regulatory Determinations on the Second CCL 

RfC reference concentration 

RfD reference dose 

RNA ribonucleic acid

ROS reactive oxygen species

RR relative risk

RSA rodent serum albumin 

RSC relative source contribution

RTECS The Registry of Toxic Effects of Chemical Substances

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

SPI Society of the Plastics Industry

SRBC sheep red blood cells 

S-TGC sinusoidal trophoblast giant cells 

Tmax time of maximum plasma concentration

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 
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Tfam transcription factor A

TG triglycerides

TH tyrosine hydroxylase 

TNFα tumor necrosis factor α 

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 trasnporter

U.S. EPA U.S. Environmental Protection Agency 

USGS U.S. Geological Service 

Vd volume of distribution 

Vmax maximum initial rate of an enzyme catalysed 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

WT wild type 
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1.0 EXECUTIVE SUMMARY PFOA

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

produce fluoropolymers. Because of strong carbon-fluorine bonds, PFOA is stable to metabolic 

and environmental degradation and is resistant to biotransformation. Extensive data in humans 

and animals demonstrate ready absorption of PFOA and distribution of the chemical throughout 

the body by noncovalent binding to plasma proteins. The liver is an important binding site with 

increased liver weight in laboratory animals one of the early, low-dose manifestations of 

exposure. At least three transport families seem to play a role in PFOA absorption, distribution, 

and excretion: organic anion transporters (OATs), organic anion transporting peptides (OATps) 

and multidrug resistance-associated proteins (MRPs). PFOA is not readily eliminated from 

humans as evidenced by the half-life of 2.3 years. In contrast, half-life values for the monkey, 

rat, and mouse are 20.8 days, 11.5 days, and 15.6 days, respectively. Differences in transporters 

may explain species differences in elimination.

PFOA is known to activate the peroxisome proliferator activated receptor (PPAR) 

pathway by increasing transcription of mitochondrial and peroxisomal lipid metabolism, sterol, 

and bile acid biosynthesis and retinol metabolism genes. However, based on transcriptional 

activation of many genes in PPARα null mice, the effects of PFOA involve more than activation 

of PPAR. Also activated are the constitutive androstane receptor (CAR), farnesoid receptor 

(FXR), and pregnane X receptor (PXR).

Epidemiology studies have examined occupational and residential populations at or near 

large-scale 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 studies. Exposures were mainly through contaminated drinking 

water and to multiple PFCs. These studies have found a positive association between serum 

PFOA concentration and higher cholesterol in the general population and in worker populations 

but no consistent trends for the low-and high-density protein lipids. A positive association has 

been shown between serum PFOA concentrations and increased liver enzymes and/or decreased 

bilirubin in both worker and general populations, chronic kidney disease in the general 

population, and the odds of experiencing early menopause. Maternal or child plasma levels of 

PFOA were positively associated with decreased antibody titers in children after vaccination, 

obesogenic effects in female children at 20 years of age, and parent reported Attention Deficit 

Hyperactivity Disorders. 

No consistent associations were identified between serum PFOA and hyperglycemia, 

type II diabetes, thyroid homeostasis, fertility, fecundity, and birth outcome. Maternal or child 

plasma PFOA levels were not associated with changes in children reaching motor or mental 

developmental milestones, child risk of hospitalization for infectious diseases, or attainment of 

puberty. 

In most animal studies, short-term and chronic exposure to PFOA resulted in an increase 

in liver weight as at least one of the critical effects. Co-occurring effects in these studies 

included changes in spleen, thymus, liver and/or developmental endpoints. In repeated dosing 

studies, rats given 0.64 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-
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generation study, male rats had increased liver and kidney weights at 1 mg/kg/day. In shorterterm studies, slightly higher doses to rats also resulted in increased liver weight and liver lesions. 

In mice, developmental toxicity and increased spleen weight at a dose of 1 mg/kg/day 

accompanied increased liver weight. Slightly higher doses resulted in decreased 

immunoglobulin levels. 

U.S. EPA has selected 0.00002 mg/kg/day as the RfD for PFOA based on the consistency 

of the response and with recognition of the use of liver weight as a common denominator for loss 

of homeostasis and protection against co-occurring adverse effects. This value is the outcome 

for modeled serum values from three rat studies and one mouse study. In two of the rat studies, 

liver effects were accompanied by developmental effects and kidney weight increases. A 

pharmacokinetic model was used to predict a serum area under the curve (AUC) concentration 

for each LOAEL and NOAEL for liver weight effects. From each AUC, the average serum 

concentration was calculated which was then used to calculate a human equivalent dose (HED). 

The total uncertainty factor (UF) applied to the HEDNOAEL from rat study was 30 which included 

a UF of 10 for intrahuman variability and a UF of 3 to account for toxicodynamic differences 

between animals and humans. For the critical studies that lacked an HEDNOAEL, and additional 

10-fold UF was added to adjust for the use of a HED based on a LOAEL. Four of six candidate 

studies resulted in the same RfD. All of the studies supporting the RfD, except one, are studies 

with exposures to PFOA for ≥ 84 days meeting the duration requirements for determination of a 

lifetime exposure value; the exception is a 17-day developmental toxicity study in the mouse. 

Under the U.S. EPA 2005 cancer guidelines, the evidence for the carcinogenicity of 

PFOA is considered suggestive because only one species has been evaluated and studies for gene 

mutations were negative. Epidemiology studies demonstrate an association of serum PFOA with 

kidney and testicular tumors among highly exposed members of the general population. The 

dose-response data for Leydig cell tumors in rats in was analyzed using the multistage cancer 

model for a dichotomous dataset to predict 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 

study results. The resulting BMDL04 was 1.99 mg/kg/day which yields a HED of 0.58 

mg/kg/day and a slope factor of 0.07 (mg/kg/day)-1. The HED of 0.58 mg/kg/day is nearly 

30,000-fold greater than the RfD. Thus, the proposed RfD, based on increased liver weight as a 

common denominator for loss of homeostasis and protection of co-occurring effects, will also be 

protective of Leydig cell tumors. 
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2.0 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 Electro-Chemical 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). Ammonium perfluorooctanoate (APFO) is the salt 

of PFOA (Figure 2-1) and is a processing aid in the manufacture of certain fluoropolymers, 

especially as an emulsifier in aqueous solution during the emulsion polymerisation of 

tetrafluoroethylene. 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 

(fluorotelomer alcohols, olefins, and perfluoroalkyl sulfonamido substances) (Wallington et al., 

2006). 

 PFOA APFO

 Sources: SIAR (2006); http://commons.wikimedia.org/wiki/File:Ammonium_perfluorooctanoate.png

FIGURE 2-1 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 (Figure 2-2) resulting in a fairly linear molecular shape (Figure 2-3) as the lowest 

energy conformer.

FIGURE 2-2 PFOA Anti-Conformer

COOH

F F F F F F F

F F F F F F F

F
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FIGURE 2-3 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. 

Some commercial PFOA is a combination of linear and branched chain isomers (80% 

linear, 20% branched; Loveless et al., 2006). The samples studied by Loveless et al. (2006) had 

the following mole percents of branched chain isomers: 12.6% internal monomethyl (non-alpha), 

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. 

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

salt APFO are provided in Table 2-1. 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 40-50°C 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. 

The pKa (acid dissociation constant) for PFOA has been reported as 1.5-2.8. 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.
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TABLE 2- 1 Chemical and Physical Properties of PFOA and APFO

Property Perfluorooctanoic Acid Ammonium Perfluorooctanoate

Chemical Abstracts Registry 

(CAS) No.

335-67-1 3825-26-1

Synonyms PFOA; Hexanoyl fluoride, 

3,3,4,4,5,5,6,6,6-nonafluoro-2-oxo-; 

Pentadecafluoro-1-octanoic acid; 

Pentadecafluoro-n-octanoic acid; 

Octanoic acid, pentadecafluoro-; 

Perfluorocaprylic acid;

Pentadecafluoroocanoic acid; 

Perfluoroheptanecarboxylic acid; 

APFO; C-8; Ammonium 

pentadecafluoroctanoate; Ammonium 

perfluorocarpylate; DS 101; FC 1015; 

FC 143; FX 1006; Fluorad® FC 143; 

Perfluorooctanoic acid ammonium 

salt; Unidyne® DS 101-20

Chemical Formula C8HF15O2 C8HF15O2NH3

Molecular Weight 414.09 431.10

Color/Physical State White powder White powder

Boiling Point 189oC Decomposition

Melting Point 45-50 oC Decomposition starts above 105oC

Density (at 20oC) 1.7921 g/cm3 0.6-0.7 g/cm3

Vapor Pressure: 4.2 (25oC)

2.3 (20oC)

128 (59.3oC)

3.7 (90.1oC)

0.0081 (20oC)

pKa 2.5

2.8

1.5-2.8

-

pH value 2.6, 1 g/L (20oC) ~5

Koc 27,000 estimated

Solubility in water (g·L-1) 9.5 (25oC)

4.1 (22oC)

> 500

Solubility in organic solvents - Heptane, Toluene: 0

Methanol, Acetone: > 500

Conversion Factors for vapor 

phase

1 ppm = 17.21 mg/m3 1 ppm = 17.21 mg/m3

Sources: HSDB (2006); SIAR (2006), EFSA (2008); RTECS (2008)
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3.0 TOXICOKINETICS

Because of strong carbon-fluorine bonds, PFOA is stable to metabolic and environmental 

degradation. PFOA, is resistant to biotransformation, making the toxicity of the parent 

compound the concern, not that of a metabolite. However, because of their impact on cellular 

receptors and proteins, they possess the ability to impact the biotransformation of dietary 

molecules, intermediate metabolites, and other xenobiotic chemicals and may alter enzyme 

activities and transport kinetics. PFOA is known to activate the peroxisome proliferator 

activated receptor (PPAR) pathway by increasing transcription of mitochondrial and peroxisomal 

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

based on transcriptional activation of many genes in PPARα null mice, the effects of PFOA 

involve more than activation of PPAR. They may also activate the constitutive androstane 

receptor (CAR), farnesoid receptor (FXR), and pregnane X receptor (PXR).

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 seem to play a role in PFOA absorption, 

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

peptides (OATps) and multidrug resistance-associated proteins (MRPs) (Launay-Vacher et al., 

2006). The transporters play a critical role in gastrointestinal absorption, uptake by the tissues, 

and excretion via the kidney. These transport systems are located at the blood brain barrier, 

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

the brain, fetus, and male reproduction organs (Ito and Alcorn, 2003; Zaïr et al., 2008). There 

are differences in transporters across species, genders, and individuals. There is more PFOAspecific information regarding the OAT and OATp families than the MRPs. Nomenclature 

abbreviations for the various transporters are not totally standardized, and there are 

inconsistencies across individual publications. 

3.1 Absorption

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

animals. There are extensive data 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 human serum data are included in Section 3.5 of this report. 

The absorption processes 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 (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 data that identify 

involvement of OATs, OATps, and MRPs in entrocytes (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). No primary data sources relative to the specificity of these transporters for 

PFOA in humans or laboratory animals were identified. 
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Oral Exposure 

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 3 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/group) 

to PFOA (96% a.i.) 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. 

Steady state serum PFOA levels were reached within 4-6 weeks in cynomolgus monkeys 

dosed with oral capsules containing 3, 10, or 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, or 6.5 mg PFOA/kg for 90 days was 7.1, 41.2, 

70.3, and 137.6 μg/mL, respectively (Palazzolo, 1993). 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 when compared to a female rat given 10 daily doses of 25 mg PFOA/kg (Kennedy et 

al., 2004). Plasma PFOA concentrations in male Sprague-Dawley rats fed a diet containing 300 

ppm PFOA for 1, 7, or 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. Correspondingly, 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 are discussed in more detail in the 

following sections. 

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/sex/group) were exposed to a single nose-only exposure of an aerosol of 0, 1, 10, 

or 25 mg/m3 PFOA. Preliminary range-finding studies demonstrated that aerosol sizes were 1.8 - 

2.0 μm mass median aerodynamic diameter (MMAD) with geometric standard deviations 

ranging from 1.9 - 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 was analyzed 

by liquid chromatography-mass spectrometry (LC-MS). PFOA plasma concentrations were 

proportional to the inhalation exposure concentrations. 

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

female Cmax occurred approximately one 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 six hours. The data are illustrative of absorption of PFOA 

via inhalation. 
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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 or 

2,000 mg/kg. Treatment resulted in elevated blood organofluorine levels that increased in a doserelated manner. O’Malley and Ebbens (1981) treated groups of 2 male and 2 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, 3 of 4 died in the mid-dose group and none from the low-dose group. Although 

these data demonstrate dermal absorption they do not provide quantitative data. 

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 cm/h and 9.49 ± 2.86 x 10-7 cm/h in rat and human skin, 

respectively.

3.2 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, plus 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 3-1. Rat, human, and monkey plasma proteins were able to bind 97 to 100% of the 

PFOA added at concentrations ranging from 1-500 ppm. Human serum albumin 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, inorganic ions, and some pharmaceuticals. Approximately 60% of the serum protein in 

humans and rats is albumin (Saladin, 2003; Harkness and Wagner, 1983). At 68%, the 

percentage bound to albumin in mice is slightly higher than in humans and rats (Harkness and 

Wagner, 1983). 

 

TABLE 3-1. 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

From Kerstner-Wood et al., 2003

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

limit (BQL)<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 
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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 human serum albumin binding using labeled 19F nuclear magnetic resonance 

(NMR) allowed the calculation of disassociation constants for PFOA binding to rodent and 

human serum albumin. No significant difference in binding to the serum albumin of rat versus 

human was detected (Table 3-2). 

 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 3-2. 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

From Han et al., 2003

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 standard deviations 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 microSEC is statistically insignificant.

 Wu et al. (2009) examined the interaction of PFOA and human serum albumin (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 may 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 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 could also be altered. 

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

PFOA to serum albumin (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) 
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while the association constant for the secondary site (K2

a

) was 0.8 ± 0.1 x 102/mol BSA at a 

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

mole of BSA decreased: K1

a

 = 0.33 ± 0.004; 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 the most relevant to the potential physiological impact of 

environmental levels of PFOA. 

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 maximum number of PFOA binding sites on HSA was 9. The site locations 

were common to the natural binding sites for fatty acids, thyroxine, warfarin, indol, and 

benzodiazepine (Figure 3-1; Salvalaglio et al., 2010). The binding site closest to tryptophan 

residue #214 had the highest binding affinity (-8.0 kcal/mol) compared to other sites (~-6--7 

kcal/mol). 

FIGURE 3-1. PFOA binding sites on human serum albumin

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 thyroxine (T4). PFOA was one of the chemicals evaluated. 

 Human TTR was incubated overnight with 125I-labeled T4, unlabled T4, and 10-10,000 

nM 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, 
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characteristics that apply to PFOA. PFOA demonstrated a high binding affinity for TTR with 

949 nM causing a 50% inhibition of T4 binding to the TTR. 

It is also 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 this 

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-dimethylaminonapthalenesulphonyl)-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 to 5% of the 

cytosolic protein in hepatocytes. Luebker et al. (2002) reported that PFOA (IC50>10μM) 

exhibited some binding to L-FABP, but the binding potential was only about 50% of that for 

perfluorooctane sulfonate (PFOS) (IC50 = 4.9 μM) and far less than that of oleic acid (IC50 = 0.01 

μM).

Oral Exposure 

Tissue distribution 

Human. No clinical studies are available in humans on administration of a controlled dose of 

PFOA or its distribution. However, samples collected in biomonitoring and epidemiological 

studies provide data showing distribution of PFOA. Olsen et al. (2001c) 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 ng/mL. PFOA 

concentrations above the LOQ were detected in 5/6 postmortem liver samples from men 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.

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. 

Monkey. Butenhoff et al. (2002; 2004b) studied the fate of PFOA in cynomolgus monkeys in a 

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

PFOA daily via oral capsule at dose rates of 0, 3, 10, or 20 mg/kg. The highest concentration 

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 six month period, 

resulting in a normalized dose of 20 mg/kg/day for the study. Serum, urine and fecal samples 

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

were collected at time of sacrifice. 
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Serum concentrations reached steady-state levels within four to six weeks in all dose 

groups. Steady-state concentrations of PFOA in serum were 77 ± 39, 86 ± 33, and 158 ± 100 

μg/ml after six weeks (Butenhoff et al., 2002) and 81 ± 40, 99 ± 50, and 156 ± 103 μg/ml 

(Butenhoff et al., 2004b) after six months for the 3, 10, and 20 mg/kg dose groups, respectively. 

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 21.9 μg/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/sex/group) doses of 3, 10, 

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

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 quantitation limit of the FID (1 μg/ml). 

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

is presented in Table 3-3. 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 

testes, 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 had similar concentrations followed by lower levels in the lung and brain.

TABLE 3-3.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

FromYlinen et al., 1990 a

 N = 6, Mean ± SD, μg/ml tissue b N=3, no SD c below the limit of quantitation of 1μg/mL
Perfluorooctanoic Acid – February 2014 3-8 

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Kemper (2003) examined the distribution of PFOA in tissues of male and female 

Sprague-Dawley rats following administration by gavage. Rats were administered 1, 5, and 25 

mg/kg 14C-PFOA 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 pharmacokinetic experiments. In those experiments, plasma was collected over 

the course of several days and PFOA concentration was analyzed. Non-compartmental 

pharmacokinetic 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 may 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 3-4 and Table 3-5 for males and females, 

respectively. Liver, kidney and blood were the primary tissues for distribution of 14C-PFOA. In 

males the fraction of the dose found in 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. Blood to kidney 

concentration ratios in females were between 1 and 2 at all dose levels and remained relatively 

constant between Tmax and Tmax/2. In males the blood to kidney ratios were 10 or higher at Tmax

and declined slightly at Tmax/2. 
Perfluorooctanoic Acid – February 2014 3-9 

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TABLE 3-4. 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

From Kemper, 2003

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.

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

(171 hours) indicate that the 40 to 60% percent of the dosed PFOA retained was present in the 

liver, blood, skin, and muscle tissues in decreasing amounts (Table 3-4). 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. 

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

retained was present in the liver, blood, kidney, muscle, and skin tissues in decreasing amounts 

(Table 3-5). 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. 
Perfluorooctanoic Acid – February 2014 3-10 

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 TABLE 3-5. 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

From Kemper, 2003 

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.

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, or 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. The mean liver 

PFOA concentration was 92 ± 6, 250 ± 32, and 243 ± 23 μg/g after 1, 3, or 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. 

Mouse. Lau et al. (2006) gavage dosed adult male and female CD-1 mice (5-7/group) with 20 

mg/kg for 7 or 17 days. The animals were sacrificed 24 hours after the last treatment. After 

seven days of treatment male mice had serum PFOA levels of 181 μg/mL and females had 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. 
Perfluorooctanoic Acid – February 2014 3-11 

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As part of a physiologically based pharmacokinetic (PBPK) modelling exercise, Lou et 

al. (2009) administered single doses of 1 or 10 mg/kg to groups of 3 male and 3 female CD-1 

mice. The mice were sacrificed for analysis of plasma, liver, and kidney tissues after 4, 8, or 12 

hours and at 1, 3, 6, 9, 13, 20, 27, 34, or 48 days after dosing. This study was repeated for a 

second analysis which 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 was generally reflected in an order of magnitude 

difference between serum concentration for 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 concentrations 

were 10 and 100 mg/L for the 1 and 10 mg/kg/day, respectively. Declines in serum 

concentrations for females 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 were also higher in males than females for each of the two doses. 

 Lou et al. (2009) also collected serum data for up to 28 days after administration of a 60 

mg/kg dose to groups of 3 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 10 mg/kg dose, and 70 days for the 1 mg/kg dose (values estimated from 

the Figure 3 of the published paper). 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 5 

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, or 50 μmol/kg PFOA (~0, 5.4, 10.8, 

or 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 3-6. 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 38 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 the existence of a capacity-limited and PPARα independent PFOA 

transport from the liver to the bile duct. 
Perfluorooctanoic Acid – February 2014 3-12 

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TABLE 3-6. PFOA Concentration in Wild-type and PPARα-null Mice

PFOA Concentration (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 12.9 ± 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

From Minata et al., 2010

Mean ± SD; ND= not detected (<0.001 μg/mL)

Tissue transporters. The liver is an important binding site for PFOA. Increased liver weight is 

one of the early, low-dose manifestations of exposure. OATps and MRPs, at least one OAT and 

the sodium-taurocholate cotransporting polypeptide (NTCP) have been identified at the 

interphace of the liver with the portal blood and the canalicular membranes within the liver 

(Kim, 2003; Kusuhara and Sugiyama, 2009; Zaïr et al., 2008). OATs and MRPs have been 

identified along the hepatocyte interface with systemic circulation.

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 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 

two days after treatment as reflected in quantification of their 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. Conversely, there were 

significant decreases (p<0.05) in the protein levels for OATp1a1, OATp1a4, and OATp1b2 

(Cheng and Klaassen, 2008) as determined by Western Blot analysis and mRNA measurements. 

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, 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. 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 four to eight week old 

Sprague-Dawley rats (10/sex/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 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 3-7). In the 5 and 6 week old 

female rats, the plasma PFOA concentrations were about 2 fold lower than in the 4-week old 

rats. However, in the 5 week old males, the concentration of plasma PFOA was about 5-fold 

higher than in the 4 week 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 sex
Perfluorooctanoic Acid – February 2014 3-13 

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but did differ between sexes. 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 

factor(s) responsible for the gender difference in elimination occurs between the ages of 4 and 5 

weeks in the rat. 

TABLE 3-7. 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

From Han, 2003

a mean ± SD; samples from 10 animals/sex/group

Hinderliter (2004, 2006a) continued the investigation of the relationship between age and 

plasma PFOA 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. Rats were not 

fasted prior to dosing. Two hours after dosing, 5 rats/sex/ per age group and dose group were 

sacrificed and blood samples were collected. The remaining 5 rats/sex/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 or 5 

weeks), except at 2 hours for the five week age group (p<0.01) which showed the lowest PFOA 

level (Table 3-8). 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 age groups at 24 hours after dosing and thus do not 

support the observations from the Han (2003) study. 

TABLE 3-8. 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

From Hinderliter 2004

*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 

age group than in the 3- or 4-week age groups at the 24-hour time period for both doses and at 

the 30 mg/kg dose group at 2-hours (Table 3-9). Plasma PFOA concentrations following a 30 

mg/kg dose were approximately 1.5 to 4 times higher than those observed following a 10 mg/kg 

dose. At 24 hours post dose, plasma PFOA levels in the female rats was significantly lower than 

the plasma PFOA levels in male rats, especially at 5 weeks of age. The data for the 5-week old 
Perfluorooctanoic Acid – February 2014 3-14 

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female rats compared to the 3- and 4-week age groups at 24 hours are consistent with the Han 

(2003) data 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 3-9. 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

From Hinderliter, 2004

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).

In a supplemental study to determine the effect of fasting (Hinderliter, 2004, 2006a), 4 

week old rats, 4 rats/sex, were administered 10 mg/kg PFOA via oral gavage. Animals (2/sex) 

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 non-fasted animals, 

respectively. Plasma PFOA concentrations in the female rats were 68.16 and 26.54 μg/ml for the 

fasted and non-fasted 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. 

Distribution during pregnancy and lactation 

Rat. An oral two-generation reproductive toxicity study of PFOA in rats was conducted (York, 

2002; Butenhoff et al., 2004a). Five groups of rats (30 sex/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/sex/group-control; 10/sex/group-treated) were collected from animals dosed with 0, 

10, or 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 the limits of quantitation (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 or 30 

mg/kg-day PFOA during days 4-10, 4-15, or 4-21 of gestation, or from gestation day 4 to 
Perfluorooctanoic Acid – February 2014 3-15 

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lactation day 21. Maternal blood samples were collected at 2 hours ± 30 minutes post-dose 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 gestation day (GD) 10 only embryos were recovered and on GD 15 and 21 

the placentas, amniotic fluid, and embryos/fetuses were collected. 

The remaining 5 rats per group were allowed to deliver their pups. On lactation days 0, 3, 

7, 14, and 21, the pups were counted, weighed (sexes separate), and examined for abnormal 

appearance and behavior. Randomly selected pups were sacrificed and blood samples were 

collected. On lactation days 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 3-10. Maternal plasma levels at 2 hrs 

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 

limit of quantitation (0.05 μg/ml). 

TABLE 3-10. Maternal Plasma PFOA Levels (μg/ml) 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

From Hinderliter et al., 2005 and Mylchreest, 2003

Mean ± SD; samples were from 5 dams/group/time point and were collected 2 hrs post-dosing 

PFOA levels in the placenta, amniotic fluid, and embryo/fetus are presented in Table 3-

11. The levels of PFOA in the placenta on gestation day 21 were approximately twice the levels 

observed on gestation day 15, and the levels of PFOA in the amniotic fluid were approximately 

four times higher on day 21 than on day 15. The concentration of PFOA in the embryo/fetus was 

highest in the day 10 embryo and lowest in the day 15 embryo; PFOA levels in the day 21 fetus 

were intermediate. 
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TABLE 3-11. Placenta, Amniotic Fluid, and Embryo/Fetus PFOA Concentrations (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

From Hinderliter et al., 2005 and Mylchreest, 2003

Mean ± SD; samples were pooled by litter and were collected 2 hrs post-dosing 

The concentrations of PFOA in the plasma of the day 21 fetus were approximately half 

the levels observed in the maternal plasma (Table 3-10). 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 in the, 

respectively for the pups. Pup plasma levels decreased between birth and lactation day 7 (Table 

3-12), and were thereafter similar to the levels observed in the milk (see Table 3-13 below). The 

pups were not separated by gender. 

TABLE 3-12. Fetus/Pup PFOA Concentration (μg/ml) 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

From Hinderliter et al., 2005 and Mylchreest, 2003

Mean ± SD; samples were pooled by litter and were collected 2 hrs post-dosing 

The concentration of PFOA in the milk was also fairly similar throughout lactation and 

was approximately 1/10th of the PFOA levels in the maternal plasma (see Table 3-11 above); 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 3-13). 

TABLE 3-13. PFOA Levels (μg/ml) in Sprague-Dawley 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

From Hinderliter et al., 2005 and Mylchreest, 2003

Mean ± SD; samples were from 5 dams/group/time point and were collected 2 hrs post-dosing 

Mouse. Fenton et al. (2009) orally dosed pregnant CD-1 mice (n=25/group) with 0, 0.1, 1, or 5 

mg PFOA/kg on GD17. On GD18, 5 dams/group were sacrificed and trunk blood, urine, 

amniotic fluid, and the 4th and 5th mammary gland were collected. One fetus/dam was 
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euthanized and retained for whole pup analysis. The remaining dams were allowed to litter. 

Biological samples as described above excluding amniotic fluid were also collected on postnatal 

days 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 postnatal days 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 3-14). 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 to 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 PND18 showing an inverse Ushaped response curve.
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TABLE 3-14. PFOA Levels (ng/ml) During Gestation and Lactation in Selected Fluids and Tissues

Tissue Day Dose

0.1 mg/kg 1 mg/kg 5 mg/kg

Dam Serum GD18 143 ± 19 1697 ± 203 7897± 663

PND1 217.5 ± 35 1957.0 ± 84 9845.6 ± 1478

PND4 110.0 ± 12 1269.4 ± 235 6776.6 ± 561

PND8 46.7 ± 21 360.8 ± 98 1961.8 ± 414

PND18 123.3 ± 41 1035.2 ± 305 5156.5 ± 1201

Amniotic Fluid GD18 99.0 ± 28 865.3 ± 191 3203.8 ± 492

Dam Urine GD18 21.9 ± 8.6 104.9 ± 69.7 666.7 ± 169

PND1 7.7 ± 1.7 116.8 ± 64 492.3 ± 119

PND4 8.4 ± 6.4 53.5 ± 15 401.5 ± 117

PND8 0.8 ± 0.22 11.6 ± 6.2 40.1 ± 17

PND18 1.8 ± 1.1 18.7 ± 8.6 91.7 ± 49

Mammary Gland GD18 18.9 ± 1.9 307.2 ± 30.4 1429± 186

PND1 27.4 ± 6.8 343.8 ± 53 1933.5 ± 194

PND4 9.6± 8.4 239.2 ± 53 1461.8 ± 267

PND8 2.4 ± 3.8 71.7 ± 22 411.8 ± 78

PND18 17.1 ± 10 239.9 ± 76 1372.8 ± 240

Milk PND2 32.5 ± 12 716.7 ± 145 1236.6 ± 1370

PND8 11.6 ± 8.1 77.4 ± 19 245.1 ± 26

PND11 5.4 ± 1.0 42.3 ± 9.1 282.5 ± 162

PND18 43.5 ± 19 251.8 ± 147 909.8 ± 308

Whole Pup GD18 136.3 ± 15 1665.8 ± 213 6256.5 ± 751

PND1 150.9 ± 21 1606.9 ± 288 7134.5 ± 1097

PND4 91.8 ± 8.9 1183.2 ±187 5071.4 ± 267

PND8 60.9 ± 16 729.0 ± 92 3118.5 ± 424

PND18 17.5 ± 11 251.9 ± 112 1391.5 ± 118

Pup Serum PND1 324.7 ± 36 3926.8 ± 480 16,286.4 ± 1372

PND4 267.6 ± 47 3020.8 ± 223 11,925.2 ± 1077

PND8 260.2 ± 56 2548.2 ± 245 9215.8 ± 594

PND18 111.8 ± 30 1124.8 ± 236 5894.3 ± 743

From Fenton et al., 2009

 Pregnant C57BL/6/Bkl mice were fed diets containing 0.3 mg PFOA/kg/day from GD1 

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). 

 Macon et al. (2011) gavage dosed CD-1 mice with 0, 0.3, 1.0, or 3.0 mg PFOA/kg from 

GD1 to GD17 or with 0, 0.01, 0.1, or 1.0 mg PFOA/kg from GD10 to GD17. In the full 

gestation experiment (GD1-17), offspring were were sacrificed on postnatal day (PND) 7, 14, 21, 

28, 42, 63, and 84, and in the half gestation experiment (GD10-17), female offspring were 

sacrificed on PND1, 4, 7, 14 and 21. Serum, liver, and brain from the offspring were analyzed 

for PFOA by HPLC-MS/MS. 
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 At the lowest dose, PFOA concentration in the serum peaked around PND7, but the two 

higher doses peaked around PND 14 (Table 3-15). Calculated blood burdens which take into 

account the increasing blood volumes and body weights for females showed an inverted Ushaped curve peaking at PND14 for all doses. In the liver, PFOA concentration decreased over 

time with the highest concentration observed at PND7. Lower concentrations of PFOA were 

detected in the brain of the offspring on PND 7 and 14. 

 

TABLE 3-15. Female Offspring PFOA Levels (ng/ml) After GD1-17 Exposure

Tissue Day Dose

0.3 mg/kg 1.0 mg/kg 3.0 mg/kg

Female Serum PND7 4980 ± 218 11026 ± 915 20700 ± 3900

PND14 4535 ± 920 16950 ± 3606 26525 ± 2446

PND21 1194 ± 394 377 ± 607 8343 ± 1078

PND28 630 ± 162 1247 ± 208 4883 ± 1378

PND42 377± 81 663 ± 185 2058 ± 348

PND63 55 ± 17 176 ± 85 ‒

PND84 16 ± 5 71 ± 8 125 

Female Liver PND7 2078 ± 90 8134 ± 740 16700 ± 749

PND14 972 ± 124 4152 ± 483 10290 ± 1028

PND21 1188 ± 182 1939 ± 637 2339 ± 1241

PND28 678 ± 130 2007 ± 560 7124 ± 1081

PND42 342 ± 87 617 ± 145 1145 ± 274

PND63 118 ± 22 320 ± 113 417 ± 160

PND84 43 ± 12 55± 12 235 ± 79

Female Brain PND7 150 ± 26 479 ± 41 1594 ± 162

PND14 65 ± 12 241 ± 20 650 ± 44

PND21 <LOQ 31 ± 5 133 ± 23

PND28 <LOQ <LOQ 62 ± 93

PND42 <LOQ <LOQ <LOQ

PND63 <LOQ <LOQ <LOQ

PND84 <LOQ <LOQ <LOQ

From Macon et al., 2011 

‒ = 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 GD10 to GD17, serum PFOA 

concentration in the female offspring declined from PND1 through the end of the experiment 

(Table 3-16). Calculated blood burden showed a gradual increase from PND1 through PND14, 

followed by a decline through PND21. 
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TABLE 3-16. Female Offspring Serum PFOA Levels (ng/ml) After GD10-17 Exposure

Tissue Day Dose

0.01 mg/kg 0.1 mg/kg 1.0 mg/kg

Female Serum PND1 284.5 ± 21.0 2303.5 ± 114.4 16305.5 ± 873.5

PND4 184.1 ± 12.1 ‒ ‒

PND7 150.7 ± 20.9 1277.8 ± 122.6 11880.3 ± 1447.6

PND14 80.2 ± 13.9 645.4 ± 114.2 6083.7 ± 662.6

PND21 16.5 ± 2.1 131.7 ± 24.5 2025.1 ± 281.9

Female Blood 

Burden (calculated)

PND1 15.2 ± 1.7 114.3 ± 5.4 926.0 ± 47.6

PND4 20.6 ± 0.1 ‒ ‒

PND7 27.3 ± 3.8 221.7 ± 24.9 1965.9 ± 256.7

PND14 27.0 ± 4.6 218.5 ± 39.8 2033.6 ± 293.5

PND21 7.9 ± 1.0 66.4 ± 12.8 984.7 ± 142.8

From Macon et al., 2011 

‒ = 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 3-17). Pregnant mice (F0, n=10-12 dams/group) were gavage dosed with 0, 1, or 5 mg 

PFOA/kg from GD1-17. A separate group of pregnant mice (n=7-10 dams/group) were gavage 

dosed with either 0 or 1 mg PFOA/kg from GD1-17 and received drinking water containing 5 

ppb PFOA beginning on GD7 and continuing until the end of the study for their offspring, except 

during breeding and early gestation, to simulate a chronic low-dose 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 + 5ppb PFOA group of the F2 generation. A decrease was 

observed for the remaining groups. 

TABLE 3-17. Serum PFOA Levels (ng/ml) 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/ PND22 74.8 ± 11.3 6658.0 ± 650.5 4772.0 ± 282.4 26980.0 ± 1288.2

F1/~PND91 86.9 ± 14.5 9.3 ± 2.6 173.3 ± 36.4 18.7 ± 5.2

Offspring F1/PND22 21.3 ± 2.1 2443.8 ± 256.4 2743.8 ± 129.7 10045 ± 1125.6

F1/PND42 48.9 ± 4.7 609.5 ± 72.2 558.0 ± 55.8 1581.0 ± 245.1

F1/PND63 66.2 ± 4.1 210.7 ± 21.9 187.0 ± 24.1 760.3 ± 188.3

F2/PND22 26.6 ± 2.4 4.6 ± 1.2 28.5 ± 3.7 7.8 ± 1.9

F2/PND42 57.4 ± 2.9 0.4 ± 0.0 72.8 ± 5.8 0.4 ± 0.0

F2/PND63 68.5 ± 9.4 1.1 ± 0.5 69.2 ± 4.3 1.2 ± 0.5

From 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 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 and cell 

debris, lysosome and mitochondria, microsome, light microsome and ribosome, and membranefree cytosol) were obtained by differential centrifugation. The radioactivity per gram of each 

fraction and the total radioactivity were measured. 
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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% TRR, lysosome and mitochondria ~14% TRR, 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. 

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 sexes 98% of PFOA in the plasma was protein bound and 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 the females. 

Based on the results, the authors concluded that subcellular distribution of PFOA in the 

rat liver was sex-dependent because the proportion of PFOA in the liver cytosol of female rats 

was almost twice that in the male. They hypothesized that the female may have a greater amount 

than males 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 thefemales (17%).

Inhalation Exposure 

In a repeated exposure study, Hinderliter (2003, 2006b) exposed 6-8 week old male and 

female rats (5/sex/group) to 0, 1, 10, or 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 three days per week. The aerosols had mass median aerodynamic diameters 

(MMAD) of 1.3 - 1.9 μm with geometric standard deviations 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 three 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 daily exposure, plasma levels had returned to baseline in females demonstrating 

clearance within 24 hours of each daily dose. 
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Dermal Exposure 

No data were identified on tissue distribution following dermal exposures. 

3.3 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 animo 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 is secreted in the 

urine as such and has a neglible rate of phase II metabolism (Ylinen et al., 1989). The potential 

for reduction of the carboxyl group does not appear to have been evaluated but is not likely to 

occur.

3.4 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 

(breastmilk) (Apelberg et al., 2007; Tao et al., 2008; Thomsen et al., 2010; Völkel et al., 2008; 

von Ehrenstein et al., 2009). 

Urinary elimination of PFOA was reported in a case history of of a single human male 

with high serum levels of perfluorinated chemicals (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 perfluorohexanesulfonic acid 

(PFHxS) (58-46.8 ng/g serum) which were present at higher levels in the serum. This 

observation that suggests excretion with bile and possible enterohepatic resorption via intestinal 

transporters limiting the loss of absorbed PFOA via feces in the absence of a binding agent such 

as cholestyramine.

Elimination half-lives differ among species. There are also significant gender differences 

in the elimination of PFOA in some laboratory animal species. Information from humans does 

not at present provide sufficient data to determin the magnitude of interindividual and gender 

differences in excretory half lives. The organic acid transport system appears to play an 

important role in renal excretion of PFOA and possibly biliary elimination. 
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Oral Exposure 

Monkey. Butenhoff et al. (2004b) studied the fate of PFOA in cynomolgus monkeys in a sixmonth oral exposure study. Groups of four to six male monkeys each were administered PFOA 

daily via oral capsule at dose rates of 0, 3, 10, or 30/20 mg/kg for six 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 two-week 

intervals and were analyzed for PFOA concentrations. 

 

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 steady-state 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, 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, 5 rats/sex/age group and dose group were sacrificed and blood samples were 

collected (See section 3.2). The remaining 5 rats/sex/ 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 sex (p<0.01, Table 3-

18). Female rats had higher urine PFOA concentrations than 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 sexes, 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 and 4/5 week male rats with 

the older rats excreting ~50% less PFOA in the urine compared to 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 3-8), 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 3-18) and 

plasma concentrations decreased (Table 3-9). 
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TABLE 3-18. 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

From 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 

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 3-19. 

TABLE 3-19. 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

From 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/group) to PFOA 

(96% a.i.) 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 
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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. 

Hormonal changes during pregnancy do not appear to cause a change in the rate of 

elimination of 14C after oral administration of a single dose of 14C-PFOA (Gibson and Johnson, 

1983). At 8 or 9 days after conception, four pregnant CD rats and two nonpregnant female CD 

rats were given a mean dose of 15 mg/kg 14C-PFOA. Individual urine samples were collected at 

12, 24, 36, and 48 hours post dose and analyzed for 14C content. Essentially all of the 14C was 

eliminated via the urine within 24 hours for both groups of rats 

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 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% and 8.3% 14C-PFOA, and the female 

mouse excreted 6.7% and 5.7% 14C-PFOA in urine and feces, respectively. The mice were 

similar in amount excreted. The male hamster excreted 90.3% and 8.2% 14C-PFOA, and the 

female hamster excreted 45.3% and 9.3%14C-PFOA in urine and feces, respectively. The male 

hamster excreted a greater amount of 14C-PFOA than the female hamster. Over 84% of 14CPFOA 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, and the female rabbit excreted 87.9% and 4.6% 14CPFOA in the urine and feces, respectively. Both rabbits excreted most of the dose by 24 hours. 

The cumulative percentage of 14C-PFOA excreted is shown in Table 3-20. 

TABLE 3-20. Cumulative Percent 14C-PFOA Excreted in Urine and Feces

Species Sex

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

From Hundley et al., 2006

Inhalation Exposure

Although no data were identified on urine or fecal excretion of PFOA following 

inhalation exposures, the Hinderliter (2003) study provides evidence of clearance following 

single and repeated inhalation exposures in Sprague-Dawley rats. Plasma PFOA concentrations 
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following a single exposure to 1, 10, or 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 the plasma levels had dropped below the 

analytical limit of quantitation (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 levels. In the repeated exposure study, male and female rats were 

exposed to the same concentrations for 6 hr/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. Little to no fecal 

excretion is anticipated for the dermal route of exposure. 

3.4.1 Mechanistic Studies of Renal Excretion

 

Several studies have been conducted to elucidate the cause of the gender difference in the 

elimination of PFOA for rats. Many of the studies have focused on the role of transporters in the 

kidney tubules. Most studies have examined the OAT transporters located in the proximal 

portion of the descending tubule. They are located 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 its binding to surfaces of 

serum proteins (particularly albumin) make 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 the kidney. 

 

OATs are located on both the basolateral (serum interphase) 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; Nakagawa et al. 2007, 2009; Klaassen and Lu 2008). OATs 1, 2, and 

3 are located on the basolateral membrane surface. The OAT 4 and 5 transporters are located on 

the apical surface of the tubular cells where they function to reabsorb the PFOA anions from the 

glomerular filtrate. Figure 3.2 diagrams the flow of organic anions such as 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. 

There are several MRP transporters that also appear to function in the kidney and move 

organic anions in and out of cells at both the basolateral (e.g. MRP2/4) and apical (e.g. MRP1) 

surfaces (Launay-Vacher et al., 2006; Klaassen and Lu, 2008; Cheng and Klaassen, 2009; 

Kusuhara and Sugiyama, 2009) and one or more OATps on each surface (Cheng and Klaassen, 
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2009; Kusuhara and Sugiyama, 2009; Yang et al., 2009). Bidirectional movement of PFOA 

across both the basolateral and apical surfaces can occur as driven by concentration gradients 

and/or active transport. There are far more data on the PFOA and OATs in the kidneys than on 

OATps and MRPs. Abreviations for individual transporters on the basolateral and apical 

surfaces differ across publications. 

FIGURE 3-2. Renal Organic Acid Transporters

From Anzai et al., 2006; Luanay-Vacher et al., 2006; Cheng and Klaassen, 2006; Klaassen and Lu, 2007; Cheng et al., 2008; Zaïr 

et al., 2008; Kusuhara and Sugiyama, 2009

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, while males excreted only 7.8% of 

the administered dose over a 7-hr period. Probenecid administration modified the cumulative 

excretion curve for males only slightly. However, in females probenecid markedly reduced 

PFOA elimination to 11.8%. The authors concluded that the female rat possesses an active 

secretory mechanism which rapidly eliminates PFOA from the body while male rats do not. 

 Kudo et al. (2002) examined the role of sex hormones and organic anion transporters on 

the renal clearance (CLR) of PFOA. Renal mRNA levels of specific organic anion transporters in 

castrated male and ovariectomized female Wistar rats were also 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 (2-fold 
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increase), but the administration of estradiol to ovariectomized 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 OATs 1-6 and 8, 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 receptors.

To identify the transporter molecules that are responsible for PFOA transport in the rat 

kidney, renal mRNA levels of specific organic anion transporters were determined in male and 

female rats under various hormonal states and compared with the CLR of PFOA (Kudo et al. 

2002). The level of OAT2 mRNA in male rats was only 13% that 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 the organic anion-transporting polypeptide (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.

 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 these 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 that in the females. 

Treatment with E2 almost abolished the expression of OATp1a1 in the kidneys and caused no 

significant change in OATp3a1. In the intact control animals there was almost no expression of 

OATp1a1 in the kidneys of females and a significantly lower (p<0.05) expression of OATp3a1. 

In the gonadectomized control animals, there was litle or no expression of OATp1a1 in either 

sex, and expression of OATp3a1 was equivalent in both sexes. 

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, 

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 only served as control cells. The transfected HEK293 

cells were incubated for one minute with or without 0, 10, or 100 μM [14C]PFOA and/or varying 

concentrations of favored OAT substrates as follows: 5 μM [14C]para-aminohippuric acid 

(OAT1), 20 nM [14C]estrone sulfate (OAT3), or 10 nM [14C]prostaglandin F2α (OAT2) to 

determine inhibitory effects of PFOA. 

PFOA significantly inhibited para-aminohippuric acid and estrone sulfate uptake 

mediated by OAT1 and OAT3, respectively. At 10 μM PFOA, uptake of 5 μM [14C]paraaminohippuric acid was inhibited by 75-85% compared to control, and 100 μM PFOA inhibited 

it to 35-45% of control. Estrone sulfate uptake by human OAT3 was inhibited by 65% compared 

to control at 10 μM PFOA and 40% of control at 100 μM PFOA. Estrone sulfate uptake by rat 
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OAT3 was inhibited to 15% of control in the presence of 10 μM PFOA and by 100% at 100 μM 

PFOA. Prostaglandin F2α uptake by OAT2 was inhibited moderately by PFOA, 75-85% of 

control at 10 μM PFOA and 65% of control at 100 μM PFOA. 

 In the second part of their study Nakagawa et al. (2007) incubated HEK293 and S2 

transfected cells with 10 μM [14C]PFOA for 1 minute to determine uptake. Time dependent 

uptake of 5 μM [14C]PFOA from 0-30 minutes 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 minutes and 

reached a plateau in about 15 minutes. Kinetic evaluations resulted in Km (substrate 

concentration at which the initial reaction rate is half maximal) values of 48.0, 51.0, 49.1, and 

80.2 μM 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 does not. 

Yang et al. (2009) investigated the role of organic anion transporter polypeptide 1a1 

(OATp1a1) in the renal elimination of PFOA. The polypeptide is located on the apical side of 

proximal tubule cells and may 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 cDNA. The transfected 

CHO cells were incubated with 4 μM [14C]PFOA for up to 10 minutes or with 0-1000 μM 

[

14C]PFOA for 2 minutes 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 μM 

[

14C]PFOA for 2 minutes in the absence or presence of inhibitors (sulfobromophthalein, 

taurocholate, probenecid, p-aminohippurate, and naringin) for inhibition studies. The transfected 

CHO cells were incubated with 2 μM 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 minutes of incubation. Vector-transfected cells 

had a significant level of uptake of PFOA attributed to non-specific 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, was described by the MichaelisMenton equation and exhibited saturable kinetics. 

Inhibition experiments with substrates of OATs and OATps showed that 

sulfobromophthalein, taurocholate, and naringin inhibited PFOA uptake to 10-30% of control 
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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 

may be an important route of PFOA distribution, and that renal reabsorption in the male rat may 

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 re-absorption of organic anions. Transformed cells derived from 

human embryonic kidney 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 μM [14C]PFOA for 15 minutes to determine 

uptake. Transfected cells were also incubated with 10 μM [14C]PFOA for 15 minutes and then 

washed with incubation medium containing 1%, 3%, and 5% BSA to investigate the contribution 

of non-specific 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 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 was also 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 may 

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 URAT1 

to determine their roles in mediating human renal reabsorption. Chinese hamster ovary (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) or 30 

seconds (URAT1). Cells transfected with OAT4 were incubated with 5 μM PFOA for up to 1 

minute in time-dependent uptake experiments. In inhibition studies, cells transfected with OAT4 

were incubated with 5 μM C14-PFOA for 10 seconds in the presence and absence of 100 μM 

BSP, probenecid, glutarate, or polycyclic aromatic hydrocarbon (PAH). Perfluorochemicals with 

differing chain lengths (C4-C12) were used in chain-length dependent inhibition experiments. 

Incubations with 3H-E3S (OAT4, OATp1A2) or 6 μM C14-uric acid(URAT1) in the presence and 

absence of 100 μM perfluorinated carboxylate acids (PFCs) lasted 10 seconds (OAT4), 30 

seconds (OATp1A2), or 1 minute (URAT1). 

 

PFOA uptake in OATp1A2 transfected HEK293 cells was not different than uptake in 

control cells. At 100 μM PFCs, 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 
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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 μM) 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 4-

fold compared to uptake in control cells. Time-dependent PFOA (5 μM) uptake by URAT1 

increased with time during the 5 minute incubation period, and a concentration-dependent 

increase in PFOA uptake was observed (0-700 μM). Urate uptake was inhibited in a chain 

length-dependent 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, the authors concluded that PFOA was not a 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 PFCs having chain lengths ranging from 2 

to 18 carbons (C). The activity of the PFC on the transporters was quantified on the basis of its 

ability to inhibit the transport of a favored radiolabeled substrate. The PFC inhibition of the 

individual transporters varied with chain length. The PFCs 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 PFCs with 5- through 10- carbon chains caused a significant decrease 

in transport of tritiated estrone-3-sulfate by OAT3 with C8 and C9 acids having the strongest 

effect. The transport of tritiated estadiol-17β-glucuronide by OATp1a1 was significantly 

inhibited by PFCs with 6 through 11 carbon chains with C10 acid having the strongest effect. 

The PFCs 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 PFCs was also evaluated by Weaver et al. (2010). The change in 

transport velocity (ng/mg protein/min) with increasing concentrations of the PFCs exhibited a 

Michaelis-Menton type response. The kinetic data were analyzed to determine the Km and Vmax, 

and those data are summarized in Table 3-21 below. 

TABLE 3-21. Kinetic Parameters of PFC Transport by OAT1, OAT3 and OATp1a1

Transporter PFC Km (μM) Vmax (nmol/mg protein/min)

OAT 1 C7 50.5 ± 13.9 2.2 ± 0.2

C8 43.2 ± 15.5 2.6 ± 0.3

OAT 3 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

From Weaver et al., 2010
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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 PFCs 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 of the C7 and C8 acids by OAT1, an 

export transporter on the basolateral surface of the renal tubular cells. 

Based on the Hinderliter (2004) study, 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 3-22). 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 ovariectomized 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 that in males seems to involve the OATps (Cheng et al., 2006). 

TABLE 3-22. 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

From 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 for the smaller amount they can transport into the glomerular filtrate via 

OATp1a1 in the apical membrane than females. This scenario may explain the inverse 

relationship between the levels of PFOA in female urine and plasma and the plateau of plasma 

PFOA in male rats when compared to their losses via urine. 

Unfortunately much work remains to be done in order to explain the gender differences 

between male and female rats and to determine whether it is relevant to humans. Similarities are 

possible because of the long half-life in humans suggests that they may be more like the male rat 

rather than the female rat. There is a broad range of half-lives in human epidemiology studies 

suggesting a variability in human transport capabilities resulting from genetic variations in 

structures and consequently in function. Genetic variations in human OATs and OATps have 

been identified as described in a review by Zaïr et al. (2008). 
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3.5 Toxicokinetic Considerations 

3.5.1 PK Models

One of the earliest PK models (Andersen et al., 2006) was done using the post-dosing 

plasma data from the Butenhoff et al. (2004b) study in cynomolgus monkeys. In this study, 

groups of 6 monkeys (3/sex/group) were dosed for 26 weeks with 0, 3, 10, or 20 mg/kg PFOA 

(high-dose =30 mg/kg PFOA for the first 12 days) and 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. 

The Andersen et al. (2006) model 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 (Figure 3-3; Andersen et al.2006) was derived from a 

published model for glucose resorption from the glomerular filtrate via transporters on the apical 

surface of renal tubule epithelial cells. 

FIGURE 3-3. Schematic for a physiologically-motivated renal resorptions pharmacokinetic 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., 2004) 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 where renal resorption was saturated. Based on the data for the dose of 20 
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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. This could be 

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 renal clearance of PFOA than the other three monkeys. 

 Tan et al. (2008) divided the second compartment in the Anderson et al. (2006) model 

into a liver compartment and a tissue compartment. A storage compartment was added between 

the filtrate compartment and urinary excretion (Figure 3-4; Tan et al., 2008). 

FIGURE 3-4. Physiologically-motivated pharmacokinetic model schematic for PFOA-exposed rats

The models were parameterized and applied to the Kemper (2003) data for male and 

female CD rats given doses of 1, 5, or 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 sex. 

 Lou et al. (2009) utilized the data they collected on the serum, liver, and kidney PFOA 

concentration (see section 3.2) in CD-1 mice to examine if one (Figure 3-5) or two compartment 

pharmacokinetic (PK) models would fit the experimental data for 1, 10, and 60 mg/kg/day single 

gavage doses. 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 squares approach. A likelihood ratio squared approach was applied to determine 

which model achieved the best fit to the data. 
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FIGURE 3-5. 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 volume of distribution (Vd), serum half-life, 

and absorption (Ka) and elimination (Ke) rate constants 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 3-23). The quantitative measures for 

liver and kidney were only available for the 1 and 10 mg/kg/day doses. 

TABLE 3-23. 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

Adsorption rate constant (Ka) 0.537 L/hr

Elimination rate constant (Ke) 0.00185 L/hr 0.00133 L/hr

Half life (T 1⁄2) 15.6 days 21.7 days

Liver Volume of distribution (Vd) 0.161 L/kg 0.120 L/kg

Adsorption 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

Adsorption rate constant (Ka) 0.527 L/hr

Elimination rate constant (Ke) 0.00151 L/hr 0.00113 L/hr

From 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. This suggests 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 

Central Compartment

Vd, C1, Free PFOA

Oral Dose

Ka

Urinary Loss

Ke
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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 60 mg/kg dose, 61 days at 10 mg/kg, and 70 days at 1 

mg/kg (values estimated from Figure 3, Lou et al., 2009). The one 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. 

The authors also tried the Andersen et al. (2006) renal-resorption model to determine if it 

provided an improved fit for the data. The Andersen et al. (2006) model fit to the data was 

superior to that of the one compartment 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. (2006) model 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 clearance, transport maximum, transport affinity constant, 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 (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. (2009) parameter estimates indicate that there may be several biological 

limitations to the Anderson 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 in order 

to optimize fit to the experimental data. It also does not include excretion via export transporters 

in the renal tubular cells and does not 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 may also be opportunities for protein binding within 

organs, including the liver, which could function to retard distribution especially at low doses 

with changes occurring when binding sites become saturated. 

 A model has also 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 (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 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. 
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Both adsorption and excretion were assumed to be first order processes as was lactation 

transfer from the dam to the litter (Figure 3-6; 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, glomerular filtration rate, 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 protein binding information. As was the case with the Lou et al. (2009) model,

Rodriguez et al. (2009) did not include parameters to adjust for transporter-mediated efflux from 

the renal tubular cells into the glomerular filtrate. 

FIGURE 3-6. Pharmacokinetic model of gestation and lactation in mice. 

 One of the limitations of the Rodriguez et al. (2009) modeling effort 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 clearance pathway for the dam.

The authors caution that the model is accompanied by a number of uncertainties because 

of the assumptions regarding the flow limitation on transport to the fetus and to maternal milk, 

the first order elimination rate constant for the pups, the milk: maternal serum partition 
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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 physiologically-based pharmacokinetic 

(PBPK) model for monkeys based on the Anderson et al (2006) and Tan et al. (2008) models, 

and then extrapolated it for use in humans (Figure 3-7). The model reflects saturable renal 

absorption of urinary PFOA by the proximal tubule of the kidney. This is represented in Figure 

3-7 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 IV dose to male Wistar rats. 

 

Tm = transporter maximum, Kt= affinity constant and Q= flow in and out of tissues 

FIGURE 3-7. Structure of the PFOA PBPK model in monkeys and humans

Existing i.v. and oral data sets from Butanhoff et al. (2004b) for the cynomolgus monkey 

were used to develop the monkey model. In the oral study (Section 3.2), animals were dosed for 

six 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 Butanhoff et al. (2004b) data for the 10 mg/kg dose showing a rapid rise to 
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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 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 or medians) for PFOA are available for occupational 

and general populations (Emmett et al., 2006; Calafat et al., 2007 a,b; Olsen et al., 2005; Hölzer 

et al., 2008; Steenland et al., 2009; Bartell et al., 2010). 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 population after reduction of the PFOA in drinking water as a result of treatment 

(Bartell et al., 2010); see section 3.5.2. In each case the major source of exposure was assumed 

to be the drinking water. 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 

compared to the measured serum values. The authors concluded that more data are needed on 

the kinetics of renal transporters, intra-human variability, and more definitive information on 

exposures in order to further refine the human model. 

Loccisano et al. (2012a) also utilized the saturable resorption hypothesis when 

developing a model for adult Sprague Dawley Rats (Figure 3-8). The structure of the model is 

similar to that for the monkey/human model depicted in Figure 3-7 but lacks the fat and skin 

compartments and includes a storage compartment to accommodate fecal loss of unabsorbed 

dietary PFOA as well as that 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. 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 utilized where appropriate. Model performance 

was evaluated primarily based on its ability to predict plasma and liver concentrations from the 

studies identified above. As was the case for the monkey model, performance was generally 

good given the limitations in the primary data sources.
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FIGURE 3-8. Structure of the PBPK Model for PFOA in the Adult Sprague Dawley Rat

 

Loccisano et al. (2012b) expanded the adult Sprague Dawley rat model described above 

to cover gestational and lactational exposure to the fetus and pups through their dams. The data 

from Hinderletter 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 Figure 3-8. 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 fats than male rats after 

sexual maturation. The lactation model linked the pups to their dams through the 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 where measured levels were greater than projection for two of three 

data points and for neonate plasma during lactation where all data points fell below the 

predictions. 

 
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3.5.2 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 and the Little Hocking Water Association in West Virginia 

(WV). Source waters for these systems had become contaminated with PFCs from the Dupont 

Works Plant in Washington, WV between 1951 and 2000. 

The Bartell et al. (2010) study was based on a series of serum measurements (8 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 

18-89 years). Most of the participants consumed public tap water (172) as their primary cource, 

but a small number (28) consumed bottled water as their source. 

 The participants were required to report primary use of home tap water for cooking, 

bathing, and showering between 2005-2007. 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 

(CDC), and 19 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/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 vs. bottled water, sex, age, consumption of local or home grown 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 home grown vegetable consumption (p<0.001). 
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 Identification of consumption of home grown 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 phrased, “Do you eat any fruits and vegetables grown at your 

own home?” Some people who answered no to the original questions andwered yes to the 

second question. 

 Changes in the source of drinking water during the study could have also 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., 2011). The PFOA levels in the finished 

drinking water were measured as 500-640 ng/L in 2006. PFOS and PFHxS were also present. 

The plasma PFOA levels in the Arnesberg 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 the PFCs in the drinking water. 

 

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 men women and children from Arnsberg and 73 from the 

reference community in their evaluation. The drinking water concentration monitoring results 

(not detects estimated as 1⁄2 the LOD, 10 ng/L) and drinking water intake estimates obtained by 

questionnaire and interview, were used to estimate PFOA exposures. Plasma PFOA samples 

were collected during a two 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 women and children than the men when 

compared to the reference community, an observation that appeared to reflect the reported lower 

drinking water intakes of the Arnsberg women 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 may have been greater in the younger subjects and older 

subjects.

Seals et al. (2011) determined half-life estimates for 602 residents of Little Hocking, OH 

and 971 residents of Lubeck, WV 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 (Little Hocking- 4 years, Lubeck-9 years) linear spline regression approach 
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in their analysis of the data. The slope of the line decreased for the second time segment as 

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 

after 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 those 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 may 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 clearance 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; Burris et al., 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. 

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 

sex of retiree, the number of years retired or worked at the production facility, or medication use 

or health conditions. 

 Harada et al. (2005a) studied the relationship between age, sex 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. (2005b) also estimated the renal clearance rate of PFOA in humans and 

found it to be only about 0.001% of the glomerular filtration rate, indicating the absence of active 
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excretion in human kidneys. There was no significant difference in renal clearance of PFOA with 

respect to sex 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 (4/sex/group) were 

administered single doses of PFOA by gavage at dose rates of 0.1, 1, 5, and 25 mg/kg PFOA. 

After dosing, plasma was collected for 22 days in males and 5 days in females. Plasma 

concentration vs. time data were then analyzed by non-compartmental pharmacokinetic methods 

(Table 3-24 and Table 3-25). 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 (2016 hours in 

males). The estimated plasma elimination half-life in this experiment was approximately 277 

hours (11.5 days) in male rats.

TABLE 3-24. Pharmacokinetic 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 

(i.v.)

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)

From Kemper, 2003

(Mean (SD))

AUCINF: area under the plasma concentration time curve, extrapolated to infininty; 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. 

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 – February 2014 3-45 

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TABLE 3-25. Pharmacokinetic 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 

(i.v.)

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)

From Kemper, 2003

(Mean (SD))

Gibson and Johnson (1979) administered a single dose of 14C-PFOA averaging 11.4 

mg/kg by gavage to groups of 3 male 10-week old CD rats. The elimination half-life of carbon14 from the plasma was 4.8 days. NRC (2005) 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. These values are 

consistent with the half-lives reported by Kemper (2003) for male and female Sprague-Dawley 

rats.

 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 3.5.1). NRC (2005) 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 or 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 were about 21 days for 

females and 30 days for males.

3.5.3 Volume of Distribution Data

Several researchers have attempted to characterize PFOA exposure and intake in humans 

(Thompson et al., 2010; Lorber and Egeghy, 2011) through pharmacokinetic modeling. As an 

integral part of model validation, the parameter for volume of distribution of PFOA within the 

body was calibrated from the available data. In the models discussed below, volume of 
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distribution was defined as the total amount of PFOA in the body divided by the blood or serum 

concentration. 

Two groups of researchers defined a volume of distribution of 170 ml/kg body weight for 

humans for use in a simple, single compartment, first-order pharmacokinetic model (Thompson 

et al., 2010; Lorber and Egeghy, 2011). 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 volume of distribution was calibrated using human serum concentration and 

exposure data from the National Health and Nutrition Examination Survey (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 value of volume of distribution was calibrated 

so that model prediction of elevated blood levels of PFOA matched those seen in residents.

Butenhoff et al. (2004b) calculated a volume of distribution from non-compartmental 

pharmacokinetic analysis of data from cynomolgus monkeys. Three males and three females 

were administered a single intravenous dose of 10 mg/kg and serum PFOA concentrations were 

measured in samples collected up to 123 days post-dosing. The volumes of distribution of PFOA 

at steady state (Vdss) were similar for both sexes at 181 ± 12 ml/kg for males and 198 ± 69 

ml/kg for females.
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4.0 HAZARD IDENTIFICATION 

4.1 Human Effects

4.1.1 Long-Term and Epidemiological Studies 

Large-scale production of PFOA occurred in the United States for several decades. The 

employees of the main production companies and the residents near the plants were exposed to 

the compound over time. There are several studies that have examined the data available on 

these occupational and residential populations. Both 3M and DuPont have been the primary U.S. 

producers and users of perfluorinated compounds, and both companies have been offering 

voluntary fluorochemical medical surveillance programs for many years to workers at plants that 

produce or use perfluorinated compounds. The monitoring data collected by 3M and DuPont 

were used in conjunction with mortality and health effect information in a number of 

epidemiological studies on their worker populations. 3M discontinued manufacturing PFOA in 

2000, but a subsidiary in Europe continued to manufacture and sell it through 2008. 

The data from NHANES have been the subject of several studies investigating the 

association between PFOA and health effects in the general population of the United States. The 

NHANES studies have examined representative members of the U.S. population (~5000 adults 

and children/year) through their surveys focusing on different health topics. The studies consist 

of an interview (demographic, socioeconomic, dietary, and medical questions) and examination 

(medical including blood and urine collection, dental, and physiological). Serum samples from 

the studies were analyzed for PFOA concentration by solid-phase extraction coupled to isotope 

dilution/high-performance liquid chromatography/ tandem mass spectrometry. 

Members of the general population living in the vicinity of the West Virginia DuPont 

Washington Works PFOA production plant in Parkersburg, WV, are the focus of an ongoing 

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 from six water districts near the plant resulting in exposures to the general 

population. The purpose of the C8 Health Project is to assess if there are any “probable links” 

between PFOA exposure and disease. During August 2005, through July 2006, about 69,000 

study participants were identified. Eligible participants included those who had consumed 

drinking water for at least one year up to and including December 4, 2004 from the Lubeck and 

Mason County water districts in West Virginia, Belpre, Little Hocking, Tuppers Plains-Chester, 

and Pomeroy water districts in Ohio, or private water source within the geographical boundaries 

of the public water sources. The participants (n=69,030; 33,242 males, 35,788 females; <10 to 

70+ years) donated a blood sample, filled out an extensive questionnaire, and received $400 in 

compensation. Serum was analyzed for perfluorochemicals, including PFOA by liquid 

chromatography separation with tandem mass spectrometry. Clinical chemistry tests were 

carried out if serum quantities were sufficient. Medical records were used to validate diseases 

reported by participants.

The results of these studies along with other population studies are described in the 

following sections. 
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4.1.1.1 Non Cancer Systemic Toxicity Studies

Several studies have been undertaken that examine the relationship between serum PFOA 

concentration and various health outcomes suggested by the standard animal toxicological 

database. Many of these monitored standard clinical chemistry parameters used within the 

occupational and public health communities as indicators of health status. In the studies of 

worker cohorts, the data collected were focused towards measures of cardiovascular risk, signs 

of organ damage, and standard haematological endpoints. Within the general population, data 

were focused on cardiovascular risk factors and diabetic or prediabetic conditions as well as 

reproductive and developmental endpoints. In reviewing the epidemiological monitoring data 

for PFOA it is important to remember that many individuals were exposed to other fluorocarbons 

and, although only serum PFOA measurements are presented in the study summaries that follow, 

many individuals had other PFCs in their serum as well. 

Serum Lipids and Cardiovascular-Related Outcomes

Occupational Population Studies

 

Serum lipids have been a focus in many of the occupational studies because data from 

animal studies shows alterations in standard profiles for total cholesterol and triglycerides 

following exposure to PFOA. A number of the studies discussed in the following paragraphs 

also examined endpoints in addition to those associated with cardiovascular disease. 

Descriptions of the outcomes related to standard clinical biochemistry endpoints and 

hematological effects are found in the sections that follow the lipid and cardiovascular-related 

observations. 

Olsen et al. (2000) analyzed data from voluntary medical surveillance examinations 

conducted in 1993, 1995, and 1997, in PFOA production plant workers to determine if serum 

PFOA concentration in workers was associated with serum lipid alterations. Cholesterol, low 

density lipoprotein (LDL), high density lipoprotein (HDL), and triglycerides were measured in 

male workers; n=111 in 1993; n= 80 in 1995, n= 74 in 1997. Multivariable regression analyses 

were adjusted for age, body mass index (BMI), alcohol use, and cigarette use. Employees’ 

serum PFOA levels were stratified into 3 categories (<1, 1- <10, and ≥10 μg/mL). Categories 

were chosen in order to maximize the number of employees in the highest exposure category. 

Serum PFOA levels ranged from 0.0-80.0, 0.0-114.1, and 0.1-81.3 μg/mL for the years 1993, 

1995, and 1997, respectively. No association was observed between serum PFOA concentration 

and total cholesterol, LDL, HDL, or triglycerides in the workers in 1993, 1995, or 1997. 

A cross-sectional analysis of the data from the 2000 medical surveillance program at the 

Decatur and Antwerp manufacturing plants was undertaken to determine if there were any 

associations between PFOA exposure and serum lipids of monitored employees (Olsen et al.,

2001a, 2003). Multivariable regression analyses were conducted to adjust for possible 

confounders including the following variables: production job (yes or no), plant, age, BMI, 

cigarettes/day, drinks/day and years worked at the plant. Mean PFOA serum levels were 1.03 

μg/mL for all male employees at the Antwerp plant and 1.90 μg/mL for all male employees at 

the Decatur plant. Male production employees at the Decatur plant had significantly higher (p 

<0.05) mean serum levels (2.34 μg/mL) than those at the Antwerp plant (1.28 μg/mL). Mean 
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PFOA serum concentrations were 0.07 μg/mL for female employees at the Antwerp plant and 

1.23 μg/mL for female employees at the Decatur plant. A positive significant association was 

reported between serum PFOA and cholesterol (p = 0.05) and triglycerides (p = 0.002). Age was 

also significant in both analyses. Alcohol consumed per day was significant in the cholesterol 

model, while BMI and cigarettes smoked per day was significant for triglycerides. 

A longitudinal analysis of the above data and previous medical surveillance results was 

carried out to determine whether occupational exposure to fluorochemicals over time was related 

to changes in serum lipids in employees of the Antwerp and Decatur facilities (Olsen et al., 

2001b, 2003). Medical surveillance data from 1995, 1997, and 2000 were analyzed using 

multivariable regression analysis. The plants were analyzed using 3 subcohorts that included 

those who participated in 2 or more medical exams between 1995 and 2000. A total of 175 male 

employees voluntarily participated in the 2000 surveillance and at least one other. Only 41 

employees were participants in all 3 surveillance periods. 

When mean serum PFOA levels were compared by surveillance year, PFOA levels in the 

employees participating in medical surveillance at the Antwerp plant increased between 1994/95 

and 1997 and then decreased slightly between 1997 and 2000. At the Decatur plant, PFOA serum 

levels decreased between 1994/95 and 1997 and then increased between 1997 and 2000. When 

analyzed using mixed model multivariable regression and combining Antwerp and Decatur 

employees, there was a statistically significant positive association between PFOA and serum 

cholesterol (p = 0.0008) and triglycerides (p = 0.0002) over time. When analyzed by plant and 

also by subcohort, these associations were limited to the Antwerp employees (p = 0.005) and, in 

particular, the 21 Antwerp employees who participated in all 3 surveillance years (p = 0.001). 

However, the association between PFOA and triglycerides was also statistically significant (p = 

0.02) for the subgroup in which employees participated in biomonitoring in 1994/95 and 2000. 

There was not a significant association between PFOA and triglycerides among Decatur workers. 

There were no significant associations between PFOA and changes over time in HDL. 

Limitations to the 2000 cross-sectional and longitudinal studies included but were not limited to 

the consistent differences in serum PFOA concentration and demographics between employees 

of the Decatur and Antwerp plants, possible confounding by other perfluorinated chemicals, and 

differences in analytical techniques. 

A cross-sectional study was undertaken to determine the relationship between serum 

PFOA and lipids in a large population of active employees (Washington Works fluoropolymer 

production plant) with potential exposure to PFOA (Sakr et al., 2007a). The employees who 

volunteered (n=1025, 782 males, 243 females) to participate in the study had a physical 

examination, provided fasting blood samples, and answered a medical and occupation history 

questionnaire in 2004. The mean age and years of work were 46.5 and 19.6 years, respectively, 

for males and 44.4 and 15.9 years, respectively, for females. Serum lipids including cholesterol, 

HDL, LDL, very low density lipoprotein (VLDL), triglyceride, uric acid (risk factor for 

hypertension) and serum PFOA concentrations were determined from blood samples. Serum 

PFOA concentration was analyzed by liquid chromatography tandem mass spectrometry. The 

relationship between PFOA exposure and lipid levels was evaluated by analysis of variance 

(ANOVA), χ2 test, t test, and linear regression models. Confounders including age, BMI, 

gender, alcohol consumption, and parental heart attack were considered in the models. 
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Serum PFOA concentration in the workers ranged from 0.0046 to 9.55 μg/mL. For those 

with current occupational exposure to PFOA, the range was 0.0174 to 9.55 μg/mL, and 0.0081 to 

2.07 μg/mL for workers with intermittent occupational exposure. The range was 0.0086 to 2.59 

μg/mL for workers with past occupational exposure and 0.0046 to 0.963 μg/mL for workers with 

no occupational exposure. Serum PFOA was positively associated with cholesterol, VLDL, and 

LDL (p<0.03) in the participating workers whether or not they were taking lipid-lowering 

medication. An association between uric acid and serum PFOA was also observed. No 

association was observed between serum PFOA and HDL or triglycerides. Limitations of the 

study included the study design (cross-sectional surveys which cannot be used to determine 

causality), and the limitations of the questionnaire which was not designed or validated for the 

study. 

Sakr et al. (2007b) conducted a longitudinal study of serum lipids in workers of the 

Washington Works plant from 1979 to 2004. Employee medical records from the medical 

surveillance program were used to obtain blood lipid (total cholesterol, LDL, HDL, triglycerides, 

and uric acid), height, and weight data. As part of the medical surveillance program, employees 

gave a detailed medical history and had a physical examination at least every 3 years. Serum 

PFOA concentration was measured every 1 to 2 years in PFOA-exposed workers and every 3 to 

5 years in non PFOA-exposed workers on a volunteer basis. This study included 454 workers 

who had 2 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 one year 

since 1979. A linear mixed effects regression model was used to analyze the data. Age, gender, 

BMI, and decade of hire were potential confounders. 

Serum PFOA concentration 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. Serum PFOA concentration was 

positively associated with total cholesterol after age, BMI, gender, and decade of hire adjustment 

in the model. Limitations of the study included inclusion of workers who had at least 2 PFOA 

measurements (only 16% of all workers in the PFOA area), lack of information on lipid-lowering 

medications and alcohol intake, and the exposure and outcome were not measured on the same 

date.

Olsen and Zobel (2007) examined the association between serum PFOA concentration 

and serum lipids in fluorochemical workers based on data collected from the 3M medical 

surveillance program in 2000. The fluorochemical workers consisted of males (age 21-67) from 

the Antwerp, Belgium (n=196), Cottage Grove, MN (n=122), and Decatur, AL (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. 

PFOA concentration was determined by liquid chromatography/tandem mass spectrometry. 

Analysis of variance, analysis of covariance, logistic regression, and multiple regression models 

were used to analyze the data with age, BMI, and alcohol consumption as covariates. 

Serum PFOA concentration ranged from 0.04 to 92.03 μg/mL for the male workers (all 

sites combined). The range was 0.01-7.04, 0.01-92.03, and 0.04-12.7 μg/mL for the Antwerp, 

Cottage Grove, and Decatur sites, respectively. The mean serum PFOA concentration was 2.21, 

1.02, 4.63, and 1.89 for all sites combined, and the Antwerp, Cottage Grove, and Decatur sites, 
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respectively. Serum PFOA (all sites combined) was not associated with total cholesterol or 

LDL. A negative association was observed between serum PFOA concentration (all sites 

combined) and HDL. However, no association was observed at the individual sites for HDL. It 

was hypothesized that the results may be due to uncontrolled residual confounding. Serum 

triglyceride was positively associated with serum PFOA at all sites combined and the Antwerp 

site. Non-adherence to fast requirement for blood collection and non-causal positive correlation 

were provided as possible hypotheses for the results. 

Costa et al. (2009) provided additional 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 to 32.5 years. In 2007, 37 men were active workers, and 

16 men were retired or had transferred to other departments and were no longer being exposed. 

Non-exposed 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. The 

medical surveillance program consisted of a yearly physical examination and blood chemistry 

tests including Apo-A and Apo-B lipoproteins, total cholesterol, HDL, triglycerides, and uric 

acid analysis. Beginning in 2000, serum PFOA was monitored yearly except in 2005. 

Determination of serum PFOA concentration was made by HPLC-electrospray-tandem mass 

spectrometry. The relationship between PFOA exposure and serum lipids was evaluated by 

ANOVA, t test, and multiple linear regression models. Confounders including age, BMI, 

smoking, alcohol consumption, and shift work were considered in the models. 

Serum PFOA concentration of the exposed workers ranged from 1.54-86.3, 0.73-91.9, 

0.34-91.9, 0.38-74.7, 0.54-46.3, 0.54-41.9, and 0.2-47.0 μg/mL for the years 2000, 2001, 2002, 

2003, 2004, 2006, and 2007, respectively. Serum PFOA concentrations in the workers decreased 

after plant renovations partially automated the PFOA production process and adopted strict 

working procedures with the use of protective devices in 2002. Total cholesterol and uric acid 

were significantly increased (p<0.05) in 34 workers currently exposed to PFOA when compared 

to 34 non-exposed workers (age and work matched). Regression analysis comparing those same 

34 exposed workers to all non-exposed workers (n=107) revealed a significant effect (p<0.05) of 

PFOA exposure on total cholesterol and uric acid levels. Serum PFOA concentration was 

correlated (p<0.05) with total cholesterol and uric acid levels in 56 workers (formally, currently, 

and never exposed) assessed concurrently from 2000-2007. No correlations were observed 

between serum PFOA concentration and Apo-A or B lipoproteins, HDL, or triglycerides in any 

of the analyses. Due to the small number of subjects, the authors were hesitant to draw firm 

conclusions, but suggested further exploration into these correlations. 

Leonard et al. (2008) examined mortality in employees at the DuPont Washington Works 

plant in West Virginia after a cross sectional survey of 1,025 active employees revealed a 

positive association between serum PFOA and cholesterol and LDL. The association led the 

authors to target ischemic heart disease (IHD) as well as many other causes of death as 

outcomes. The cohort consisted of 6,027 employees (4,872 males and 1,155 females) who had 

ever worked at the plant from 1948 through 2002. The DuPont Epidemiology Registry and US 

National Death Index were used to obtain causes of death. Standardized mortality ratios were 

estimated using three reference populations; the populations of the US and West Virginia and the 

DuPont regional worker reference population excluding workers at the Washington Works plant. 
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A non-signifcant elevation for IHD mortality (SMR= 109, 95% CI: 96, 124) and a significant 

increase in diabetes mortality [standardized mortality ratio (SMR) = 197, 95% CI: 123, 298] 

were observed for Washington Works plant workers compared to the DuPont regional worker 

reference population. No increases were observed in these two outcomes when compared to the 

national and state populations. No significant associations were observed for the other causes of 

death. Using populations spanning from 1948 to 2002 increased the possibilities for loss to 

follow-up and inadequate records. Also caution must be exercised in interpreting the results of 

mortality analyses that examine diseases that do not uniformly cause death. 

Sakr et al. (2009) further examined IHD mortality in workers at the Washington Works 

plant because of the association between PFOA exposure and increased serum lipid levels, a risk 

factor for heart disease. The cohort consisted of 4,747 employees (4,460 males and 102 females) 

who had worked in the plant from 1948 through 2002. The average duration of employment was 

22.5 years. Serum PFOA levels were linked with job titles and used to determine PFOA 

exposure for each job. Jobs were divided into categories for low, medium, or high exposures, 

and intensity factors were assigned that corresponded to worker mean serum PFOA 

concentration for all the jobs in each category. The cumulative exposure was calculated by 

multiplying the time spent in each job by the mean intensity estimate from the job exposure 

matrix. The exposure-response analysis of cumulative exposure to PFOA was used to estimate 

relative mortality risks. There were 773 total deaths in employees and 239 of the deaths were 

from IHD. Males accounted for 235 of the 239 deaths from IHD. Mortality from IHD was not 

found to be associated with PFOA exposure. These results were similar to those reported by 

Leonard et al. (2008) in which a non-significant elevation for IHD mortality was observed 

relative to expected deaths for a regional DuPont employee population. 

Lundin et al. (2009) also investigated occupational exposure to PFOA and death from 

IHD and other vascular diseases in workers at the Cottage Grove manufacturing plant as reported 

by Gilliland and Mandel (1993). Cerebrovascular disease, diabetes, and ischemic heart disease 

cases were included as the causes of deaths of 3,993 workers. A total of 807 workers died in the 

follow-up period. The cohort differed from Gilliland and Mandel (1993) in that the period of 

enrolment was extended to 1997 versus 1983, later follow-up (2002 versus 1989), and the 

inclusion criteria were modified. Standardized mortality ratios (SMRs) and 95% confidence 

intervals (CIs) were calculated. Mortality from cerebrovascular disease and ischemic heart 

disease in PFOA workers was not elevated compared to the general population of Minnesota. 

Within the cohort, the risk for cerebrovascular disease, SMR=1.6, 95% CI 0.5-3.7, was elevated 

in workers who had ever worked a job in which they were exposed to PFOA. The authors found 

diabetes mortality to be associated with moderate PFOA exposure (Hazard Ratio 3.7, 95% CI: 

1.4-10.1) among subjects whose cause of death had been identified as diabetes. No deaths from 

diabetes had occurred among workers with high exposure, based on job classification.

The studies discussed above did not uniformly monitor the same endpoints. Table 4-1 

provides a summary of the results to assist the reader in integrating the findings of the studies. 
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↑= positive association; ↓= negative association; ↔= no association; TC= total cholesterol; VLDL=very low density lipoprotein; 

LDL= low density lipoprotein; HDL=high density lipoprotein; TG= triglycerides; UA=uric acid; NM= not monitored 

General Population Studies

Several epidemiology studies examined serum lipids, lipoproteins, and/or uric acid 

among cohorts from the general population. Three studies focused on populations serviced by 

water districts contaminated by the Washington Works PFOA production plant in Ohio and West 

Virginia, and two utilized data from the NHANES program. 

Emmett et al. (2006) examined the association of serum PFOA concentration with serum 

total cholesterol 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 the Little Hocking 

water service. The subjects completed questionnaires (demographic, occupational, health 

conditions, etc) and provided blood samples. PFOA concentration was determined by high 

performance liquid chromatography tandem mass spectrometry. Regression models were used to 

analyze the data. The median serum PFOA concentration was 354 ng/mL. No association was 

observed between serum PFOA and total cholesterol. 

Steenland et al. (2009) examined the association of PFOA with serum lipids in adult 

participants of the C8 Health Project (n=46,294; 18 to ≥80 years). Serum samples were 

separated into deciles or quartiles for analysis. Total cholesterol, HDL, triglycerides, LDL, and 

non-HDL (total cholesterol 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 included gender, quartile body mass index, 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 80 ng/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 

total cholesterol, LDL cholesterol, triglycerides, total cholesterol/HDL ratio, and non-HDL. The 

predicted increase in total cholesterol from lowest to highest serum PFOA concentration decile 

was 11-12 mg/dL. No association between mean level of serum PFOA and HDL cholesterol was 

observed. 

TABLE 4-1. Association of Serum PFOA with Serum Lipids and Uric Acid in Studies of Occupational 

Populations

Study TC VLDL LDL HDL TG UA

Olsen et al., 2000 ↔ NM ↔ ↔ ↔ NM

Olsen et al., 2001a, 2003 ↑ NM NM ↔ ↑ NM

Olsen et al., 2001b, 2003 ↑ NM NM ↔ ↑ NM

Sakr et al., 2007a ↑ ↑ ↑ ↔ ↔ ↑

Sakr et al., 2007b ↑ NM ↔ ↔ ↔ NM

Olsen and Zobel, 2007 ↔ NM ↔ ↓ ↑ NM

Costa et al., 2009 ↑ NM NM ↔ ↔ ↑
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 The odds ratio 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 1.40 (95% CI: 1.29-1.51). The results of the study were consistent with occupational 

studies that found a positive association between PFOA exposure and serum lipids. Data 

interpretation was limited by the lack of cumulative PFOA exposure data for the individual 

subjects and by the cross-sectional study design which prevented determination of whether an 

increase in PFOA serum levels corresponded with an increase in cholesterol levels. 

Frisbee et al. (2010) examined the association of PFOA with serum lipids in children 

(n=6536; 1-11.9 years) and adolescent (n=5934; 12.0-17.9 years) participants of the C8 Health 

Project. Total cholesterol, HDL, LDL, and triglycerides were measured from serum samples. 

The data were analyzed by linear regression using the natural log-transformed values for all 

variables or by analysis of covariance. For some analyses, data were grouped into quintiles with 

age-group and sex-specific groupings. Covariates of the model included age, gender, body mass 

index z score, exercise, and length of fast. Triglyceride values were used only if fast had lasted 6 

hours or more. The mean serum PFOA concentration for children was 77.7 ng/mL (82.1 ng/mL 

for males; 73.1 ng/mL for females) and for adolescents, the mean was 61.8 ng/mL (69.3 ng/mL 

for males; 53.7 ng/mL for females).

Total cholesterol, LDL, and triglycerides were positively associated (p≤0.02) with serum 

PFOA concentration after covariable adjustment. Assessment of the quintile trends showed 

significant differences (p≤0.02) between the first and fifth quintile for total cholesterol and LDL 

for children and adolescents of both sexes combined and separated. A significant difference 

(p=0.04) was observed for fasting triglycerides in female children only. An increased risk of 

abnormal total cholesterol and LDL were positively associated with serum PFOA. The odds 

ratios 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 total cholesterol, 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. 

 Steenland et al. (2010) examined the association of serum PFOA concentrations with uric 

acid levels in adult subjects (n=53,458; 20 to ≥80 years) participating in the C8 Health Project 

from 2005-2006. The reference range for uric acid is 2.0 to 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, body mass index, education, smoking, alcohol consumption, and 

serum creatinine. The mean serum PFOA concentration was 86.4 ng/mL. 

 The mean uric acid level was 5.58 mg/dL with an interquartile range 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 odds ratio 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. As was the case for the Steenland et al. (2009a) report, data interpretation was limited by 

the lack of cumulative PFOA exposure data and by the cross-sectional study design. 
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 Lin et al. (2009) investigated the association between serum PFOA, HDL cholesterol, and 

triglycerides in adolescents (12-20 years) and adults (>20 years) of the United States by 

analyzing data from the 1999-2000 and 2003-2004 National Health and Nutrition Examination 

Survey (NHANES) study. The study population included 1443 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, sex, 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. Serum 

PFOA concentration was not associated with HDL or triglycerides. Data interpretation was 

limited by the cross-sectional study design which did not permit causal inference. 

 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 study. The population (n=1445; 

773 male, 672 female) included persons between the ages of 12 years and 80 years with no 

missing covariate information who were not pregnant, breast-feeding, taking insulin or 

cholesterol medicine, or undergoing dialysis. Cholesterol (total, HDL, LDL, VLDL) was 

measured from serum samples. Weight, height, and waist circumference were used in 

consideration of body size outcome. Data for covariates predicting cholesterol and body weight 

including age, sex, 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 sex and the age groups 12-19 years, 20-59 years, and 

60-80 years. 

A positive association was found between total (TC) and non-HDL (TC-HDL, ~70-80% 

TC) cholesterol and serum PFOA (9.8; 95% CI, -0.2 to 19.7). No association was found between 

serum PFOA concentration and HDL, LDL, or VLDL. No association was found between serum 

PFOA concentration and body weight. Data interpretation was limited by the cross-sectional 

study design and lack of multiple cholesterol measurements.

Table 4-2 provides a summary of the results from the general population studies using the 

same format as Table 4-1. 

TABLE 4-2. Association of Serum PFOA with Serum Lipids and Uric Acid in Studies of General 

Populations

Study TC VLDL LDL HDL Non-HDL TG UA

Emmett et al., 2006 ↔ NM NM NM NM NM NM

Steenland et al., 2009a, b ↑ NM ↑ ↔ ↑ ↑ ↑

Frisbee et al., 2010 ↑ NM ↑ ↔ NM ↔ NM

Lin et al., 2009 NM NM NM ↔ NM ↔ NM

Nelson et al., 2010 ↑ ↔ ↔ ↔ ↑ NM NM

↑= positive association; ↓= negative association; ↔= no association; TC= total cholesterol; VLDL=very low density lipoprotein; 

LDL= low density lipoprotein; non-HDL= TC(VLDL,IDL, LDL)-HDL; HDL=high density lipoprotein; TG= triglycerides; 

UA=uric acid; NM= not monitored 
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Clinical Chemistries (Biochemical and Hematological)

Occupational Population Studies

Many of the studies that investigated alterations in serum lipids also looked at alterations 

in clinical chemistry and hematological parameters. The results that apply to occupational 

cohorts are described below. 

Olsen et al. (2000; 2001a,b; 2003) examined alkaline phosphatase (ALP), gammaglutamyl transpeptidase (GGT), aspartate aminotransferase (AST), alanine transaminase (ALT), 

total and direct bilirubin, renal enzymes, blood glucose, and hematology in the workers of PFOA 

production plants as described above. No association was observed between serum PFOA 

concentration and the parameters explored in the workers in 1993, 1995, 1997, or 2000 by crosssectional or longitudinal analyses. 

 Sakr et al. (2007a) examined the relationship between serum PFOA and several clinical 

chemistry parameters. A complete blood count, metabolic panel (glucose, BUN, creatinine, iron, 

uric acid, electrolytes, creatinine kinase, lactic dehydrogenase (LDH), alkaline phosphatase 

(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 

positively associated (p<0.05) with GGT in all of the participating workers. An association 

between serum PFOA concentration and iron, LDH, calcium, and potassium was observed; 

however, the authors concluded that the associations were unlikely to have any clinical 

significance attributable to PFOA because of the small magnitude of the effect based on the 

parameter estimates, and the expectation of finding statistically significant associations by 

chance alone due to the number of comparisons performed. 

 Sakr et al. (2007b) also conducted a longitudinal study of liver enzymes in workers of the 

Washington Works plant described previouly. Hepatic clinical chemistry (GGT, AST, ALT, 

ALP, total bilirubin), height, and weight data were analyzed. Serum PFOA concentration was 

positively associated with AST (p=0.009) and negatively associated with total bilirubin 

(p=0.006) after adjustment for age, BMI, gender, and decade of hire in the model. No 

association was observed between serum PFOA concentration and GGT, ALT, and ALP. A 

limitation of the study was the lack of alcohol consumption data for the workers which could 

lead to confounding of the liver enzyme results. 

 In addition to serum lipids, Olsen and Zobel (2007) examined the association between 

serum PFOA concentration and liver enzymes in fluorochemical workers. Serum samples were 

analyzed for glucose, ALP, GGT, AST, ALT, and bilirubin concentrations. Serum PFOA was 

not associated with ALP or ALT. A negative association between total bilirubin and serum 

PFOA concentration (p<0.05) was observed at all sites combined. At the Decatur site, serum 

PFOA was positively associated with ALP, ALT, and GGT, and negatively associated with total 

bilirubin (p<0.05). The authors suggested the associations may be due to demographics and 

lifestyle rather than due to serum PFOA concentrations.

 Costa et al. (2009) also examined correlations between serum PFOA concentration and 

hematological parameters, α2 globulins, immunoglobulins (IgA, IgG, IgM), glucose, liver 
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enzymes, and kidney enzymes in workers of a PFOA production plant in Italy. Serum PFOA 

concentration was positively correlated with ALT, GGT, ALP, and α2 globulins (p<0.05), and 

negatively correlated with total bilirubin (p<0.01) in 56 workers assessed concurrently over the 

last 7 years. Body mass index was also correlated with ALT and ALP suggesting a possible 

interaction. The authors hypothesized that PFOA may interfere with metabolism, but were 

hesitant to draw firm conclusions due to the small number of subjects. No associations were 

observed between serum PFOA concentration and hematology, immunoglobulins, glucose, and 

kidney enzymes. 

 The occupational epidemiology studies that examined standard biochemical and 

hematological endpoints looked for elevated levels of one or more enzyme indicative of effects 

on the liver, but did not always examine the same enzymes. Two of the studies did not report on 

renal enzymes, blood glucose levels, or hematology. Table 4-3 summarizes the observations 

from the studies described above. 

TABLE 4-3. Associations of Serum PFOA with Serum Clinical Biochemistry and Hematology Measures

Study Liver Enzymes Bilirubin Renal Enzymes Glucose Hematology

Olsen et al., 2000, 

2001a,b; 2003 ↔ ↔ ↔ ↔ ↔

Sakr et al., 2007a ↑GGT ↔ ↔ ↔ ↔

Sakr et al., 2007b ↑AST ↓ NM NM NM

Olsen and Zobel, 2007 ↑GGT, ALP, ALT (Decatur) ↓ NM NM NM

Costa et al., 2009 ↑GGT, ALP, ALT ↓ ↔ ↔ ↔

↑= positive association; ↓= negative association; ↔= no association The entries for liver enzymes identify only 

those that were observed to have a positive association with serum PFOA. NM = Not Monitored

General Population Studies

Emmett et al. (2006) found no association between serum PFOA and liver enzymes, renal 

enzymes, and hematologic parameters in 371 residents of the Little Hocking water district in 

Ohio. 

 Lin et al. (2010) investigated the association between low-dose serum PFOA and liver 

enzymes in the adult population of the United States by analyzing data from the 1999-2000 and 

2003-2004 NHANES study. The study population included 2,216 adults (1076 men, 1140 

women) 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 PFCs, liver 

function, or the five components of metabolic syndrome. Regression models were used to 

analyze the data and adjust for confounders. 

 Serum PFOA concentration was divided into quartiles (Q1: ≤ 2.90; Q2: ≤ 4.20; Q3: ≤ 

5.95; Q4: > 5.95 ng/ml). Unadjusted liver enzymes, serum ALT and log-GGT increased across 

quartiles of PFOA (p ≤ 0.012) but total bilirubin showed no trend. The linear regression models 

were adjusted for: 

• age, gender, and race/ethnicity 

• age, gender, race/ethnicity, lifestyle (smoking status, drinking status, education level), 

• biomarker data (BMI, metabolic syndrome, iron saturation status, insulin resistance). 
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 A positive association was found between serum PFOA concentration and serum ALT and logGGT. One unit increase in serum log-PFOA was associated with an increase of 1.86 units of 

serum ALT and a 0.08 unit increase in log-GGT. An association between serum PFOA 

concentration and serum ALT was significant in the following categories: non-Hispanic 

Caucasians, lower education level, higher BMI, non-smoking, lower alcohol consumption, 

higher insulin resistance, and those with metabolic syndrome. An association between serum 

PFOA concentration and GGT was significant in the following categories: non-Hispanic 

Caucasians, higher BMI, lower alcohol consumption, and higher insulin resistance. The ALT 

level trend across serum PFOA quartiles was significant (p=0.003) in adjusted models with 

subjects having a BMI ≥ 30 kg/m2

. Data interpretation were limited by the cross-sectional study 

design, and the fact that other environmental chemicals (possible covariates or explanatory 

variables) were not included in the study, medication use was not included, and liver tissue status 

was unknown. 

Similar to the above study, Gallo et al. (2012) 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. The association of ALT, GGT, and direct bilirubin with PFOA was 

assessed using linear regression models adjusted for various confounders. The ln-transformed 

values of ALT were significantly associated with ln-PFOA and showed a steady increase in fitted 

levels of ALT per decile of PFOA, leveling off after approximately 30 ng PFOA/mL. Fitted 

values of GGT by deciles of PFOA showed a slight positive trend when adjusted for insulin 

resistence and BMI, but this was not confirmed in the logistic model analysis. No association 

was seen with direct bilirubin and PFOA levels. Limitations of the study include the crosssectional design and self-reported lifestyle characteristics. Only a small number of ALT values 

were outside the normal range making the results difficult to interpret in terms of health. 

 Only the data from Lin et al. (2010) and Gallo et al. (2012) provide any suggestion of 

effects on the liver at levels of serum PFOA observed in the general population. A significant 

association with ALT and GGT in some population groups and in individuals with a higher BMI 

was observed in the US adult population. 

Although there were no changes in kidney enzymes in the studies discussed of worker 

populations above, there is the potential for an impact of PFOA on the on the function of tubular 

resorption as a result its utilization of tubular transporters for excretions (See Section 3-4). 

Shankar et al. (2011) used data from the NHANES study to determine whether there was a 

relationship between serum PFOA levels and chronic kidney disease. A total of 4,587 adult 

participants (51.1% women) with PFOA measurements available from the 1999-2000 and 2003-

2008 cycles of the survey were examined. Chronic kidney disease was defined as glomerular 

filtration rate of less than 60 mL/minute/1.73 m2

. Serum PFOA levels were categorized into 

quartiles: quartile 1 = <2.8 ng/mL; quartile 2 = 2.8-4.1 ng/mL; quartile 3 = 4.2-5.9 ng/mL; 

quartile 4 = >5.9 ng/mL. The multivariable odds ratio for chronic kidney disease for individuals 

in quartile 4 was 1.73 (95% CI: 1.04, 2.88; p for trend = 0.015) compared with individuals in 

quartile 1. This association was shown to be independent for confounders of age, sex, 

race/ethnicity, body mass index, diabetes, hypertension, and serum cholesterol level. However, 

the authors noted that because of the cross-sectional nature of the study, the possibility of reverse 

causality could not be excluded. A low glomerular filtration rate would diminish the removal of 

PFOA from serum for excretion by the kidney, thus increasing the serum PFOA levels. Had 
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glomerular filtration rate been the variable, those in the highest PFOA quartiles could have had 

been those in the quartile with the lowest glomerular filtrations rates.

 

Diabetes and Related Endpoints

 

General Population 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; 20 to ≥80 years). Serum PFOA concentration 

was divided into deciles using the population distribution. Serum PFOA (deciles), body mass 

index, gender, family history of diabetes, race, use of cholesterol-lowering medicine, and use of 

blood pressure-lowering 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 non-fasting blood samples. The mean serum PFOA 

concentration for the entire study population was 86.8 ng/mL and 91.3 ng/mL for subjects with 

type II diabetes validated by medical review (n=3539). 

There was no association was between serum PFOA concentration and fasting serum 

glucose level in participants who were not characterized as diabetics. The mean serum PFOA 

concentration in subjects who self-reported type II diabetes (n=4278) was 92.9 ng/mL and 122.7 

ng/mL in subjects diagnosed in the last 10 years (n=1055). No association was observed 

between type II diabetes and serum PFOA concentration. The odds ratio 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 population studied. Data 

interpretation was limited by the cross-sectional study design which made it difficult to 

determine if PFOA exposure preceded disease. 

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 and glucose homeostasis and metabolic syndrome in 

adolescents (12-20 years) and adults (>20 years) as described previously. 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. 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. 

Nelson et al. (2010) examined the relationship between polyfluoroalkyl chemical serum 

concentration, including PFOA, and insulin resistance as previously described. Fasting insulin 

and fasting glucose were used to determine the homeostatic model assessment for insulin 

resistance. No association was found between serum PFOA concentration and insulin resistance.
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4.1.1.2 Thyroid Effects

Occupational Population Studies

Thyroid hormones were monitored in the medical surveillance programs at the various 

PFOA manufacturing plants and an association with serum PFOA concentration was examined 

in several occupational 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 (Olsen et al., 1998) and analyzed to determine a relationship between thyroid 

stimulating hormone (TSH) and PFOA concentration. Employees were placed into 4 exposure 

categories based on their serum PFOA levels: 0-1 ppm, 1- < 10 ppm, 10- < 30 ppm, and >30 

ppm. 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 ppm 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 ppm 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. 

Olsen et al. (2001a, 2003) and Sakr et al. (2007a) found no association between serum 

PFOA concentration and TSH, T4, and T3 levels in the Decatur, Antwerp, and Washington 

Works plant employees. Costa et al. (2009) also found no association between serum PFOA 

concentration and thyroid hormones levels in workers at a PFOA production plant in Italy. 

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 as described previously. 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) 

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.

 

General Population Studies

 

Emmett et al. (2006) examined the association of serum PFOA with thyroid disease in 

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 thyroid disease (387 ng/mL) compared to subjects 

without thyroid disease (451 ng/mL). 

Pirali et al. (2009) measured intrathyroidal levels of PFOA in thyroid surgical specimens 

to determine if a relationship existed between PFOA and the clinical, biochemical, and 
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histological phenotype of thyroid disease patients. Serum PFOA concentration was also 

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 Chi2 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 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 4-13.7, 1.2-16.6, 5.1-9.6, and 3.9-12.5 ng/ml, respectively. The concentration of 

PFOA in the thyroid and serum was similar between control and thyroid patients at the time of 

measurement. Age, sex, residence, working activity, malignant/nonmalignant conditions, 

antibodies, thyroid hormone concentrations, and ultrasound parameters were not associated with 

thyroid or serum PFOA concentration. There was also no correlation between serum and thyroid 

PFOA concentration. 

Bloom et al. (2010) investigated the associations between serum PFCs, including PFOA, 

and thyroid stimulating hormone (TSH) and free thyroxine (T4). 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 completed a questionnaire 

and provided a blood sample for serum analysis. The questionnaire contained questions about 

sport-fish and game consumption, lifestyle, demographic factors, and medical history. The 

serum samples were analyzed for TSH and free T4 in 2003 by immunometric chemiluminescent 

sandwich assay and for PFCs in 2006 by ion pair extraction high-performance chromatography 

with electrospray tandem mass spectrometry. 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 free T4 (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 1.33 ng/mL and ranged from 0.57 to 2.58 ng/mL. The males had a significantly higher 

serum PFOA concentration compared to females (1.47 ng/mL vs. 1.05 ng/mL; p = 0.047). There 

was no association between serum PFOA concentration and TSH or free 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-
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2000, 2003-2004, and 2005-2006 NHANES study. The population included 3,966 adults (2066 

women, 1900 men) older than 18 years. 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 liquid chromatography/ tandem mass spectrometry. Data on diseases diagnosed by 

a physician and confounding factors including year of NHANES study, age, gender, 

race/ethnicity, education, smoking status, BMI, and alcohol consumption were obtained from the 

questionnaire. Regression models were used to analyze the data and adjust for confounders.

Serum PFOA concentration was divided into quartiles for each sex. In women, serum 

PFOA concentration ranged from 0.1-123.0 ng/ml (Q1: 0.1-2.6; Q2: 2.7-4.0; Q3: 4.1-5.7; Q4: 

5.7-123.0), and in men, serum PFOA concentration ranged from 0.1-45.9 ng/ml (Q1: 0.1-3.6; 

Q2: 3.7-5.2; Q3: 5.3-7.2; Q4: 7.3-45.9). Women in PFOA quartile 4 were more likely to report 

current thyroid disease [odds ratio (OR) = 2.24, 95% CI: 1.38-3.65, p=0.002] compared to 

women in quartiles 1 and 2. No association between serum PFOA concentration and thyroid 

disease was observed in men. Data interpretation was limited by the cross-sectional study 

design, lack of specific thyroid disease diagnosis, and single serum samples for PFOA 

measurements taken at the same time as disease status.

Chan et al. (2011) examined the association between hypothyroxinemia and serum PFOA 

concentration in pregnant Canadian women (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 Dec 15, 2005 and June 22, 2006 as part of an elective prenatal screen for birth 

defects. Serum samples were anlayzed for TSH and free T4 concentrations and PFOA. The 

cases (n=96) had normal TSH concentrations and free T4 concentrations in the lowest 10th

percentile (≤8.8 pmol/L). The controls (n=175) had normal TSH concentrations and free T4 

concentrations between the 50th and 90th percentiles (12-14.1 pmol/L). Maternal age, weight, 

and gestational age at blood draw, dichotomized at 50th percentiles were included as 

confounders, and race was included at a covariate. Chi-square tests and regression models were 

used to analyze the data. The geometric mean serum PFOA concentration in the cases was 3.10 

nmol/L and 3.32 nmol/L in the controls. There was no association between serum PFOA 

concentration and hypothyroxinemia in pregnant women. The authors noted that hormone levels 

change over pregnancy and time of blood could have influenced the results. They also noted that 

the population may have been skewed because only women who chose to be screened were 

included. 

 

There are no strong data linking serum PFOA to changes in thyroid hormones or thyroid 

disorder and indicated by the Table 4-4 compilation of the studies in the occulational and general 

population. 
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TABLE 4-4. Association of serum PFOA with the prevalence of thyroid disease and thyroid hormone levels 

in studies of general and worker populations

Study Population Thyroid Disease TSH T3 T4

Bloom et al. (2010) General ↔ ↔ NM ↔

Chan et al. (2011) General ↔ ↔ NM ↔

Costa et al. (2009) Occupational NM ↔ ↔ ↔

Emmett et al. (2006) General ↔ NM NM NM

Melzer et al. (2010) General ↔ men

↑ women

NM NM NM

Olsen et al. (1998) Occupational NM ↑ NM NM

Olsen et al. (2001a, 2003) Occupational NM ↔ ↔ ↔

Olsen and Zobel (2007) Occupational NM ↔ ↑ ↔ serum

↓ free

Pirali et al. (2009) General ↔ NM NM NM

Sakr et al. (2007a) Occupational NM ↔ ↔ ↔

↑= positive association; ↓= negative association; ↔= no association; NM = Not Monitored 

4.1.1.3 Steroid Hormones

Occupational Population Studies

Olsen et al. (1998) examined several hormones including cortisol, estradiol, follicle 

stimulating hormone, dehydroepiandrosterone sulfate, 17 gamma-hydroxyprogesterone (a 

testosterone precursor), free testosterone, total testosterone, luteinizing hormone, prolactin, and 

sex hormone-binding globulin in male workers at the Cottage Grove, Minnesota, production 

plant for 1993 and 1995. No association between serum PFOA and any hormone was observed, 

but some interesting trends were noticed. When the mean measures of the various hormones 

were compared by exposure categories, there was a statistically significant (p = 0.01) elevation in 

prolactin in 1993 only for the 10 workers whose serum PFOA levels were between 10 and 30 

ppm compared to the lower 2 exposure categories. 

Estradiol levels in the >30 ppm PFOA group in both years were 10% higher than 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 postulate that the study may 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 (55 ppm PFOA in the CD 

rat). Only 3 employees in this study had PFOA serum levels this high. They also 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. (2007a) study, an association was observed between serum PFOA and 

serum estradiol (p=0.017) and testosterone (p=0.034) in male workers of the Washington Works 

plant; however, circadian variations of hormones were not taken into consideration during 

analysis. The biological signifificance of the results is unknown.

Costa et al. (2009) found no association between serum PFOA concentration and 

estradiol or testosterone in workers at a PFOA production plant from 2000 to 2007. 
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General Population Studies

Knox et al. (2011) examined the endocrine disrupting effects of perfluorocarbons in 

women from the C8 Health by analyzing the relationship between serum PFOA, serum estradiol

concentration, and menopause onset. The population included women over age 18 years 

(n=25,957). Serum PFOA and estradiol concentrations were determined from blood samples. 

Women who were pregnant, had full hysterectomies, and taking any prescription hormones or 

selective estrogen receptor modulators and fertility agents were excluded from estradiol analysis. 

Serum PFOA concentration was grouped into quintiles (natural log transformation): 1=0.25-

11.2; 2=11.3-19.8; 3=19.9-36.7; 4=36.8-84.9; and 5=85.0-22412.0 ng/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 hysterectomies. Logistic 

regression models were adjusted for smoking, age, BMI, alcohol consumption, and regular 

exercise. PFOA concentration in women who had a hysterectomy was significantly higher than 

women who had not had a hysterectomy. Serum PFOA and estradiol concentrations were not 

associated. Menopause analysis in the oldest group of women showed that all quintiles were 

significantly higher than the lowest, and in women between the ages of 42 and 51 years, quintiles 

3-5 were significantly higher than the lowest. The authors suggested that PFOA may be 

associated with endocrine disruption in women. Data interpretation was limited by the crosssectional study design and survey-reported menopause without age or independent confirmation. 

7.1.1.4. Immunotoxicity

As described previously, no association between serum PFOA concentrations and 

immunoglobin levels was found in male workers (Costa et al., 2009). 

Fei et al. (2010b) examined the association between prenatal PFOA exposure and 

hospitalizations for infectious diseases in early childhood. A total of 577 hospitalizations 

occurred with 363 being due to infectious disease. A lower risk of hospitalization due to 

infections was associated with higher maternal PFOA concentration. The incidence rate ratio 

(IRR) was significantly different only for the second quartile of exposure (IRR=0.71, 95% CI 

0.53-0.94) compared to the first. A slightly higher risk for hospitalizations was observed in girls 

with higher maternal PFOA concentrations (IRR= 1.00, 1.20, 1.63, 1.74 for first, second, third, 

and fourth quartiles, respectively). The risk for boys was below 1.0 for each quartile. Overall, 

there was no association between hospitalizations due to infectious diseases and maternal PFOA 

exposure. 

Okada et al. (2012) investigated the relationship between maternal PFOA concentration 

and infant allergies and infectious diseases during the first 18 months of life as well as cord 

blood IgE levels. The prospective birth cohort was based on infants delivered at the Sapporo 

Toho Hospital in Sapporo, Hokkaido, Japan between July 2002 and October 2005. PFOA levels 

were measured in maternal serum taken after the second trimester (n=343) and total IgE 

concentration was measured in cord blood (n=231) at the time of delivery. Infant allergies and 

infectious diseases were assessed in a maternal self-administered questionnaire at 18 months 

post-delivery. Polynomial regression analyses, adjusted for potential confounders, were 

performed on log-transformed data. Mean maternal PFOA concentration was 1.4 ng/mL and 
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cord blood IgE level was 0.62 IU/mL. Cord blood 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 incidence of food allergy, eczema, 

wheezing, or otitis media in infants at 18 months of age. Limitations of the study include the 

small sample size, potential selection bias of the population, and accurate diagnosis of disease in 

the infants. 

Antibody responses to diphtheria and tetanus toxoids following childhood vaccinations 

were assessed in context of exposure to 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; postnatal exposure was assessed from serum 

collected from the child at 5 years of age. Multiple regression analyses with covariate 

adjustments were used to estimate the percent difference in specific antibody concentrations per 

2-fold 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 2-fold increase in 

PFOA was associated with differences of -36% (95% CI, -52% to -14%) and -25% (95% CI, -

43% to -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 IU/mL. The 

odds ratios of antibody concentrations falling below this level were 4.20 (95% CI, 1.54 to 11.44) 

for tetanus and 3.27 (95% CI, 1.43 to 7.51) for diphtheria when age 7 antibody levels were 

correlated with age 5 PFOA serum concentrations.

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 non-asthmatic 

controls. Between 2009 and 2010, asthmatic children were recruited from two hospitals in 

Northern Tiawan, while the controls were part of a cohort population in seven public schools in 

Northern Tiawan. Serum was collected for measurement of ten 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 asthema severity 

score (including frequency of attacks, use of medicine, and hospitilazation) during the previous 

four weeks. Associations of perflourinated compound quartiles with concentrations of 

immunological markers and asthma outcomes were estimated using multivariable regression 

models. Nine of ten perfluorinated compounds were detectable in ≥84.4% of all children with 

levels generally higher in asthmatic children compared with non-asthmatics. Serum 

concentrations of PFOA in asthmatic and non-asthmatic children were 1.5±1.3 and 1.0±1.1 

ng/mL, respectively; four other compounds were measured at higher concentrations with the 

highest levels for PFOS and perfluorotetradecanoic acid. The adjusted odds ratios for asthma 

association with the highest versus lowest quartile levels were significantly elevated for seven of 

the compounds. For PFOA the odds ratio was 4.05 (95%CI: 2.21, 7.42). In asthmatic children, 

absolute eosinophil counts, total IgE, and eosinophilic cationic protein concentration were 

postitively associated with PFOA levels with a significant monotonic trend with increasing 
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serum concentration. None of these biomarkers was significantly associated with PFOA levels 

in non-asthematic children. Serum PFOA levels were not significantly associated with asthma 

serverity scores among the children with asthma. 

4.1.1.4 Reproductive & Developmental Endpoints

Several studies have examined the relationship between PFOA exposures and 

reproductive, gestational, and developmental endpoints as well as postnatal growth and 

maturation. The reproductive endpoints included sperm count (Joensen et al., 2009) and 

fecundity (Fei et al., 2009). The gestational endpoints included gestational age (Nolan et al., 

2009), and fetal growth (Fei et al., 2007, 2008a; Washino et al., 2009). The developmental 

endpoints included birth weight (Apelberg et al., 2007; Monroy et al., 2008; Nolan et al., 2009; 

Stein, et al., 2009), miscarriage or pre-term birth (Stein, et al., 2009), birth defects (Stein et al., 

2009), and infant development during the first two years (Anderson et al., 2010; Fei et al., 2008b, 

2010a,b). Postnatal growth and maturation included behavioural assessment (Fei and Olsen, 

2011; Hoffman et al., 2010), onset of puberty (Lopez-Espinosa et al., 2011; Christensen et al., 

2011), and risk of adult obesity (Halldorsson et al., 2012). As a group, the studies do not suggest 

an effect of serum PFOA at the levels examined on reproductive and developmental outcomes. 

A few studies indicate an impact on anthropometric measures (e.g. body weight, birth length) but 

the associations are not strong. 

Andersen et al. (2010) and Fei et al. (2007, 2008a, b, 2009, 2010a,b) conducted a series 

of studies examining the association of maternal plasma PFOA concentration and various 

reproductive and developmental outcomes in Danish women and their offspring. The women 

(n=1400) and their infants were randomly selected from the Danish National Birth Cohort, 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 women 

participated in telephone interviews (at 12 and 30 weeks gestation and when children were 6 and 

18 months 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 sex, 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 1399 women, from the second blood samples of 200 women, and from cord blood 

samples of 50 infants by solid phase extraction high performance liquid chromatography-tandem 

mass spectrometry. PFOA concentrations were divided into quartiles (Fei et al., 2009, 2010a), 

with the lowest quartile designated as the reference group, as follows: <lower limit of 

quantification (LLOQ)-3.90, 3.91-5.20, 5.21-6.96, and ≥ 6.97 ng/mL. Regression models were 

used to analyze the data. Results of these studies are included in the discussion of results for 

specific endpoints below.
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Reproductive Outcome 

Fei et al. (2009) examined the association between plasma PFOA concentration and 

longer time to pregnancy as a measure of fecundity in 1240 women. The time to pregnancy 

(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 women had a TTP within the first 2 months of trying to get 

pregnant and 379 had a TTP of ≥6 months with 188 of those women having a TTP of >12 

months. The mean plasma PFOA concentration was 5.6 ng/mL for women who planned their 

pregnancies, and for TTP <6 months, 6-12 months, and >12 months was 5.4, 6.0, and 6.3 ng/mL, 

respectively. Plasma PFOA concentration was significantly greater (p<0.001) in women who 

had TTP >6 months compared to those whose TTP was <6 months. The women with a TTP >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 women with >3.91 ng/mL plasma PFOA concentration compared to 

women with <3.91 ng/mL plasma concentration. The fecundity odds ratio 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). Although the results of the study suggest that plasma PFOA 

concentration may reduce fecundity, the authors noted that selection bias, the unknown quality of 

the sperm, unknown frequency and timing of intercourse, and abnormal hormone levels may 

have an impact on the results and fecundity. 

Nolan et al. (2009, 2010) published two studies that examined birth outcome in infants 

born to mothers whose partial or full public water source was from the Little Hocking Water 

Association (LHWA) in Washington County, OH, or from the same geographic area but with 

water not provided by the LHWA. The population serviced by LHWA had serum levels of 

PFOA (6.78 μg/L) that were approximately 50 times higher than serum levels of the general 

population. Archival data for births, including birth weight, gestation age, plurality, neonatal 

sex, race, mother’s age, and mother’s zip code, in Washington County, OH, from January 1, 

2003 to September 1, 2005 were obtained for analysis. For mothers with public water service

provided completely, partially, or not at all by LHWA, the incidence of preterm births was 

10.7%, 11.3%, and 13.4%, respectively (Nolan et al., 2009). The national incidence of preterm 

birth is 12.7%. No differences were found between gestational age of infants born to mothers 

whose partial or full water source was serviced by LHWA and infants born to mothers whose 

water source was serviced by another provider. Exposure to high levels of PFOA through the 

drinking water was not associated with preterm births. 

 

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 

reported cases were low in number. 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. An increased likelihood (crude OR 11, 95% CI: 1.8-64) of anemia and dysfunctional 
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labor (crude OR 5.3, 95% CI: 1.2-24) in mothers with water provided by LHWA was found, but 

the reported cases were low in number. No association was found between PFOA and increased 

incidence of congenital anomalies, most labor and delivery complications, and maternal risk 

factors.

Stein et al. (2009) examined the association between PFOA serum levels and pregnancy 

outcome in a population of women living in the Mid-Ohio Valley near Parksburg, WV from 

2000 to 2006. The women were part of the C8 Health Project. Pregnancies were restricted to 

those occurring in the 5 years prior to C8 Health Project enrollment (August 2005 through July 

2006) to ensure that the serum PFOA levels measured during enrollment were the same levels 

that occurred during pregnancy. Restrictions on residency including living in the same water 

district from pregnancy to enrollment were also applied. Preeclampsia and birth defects were 

included in the analysis as well. Maternal age, education, and smoking status were possible 

cofounders. A total of 1,845 pregnancies for 1,505 women met the requirements for the study. 

Regression models were used for statistical analysis. The median PFOA serum level was 21.2 

ng/mL, and the mean serum PFOA level was 48.8 ng/mL. 

No association was observed between serum PFOA levels and miscarriage, low weight 

birth, or pre-term birth. Serum PFOA levels above the median were weakly associated with 

preeclampsia. A weak association (odds ratio 1.7, CI: 0.8-3.6) was observed between PFOA 

levels above the 90th percentile and birth defects. Data interpretations were limited by the 

quality of the self-reported pregnancy outcome especially in preeclampsia, miscarriage, and 

preterm births.

Joenson et al. (2009) examined the association between PFCs, including PFOA, and 

testicular function in 105 Danish men who provided semen and blood samples as part of 

reporting for the military draft in 2003. The men chosen had the highest (ranging from 30.1 to 

34.8 nmol/L; n=53; 18.2-24.6 years) and lowest (ranging 10.5-15.5 nmol/L; n=52; 18.2-25.2 

years) testosterone concentrations. Regression models were used to analyze associations 

between PFCs and testicular function. Median serum PFOA concentration was 4.4, 5.0, and 4.9 

ng/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. 

Anthropometric Endpoints

Apelberg et al. (2007) measured PFOA in the cord blood of 293 newborns (singleton

births) born November 26, 2004 through March 16, 2005 in Baltimore, MD, at Johns Hopkins 

Hospital. Maternal and infant data including maternal birth cohort, social class, place of 

residence, past pregnancies, insurance type, BMI, age, race, education, marital status, parity, 

gestational age, smoking status, and infant sex were collected from the hospital database and 

forms filled out at time of delivery. All samples of cord blood contained PFOA (geometric mean 

1.6, range 0.3-7.1 ng/mL). Higher concentrations (p<0.03) of PFOA were measured in infants 

born to obese mothers. Concentration was evenly distributed by race, but not by sex. Male 
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infants had significantly lower (p<0.01) PFOA concentrations than females. Lower 

concentrations (p<0.02) were also measured in instances of multiparous birth. No correlations 

were observed between PFOA and maternal age, gestational age, education, insurance type, 

marital status, smoking status, and place of residence. Birth weight, head circumference, and 

ponderal index were negatively associated with PFOA.

At birth, Fei et al. (2007, 2008a) examined the association between plasma PFOA 

concentration and infant birth weight, gestation age, placental weight, birth length (skeletal 

growth), head circumference (brain), and abdominal circumference (liver size). The mean 

plasma PFOA concentration from the first blood sample, the second blood sample, and the cord 

blood sample was 5.6, 4.5, and 3.7 ng/mL, respectively. Plasma PFOA concentration ranged 

from below the limit of quantitation (≤1.0 ng/ml) to 41.5 ng/ml. Mean infant birth weight was 

3623 g; 24 infants were categorized as low birth weight infants (<2500 g). The mean gestational 

age was 280.4 days, and 53 were preterm births and 139 were post-term births. Plasma PFOA 

concentration was significantly inversely associated with birth weight among normal-weight 

women (adjusted β=-10.63 g; 95% CI, -20.79 to -0.47 g). A statistically significant decrease was 

observed in birth length in the second and fourth quartiles of PFOA exposures compared to the 

first quartile after adjustment for confounders. When maternal PFOA concentration was 

considered as a continuous variable in relation to fetal growth indicators, a statistically 

significant negative association between abdominal circumference and birth length were 

observed after adjustment for confounders. The authors noted that any associations may not be 

causal, but rather due to chance or confounding factors. 

Monroy et al. (2008) examined the association between maternal serum and umbilical 

cord PFCs, including PFOA, and birth weight. The population (n=101 pregnant mothers) was 

selected from those recruited from January 2004 to June 2005, to participate in the Family Study 

in Canada. As part of the study, the women provided blood samples during the second trimester 

of pregnancy and filled out an obstetrical history questionnaire. Another blood sample was 

provided at delivery as well as blood from the umbilical cord. Age, occupation, pre-pregnancy 

height and weight, education, and infant gender and birth weight were obtained post delivery. 

Paired t-test and linear regression analysis were used to analyze the data. The mean serum 

PFOA concentration was 2.54, 2.24, and 1.94 ng/mL at 24-28 weeks, delivery, and in the 

umbilical cord, respectively. There was no association between serum or umbilical cord PFOA 

concentration and birth weight. 

As mentioned above, Nolan et al. (2009, 2010) published two studies that examined birth 

outcome in infants born to mothers whose partial or full public water source was from the Little 

Hocking Water Association (LHWA) in Washington County, OH, or from the same geographic 

area but with water not provided by the LHWA. The population serviced by LHWA had serum 

levels of PFOA (6.78 μg/L) that were approximately 50 times higher than serum levels of the 

general population. Archival data for births, including birth weight, gestation age, plurality, 

neonatal sex, race, mother’s age, and mother’s zip code, in Washington County, OH, from 

January 1, 2003 to September 1, 2005 were obtained for analysis. Analysis of variance and 

regression models or logistic regression models were used for statistical analysis. The noted 

limitations included lack of individual exposure levels, possible misclassification of mothers by 

zip code, misclassification of socioeconomic status, possible reporting error in the database used 

to abstract the data, and using data from only live births for analysis. 
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Nolan et al. (2009) examined the association between birth weight with PFOA in a total 

of 1,555 singleton neonates (777 male, 778 female): 11% were born to mothers fully serviced by 

LHWA, 13% born to mothers partially serviced by LHWA, and 76% born to mothers not 

serviced by LHWA. For mothers with public water service provided completely, partially, or not 

at all by LHWA, the incidence of low birth weight was 3.6%, 3.8%, and 8.1%, respectively. The 

national incidence of low birth weight is 8.2%. No differences were found between birth 

weights of infants born to mothers whose partial or full water source was serviced by LHWA and 

infants born to mothers whose water source was serviced by another provider. 

 

Washino et al. (2009) investigated the correlation between prenatal exposure to 

perfluorinated chemicals and reduced fetal growth in a population of women (n=428, average 

age 30.5 years) living in or near Sapporo, Japan. Blood samples were obtained from the women 

either after the second trimester or after delivery for PFOA and PFOS concentration 

determination. PFOA was detected in 92.8% of the blood samples, and PFOS was detected in all 

samples. Maternal serum PFOA concentrations ranged from <0.5 (limit of detection) to 5.2 

ng/mL. A regression model adjusted only for gestational age showed a correlation between 

log10-transformed PFOA concentrations and decreased birth weight. The same model showed a 

correlation between log10-transformed PFOA concentrations and decreased birth weight in male 

infants. However, when the regression model was adjusted for multiple factors, including 

maternal age, education level, smoking status, BMI, parity, how infant was delivered, infant sex, 

gestational age, and when blood sample was taken, no correlation was found. No correlations 

were observed between PFOA and birth length, chest circumference, or head circumference. 

Data interpretation is limited by small sample size, selection bias, and possible errors in birth 

length, chest circumference, and head circumference measured due to molding during birth. 

Hamm et al. (2009) investigated whether increasing maternal exposure to PFCs, 

including PFOA, was associated with fetal growth and gestation length. Pregnant women living 

in and around the city of Edmonton, Alberta, Canada, (n=252; age ≥18 years) submitted serum 

samples at 15-16 weeks gestation as part of a screening for birth defects between December 15, 

2005 and June 22, 2006. The serum samples were included in the study if the mother had given 

birth to a live singleton with no evidence of malformations at greater than or equal to 22 weeks 

of gestation. Maternal serum PFOA concentrations ranged from below the limit of detection to 

18 ng/mL and had a mean of 2.1 ng/mL. A total of 21 births (8.3%) were classified as preterm 

deliveries and 16 infants (6.3%) were considered small for gestational age. The mean infant 

birth weight was 3349 g and the mean birth weight z-score was 0.062. The natural log of 

maternal serum PFOA concentration was associated with changes in birth weight of -37.4 g 

(95% CI: -86.0-11.2) after adjustment for covariates. No association was observed between the 

untransformed serum PFOA concentration and birth weight. There was no association between 

maternal serum PFOA concentration and birth weight z-score, gestation length, or preterm 

delivery. Data interpretation may have been limited by the small sample size, selection bias, and 

possible errors in birth length, chest circumference, and head circumference measurements. 

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=1010) reported the information during an interview and weight and length measurements 

were used to calculate BMI. Maternal plasma PFOA concentration was inversely associated 
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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 

boys (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. No associations were observed between maternal plasma PFOA concentration 

and the endpoints for female children in the adjusted models. 

Neurodevelopmental Endpoints

 

Fei et al. (2008b) examined the association between plasma PFOA concentration in 

pregnant women 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 women having higher plasma 

PFOA concentrations reached developmental milestones at the same times as children born to 

women 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. 

Fei and Olsen (2011) examined the association between prenatal PFOA exposure and 

behavior or coordination problems in the children at age 7. 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 5.7 ng/mL, and PFOA levels were divided into quartiles: <LLOQ-3.95, 3.96-5.32, 5.35-

7.11, and 7.14-21.90 ng/mL. A child having higher scores in total difficulties, emotional 

symptoms, and hyperactivity was negatively associated with the second (OR=0.56, 95% CI 0.27-

1.19; p<0.05) or third (OR=0.36, 95% CI 0.15-0.82; p<0.05) PFOA quartiles when compared 

with women in the lowest quartile. Odds ratios adjusted for parity, maternal age, pre-pregnancy 

BMI, pregnancy smoking and alcohol consumption, socio-occupational status, child gender, 

breastfeeding, birth year, home density, gestational age at blood draw, and 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 exposure was found. 

Hoffman et al. (2010) examined the associations between perfluorochemicals, including 

PFOA, and diagnosis of attention deficit hyperactivity disorder (ADHD) using the NHANES 

data from 1999-2000 and 2003-2004. The study population included 571 children aged 12-15 

years including those who had been diagnosed as having ADHD (n=48) and/or taking ADHD 

medications (n=21). Age, gender, and race/ethnicity were included as covariates, and 

socioecomonic status, health insurance coverage and having a routine health care provider, living 

with someone who smokes, birth weight, admittance to NICU, maternal smoking, and preschool 

attendance were confounders. Regression models were used to analyze the data. The median 

serum PFOA levels were 4.4 μg/L and ranged from 0.4 to 21.7 μg/L. Serum PFOA was 

positively associated with parental report of ADHD (OR=1.12, 95% CI 1.01-1.23). The odds 
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ratio for serum PFOA and parental report of ADHD and ADHD medication use was 1.19 (95% 

CI 0.95-1.49). Data interpretation were 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 that were not included in the available 

data. 

Postnatal Development 

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 

(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=3076) and females (n=2931) 

aged 8-18 years. Serum PFOA concentrations were divided into quartiles: <11.4, 11.4-23, >23-

58, and >58 ng/mL. Confounders including age at survey, BMI, BMI z-score, height, family 

income, ethnicity, smoking status, alchohol consumption, date and time of sample collection 

were included in the logistic regression models used to analyze the data. The median PFOA 

concentrations were 26 and 20 ng/mL in boys and girls, respectively. No association between 

PFOA concentration and puberty was observed for boys. Reduced odds of having reached 

puberty was associated with higher PFOA exposure in girls (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) was also observed (OR 0.83, 95% CI 

0.74-0.93). The results suggested that PFOA was associated with a later age of menarche. The 

authors expressed caution in interpretion of the data because of lack of serum PFOA 

concentration prior to puberty, PFOA concentration was measured after the attainment of 

puberty, and lack of secondary sexual maturation data (physical, Tanner criteria, and biomarker 

measurements).

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. The study 

population from the Avon Longitudinal Study of Parents and Children included single-birth 

female subjects who had completed at least 2 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 were 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 including a those who experienced menarched prior to age 11.5 years 

(n=218) and a random sample of those who experienced menarche after age 11.5 (n=230). 

Confounders including the mother’s pre-pregnancy 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 3.9 and 3.6 ng/mL for the early menarche and non-early menarche groups, 

respectively. There was no association between age at menarche and prenatal PFOA exposure. 

The authors noted that a modest non-significant association between the odds of earlier 

menarche and prenatal serum PFOA concentrations above the median. For all models, the 

confidence intervals included the null value of 1.0. The limitations of the study included having 

a small sample size, using a single maternal gestational serum sample for perfluorinated 

chemical measurement, and missing information from the controls and cases. 
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Halldorsson et al. (2012) examined prenatal exposure to PFOA and the risk of being 

overweight at 20 years of age in a prospective study. A birth cohort consisting of 665 motheroffspring 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 3.7±2.0 ng/mL with quartiles of 2.4±0.6, 3.3±0.4, 4.2±0.5, and 5.8±1.9 

ng/mL. In covariate adjusted analyses, female offspring whose mothers were in the highest 

quartile had 1.6 kg/m2

 (95% CI: 0.6, 2.6) higher BMI and 4.3 cm (95% CI: 1.4, 7.3) higher waist 

circumference compared to 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 [relative risk (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 

sample 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 overweight or obesity in male offspring. 

4.1.2 Cancer 

Occupational Population Studies

Several occupational studies examining cancer mortality have been conducted at 3M’s 

Cottage Grove facility in Minnesota. These studies have reported associations with prostate 

cancer mortality and cerebrovascular disease mortality (Alexander 2001a,b; Gilliland and 

Mandel, 1993). 

Lundin et al. (2009, previously described) further examined occupational exposure to 

PFOA and death from cancer in Cottage Grove employees explored by Gilliland and Mandel

(1993) using a longer period of enrolment and later follow-up. Cancers outcomes including 

liver, bladder, pancreas, testes, and prostate were examined in the 3,993 workers. Mortality from 

liver, pancreatic, prostate, and testicular cancer in PFOA workers was not elevated compared to 

the general population of Minnesota. Within the cohort, the SMR for prostate cancer, SMR=2.1, 

95% CI 0.4-6.1, was elevated in workers who had ever worked a job in which they were exposed 

to PFOA. 

Researchers suggested that pancreas acinar cell adenomas seen in rats exposed to PFOA 

may be the result of increased cholecystokinin (CCK) levels secondary to blocked bile flow 

(Obourn et al., 1997). 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. 

As a result, 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 3 categories (<1, 1- <10, and ≥10 ppm). The mean CCK 

values for the 3 PFOA categories were 33.4, 28.0, and 17.4 pg/mL, respectively. The means in 

the 2 serum categories < 10 ppm were at least 50% higher than in the ≥ 10 ppm category. A 

statistically significant (p = 0.03) negative association between mean CCK levels and the three 

PFOA categories was observed. 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 tests, 
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hypolipidemia, or cholestasis were observed in the workers. The authors suggested that the lack 

of a positive association between PFOA and CCK in workers may 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.

Limited data are available on mortality and cancer incidence at DuPont’s Washington 

Works Plant in Parkersburg, WV. Several studies of this population have been published 

(Walrath and Burke, 1989; Karns and Fayerweather, 1991; DuPont, 2003). In the most recent 

report (DuPont, 2003) cancer incidence for active employees was reported for the 1959-2001 

time period and mortality data were reported for active and retired employees for 1957 through 

2000 (DuPont, 2003). No information on employee exposure, lifestyle factors, employee 

demographics, or information on other chemicals used at the plant, is presented in the DuPont 

(2003) report.

Cancer cases were identified through a combination of company health and life insurance 

claims and company cancer and mortality registries. Standardized incidence ratios (SIRs) were 

only calculated for those cancers for which 5 or more cases were observed, which included 14 

types of cancer. Two of those cancer types were elevated and statistically significant (p = 0.05): 

bladder [standardized incidence ratio (SIR) = 1.9; 95% CI (1.15-3.07)] and kidney and urinary 

organs [SIR = 2.3 (95% CI = 1.36-3.65)]. All of the reported cases were male. Some other cancer 

types with elevated SIRs (not statistically significant at p = 0.05) included myeloid leukemia 

(2.02), cancer of the larynx (1.77), multiple myeloma and immunoproliferative (1.72), malignant 

melanoma of skin (1.3), testicular cancer (1.46), and brain cancer (1.2). 

Two separate analyses of leukemia incidence were conducted prior to the DuPont (2003) 

effort (Walrath and Burke, 1989; Karns and Fayerweather, 1991). Walrath and Burke (1989) 

reported a statistically significant (p = 0.10, significance level chosen by researchers) elevated 

odds ratio (OR) of 2.1 for leukemia incidence for male employees working at the plant from 

1956-1989. Eight cases (all male) of different types of leukemia were identified. The OR 

remained elevated when the workers were divided into wage and salaried employees (2.2 and 

2.0, respectively). In the follow-up case-control study by Karns and Fayerweather (1991), four 

controls were selected from the plant for each case, matched on gender, age and payroll status. 

Matched odds ratios were significantly elevated (p = 0.10) for employees who had previously 

worked as custodians and engineers, 8.0 (90% CI, 1.1-60.0) and 7.9 (90% CI, 1.0 - 76.0), 

respectively and remained elevated (although not statistically significant) for these same job 

categories within the plant (OR= 4.0 and 5.1, respectively). Matched ORs were also reported 

based on the area of the plant where the cases worked; however, no statistically significant 

association (p = 0.10) was identified. 

The data reported above do not show any statistically significant (p = 0.10) elevations in 

leukemia deaths, possibly because the number of cases was very small and divided among 

different types of leukemia. The Washington Works data provide some insight as to where more 

medical surveillance should be concentrated at this plant but provide little information about the 

relationship of PFOA to mortality or cancer incidence since no exposure information, use of 

other chemicals, or lifestyle information was collected on these employees.
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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 (50-65 years of age) 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=1240) consisted of 1,111 men and 

129 women whose median age was 59 years having prostate cancer (n=713), bladder cancer 

(n=332), pancreatic cancer (n=128), and liver cancer (n=67). The individuals (n=772) in the 

subcohort comparison group were randomly chosen from the cohort study and consisted of 680 

men and 92 women whose median age was 56 years. The participants answered a questionnaire 

upon enrollment in the cohort study, and data on known confounders were obtained from the 

questionnaires. Study participants provided plasma samples which were analyzed for PFOA 

concentration by high pressure liquid chromatography coupled to tandem mass spectrometry. 

 

The plasma PFOA concentrations for cancer patients were as follows: men 6.8 ng/ml, 

women 6.0 ng/ml, prostate cancer 6.9 ng/ml, bladder cancer 6.5 ng/ml, pancreatic cancer 6.7 

ng/ml, and liver cancer 5.4 ng/ml. The plasma PFOA concentrations for the subcohort 

comparison group were 6.9, 5.4, and 6.6 ng/mL for the men, women, and combined, 

respectively. Incidence rate ratios, crude and adjusted for confounders, did not indicate an 

association between plasma PFOA concentration and prostate, bladder, pancreatic, or liver 

cancer. 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. 

More recently, Vieira et al. (2013) investigated the relationship between PFOA exposure 

and cancer among the residents living near the DuPont plant in Parkersburg, WV. This analysis 

included incident cases of 18 cancers diagnosed from 1996-2005 in five Ohio counties and eight 

West Virginia counties which 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 from data collected previously from participants in the C8 Health Project and linked to 

the residence or water district of the cancer cases. Indivudual-level exposure was categorized as 

very high, high, medium, low, or unexposed based on serum concentrations of >110 μg/L, 30.8-

109 μg/L, 12.9-30.7 μg/L, 3.7-12.9 μg/L, and unexposed (background levels not given), 

respectively. Logistic regression was applied to individual-level data to calculate odds ratios 

(OR) and confidence intervals (CI) for each cancer category.

Data were first analyzed by water district. The adjusted odds ratios were increased for 

testicular cancer (OR: 5.1, 95% CI: 1.6, 15.6; n=8), and for kidney cancer (OR: 1.7, 95% CI: 0.4, 

3.3; n=10) in 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. Residents of Little Hocking also had increased odds 

ratio 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). 
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Analysis of exposure category and cancer type showed an association between the very 

high PFOA exposure category and several cancers, but ORs for lower exposure categories 

generally did not support a positive dose-response relationship. 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 compared to the unexposed. 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 

due to small numbers. ORs for the other cancer categories were all <1.0. Ovarian cancer, nonHodgkin’s lymphoma, and prostrate 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).

4.2 Animal Studies 

4.2.1 Acute Toxicity 

Oral Exposure 

Dean and Jessup (1978) reported an oral LD50 of 680 mg/kg and 430 mg/kg 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 (1976c) reported 

an oral LD50 of less than 1000 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.

Inhalation Exposure

Rusch (1979) reported no mortality in male or female Sprague-Dawley rats following

inhalation exposure to 186,000 mg/m3

 for one hour. Kennedy et al. (1986) reported a 4-hour 

LC50 of 980 mg/m3

 for groups of 6 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 2000 

mg/kg (Glaza, 1995). Kennedy (1985) determined a dermal LD50 of 4300 mg/kg for rabbits, 7000 

mg/kg for male rats, and 7500 mg/kg for female rats. The animals lost body weight, 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, 1976d), 

but is not an irritant in rabbits when washed from the eye (Gabriel, 1976a). Markoe (1983) found 

PFOA to be a skin irritant in rabbits, while Gabriel (1976b) did not.

4.2.2 Short-Term Studies 

Oral Exposure 

Monkey. In a range-finding study, Thomford (2001a) administered PFOA to male Cynomolgus 

monkeys as an oral capsule containing 0, 2, or 20 mg/kg-day PFOA for 4 weeks. There were 3 

monkeys in the 2 and 20 mg/kg-day groups and one monkey in the control group. Animals were 
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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 one 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, thyroid stimulating hormone, total and free triiodothyronine, and total and free 

thyroxin). Blood samples were also collected from each animal on day 2 (approximately 24 

hours after the first dose) for clinical chemistry measurements. 

At scheduled necropsy, liver samples were collected for palmitoyl coenzyme A (CoA) 

oxidase activity (a biomarker for peroxisome proliferation) and serum PFOA. Liver, testes, and 

pancreas were collected and assayed for cell proliferation using proliferation cell nuclear 

antigen. 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 treated monkeys. No adverse effects were noted in either gross or clinical pathology 

studies. Under the conditions of this study, the no observed adverse effect level (NOAEL) was 

20 mg/kg and no lowest observed adverse effect level (LOAEL) was established.

Rat. Pastoor et al. (1987) dosed male Crl:CD (SD) BR rats (n=6/group) for 1, 3, or 7 days with 0 

or 50 mg PFOA/kg. Liver sections were collected at necropsy and stained with hematoxylin and 

eosin. Sections were also examined by electron microscopy. DNA content was also determined 

from 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 not different in rats treated for 1 or 3 days compared to control rats. 

Liver weight of rats treated for 1 day was not different from 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% decreased (p<0.05) was observed in relative 

hepatic DNA/g liver, but no difference was observed between total amount of hepatic DNA/liver 

when compared to the total amount of DNA/liver in control rats. 

The hepatocytes of rats treated with PFOA for three days were enlarged with partially 

occluded sinusoids, had numerous basophilic granules, eosinophillic granular material in the 

cytoplasm, and fewer perinuclear glycogen vacuols compared to control hepatocytes. Enlarged 

with hyperplastic smooth endoplasmic reticulum, increased mitochondria, increased 

peroxisomes, decreased rough endoplasmic reticulum, and increased autophagosomes with 

electron-dense material were also observed in the hepatocytes.

Loveless et al. (2008) administered 0, 0.3, 1, 10, or 30 mg linear PFOA/kg by oral gavage 

to groups of male CD rats (n=10/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 measurements. Tissues were collected for weight and microscopic examination. 
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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, minimum to mild (0.3, 1 mg/kg/day) and moderate (≥10 mg/kg/day) 

hepatocellular hypertrophy, and focal necrosis (≥10 mg/kg/day) were observed. Although not 

statistically significant, serum corticosterone was increased at ≥ 10 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 4.3.2 of this report. 

Cui et al. (2009) exposed male Sprague-Dawley rats (10/group) to PFOA (96% a.i.) at 0, 

5, or 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 (hepatic, renal, gonal 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, 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 effects were observed 

in the kidneys of the low dose animals, but turbidness and tumefaction (swelling) in the 

epithelium of the proximal convoluted tubule was observed at 20 mg PFOA/kg. Under the 

conditions of this study, the LOAEL was 5 mg/kg/day based on liver and pulmonary effects, and 

no NOAEL was established. 

Male Sprague-Dawley rats (n=10/group) were fed diets containing 0 or 300 pm PFOA for 

1, 7, or 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. Liver 

DNA content and concentration were determined, and plasma was collected for analysis (study 1 

only) of liver enzymes, cholesterol, triglycerides, and glucose. Hepatic cell proliferation and 

apoptosis were also 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 4-4). 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 ratio were significantly increased in Wyeth 14,643 diet fed rats 

in both studies. 

 
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After 1 day of eating the PFOA diet, plasma AST was significantly decreased and 

triglycerides were significantly increased. After 7 and 28 days on the PFOA diet, total 

cholesterol, triglycerides, and glucose levels were significantly decreased. AST was also 

significantly decreased after 28 days on the PFOA diet. After 1 day on the Wyeth 14,463 diet, 

AST and total cholesterol were significantly decreased. After 7 and 28 days on the Wyeth 

14,643 diet, ALT, total cholesterol, 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. Liver DNA concentration was significantly decreased (p<0.05) in all PFOAexposed rats except those treated for 1 day in the second study, but liver DNA content was not 

altered by PFOA. 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. Liver DNA 

content in the Wyeth 14,643 treated rats was significantly increased after 7 and 28 days in both 

studies.

Labelling indices for hepatic cell proliferation, as measured by 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 comparisons. 

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 4-5). 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. 

TABLE 4-5. Hepatic Effects of Rats Exposed to PFOA

Day

Study 1 Study 2

Control 300 ppm 

PFOA 

50 ppm

Wyeth 

14,643

Control 300 ppm 

PFOA 

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*

From Elcombe et al., 2010

*Significantly different from control (p<0.05); ND=No Data

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 after 7 days, 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 compared to those in PFOA-diet fed rats. 

Mouse. Kennedy (1987) fed male and female CRL:CD-1 mice diets containing 0, 30, 300, or 

3000 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 3000 ppm. Body 
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weight was decreased at 300 ppm and one female died at that dose. Both male and female mice 

had significantly increased (p<0.05) absolute and relative liver weights at all doses compared to 

the control. The 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, or 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, or 30 mg linear PFOA/kg by oral gavage 

to groups of male CD-1 mice (n=20/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 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 

corticosterone levels were significantly increased in mice dosed with 10 mg/kg and elevated in 

those dosed with 30 mg/kg/day. Total 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. Data on several immunological endpoints were reported as part of the Loveless et al. 

(2008) publication. The immunological data are included in Section 4.3.2 of this report. 

 

Son et al. (2008) administered 0, 2, 10, 50, or 250 mg/L PFOA (0, 0.49, 2.64, 17.63, or 

47.21 mg/kg PFOA) in the drinking water to 4 week old male 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 transcriptase-polymerase chain reaction (RT-PCR). 

The mice exposed to 250 mg/L PFOA had significantly reduced (p<0.05) food and water 

consumption, and body weight gain (p<0.05) compared to control mice. Body weight gain was 

also 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 control mice. Relative kidney weight was not 

affected by PFOA exposure. At ≥ 10 mg/L PFOA, plasma ALT activity was significantly 

increased, and at ≥ 50 mg/L PFOA, plasma AST activity was significantly elevated compared to 

the activity level in 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 in 

the liver were observed with staining for caspase 3. Exposure to PFOA did not affect kidney 
Perfluorooctanoic Acid – February 2014 4-35 

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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. 

Wolf et al. (2008a) gavage dosed wild-type 129S1/SvlmJ mice (n=7-8/group) and 

PPARα-null mice (129S4/SvJae-PPARαtm1Gonz/J, n=6-8/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/group) with 0, 1, or 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 proliferating cell nuclear antigen 

(PCNA). Hepatocyte hypertrophy and vacuolation, observed in both strains of wild-type mice, 

were assigned a score from 0-4 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-

1000 hepatocyte nuclei/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 wild-type mice exposed to PFOA and were also significantly increased (p<0.05) in the wildtype 129S1/SvlmJ mice exposed to Wyeth 14,643 (Table 4-6). 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 not different from control 

values. 
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TABLE 4-6. 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 PFOA 2.26±0.24* 6.5±0.5* 2.1±0.9 0.7±0.5

10 mg/kg 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 PFOA 1.22±0.22* 1.6±0.2* 2.0±0.0* 0.7±0.6

3 mg/kg PFOA 1.70 ±0.12* 6.4±0.4* 2.0±0.0* 1.0±0.4

10 mg/kg PFOA 2.20±0.23* 8.3±0.2* 4.0±0.0* 2.4±0.9*

50 mg/kg 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 PFOA 1.2±0.14* 4.5±0.2* 1.9±0.6* 0.6±0.4

3 mg/kg PFOA 1.46±0.21* 5. 8±0.3* 3.0±0.0* 0.6±0.3

10 mg/kg PFOA 2.8±0.18* 9.4±0.6* 4.0±0.0* 7.7±3.0*

50 mg/kg Wyeth

14,643

1.07±0.24 3.9±0.5 1.4±0.5 0.6±0.5

From Wolf et al., 2008a 

* Statistically different from control, p< 0.05

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 were similar, but treatment with PFOA resulted in a different 

response. Table 4-7 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 endoplasmic reticulum (ER), mitochondrial, 

peroxisomes, and lipid-like cytoplasmic vacuoles. 
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TABLE 4-7. 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

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

Limited Limited Not reported Not reported Numerous*

PPARα-null/

Wyeth

Prominent Prominent Numerous Absent Scattered

From Wolf et al., 2008a. 

*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 4-7 that both 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 PFOAdosed PPARα-null mice had numerous large non-membrane 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, and 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, or 

0.3 mg/kg/day PFOA (n=4-6/group) for 2 weeks and sacrificed 18-20 hours following the last 

dose. Blood was collected and analyzed for triglycerides and cholesterol concentrations and 

ALT measurements. Livers were collected and analyzed for triglycerides and cholesterol 

concentration, and histopathological changes. The differences in the observations for the three 

strains of mice are summarized in Table 4-8. 
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TABLE 4-8. 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)

From Nakamura et al., 2009

Mice were dosed with with 0, 0.1, or 0.3 mg/kg/day PFOA

↑=significant increase (p<0.05); ↓= significant decrease (p<0.05); ND= no differences 

Body weight of the hPPARα mice was slightly lower than mPPAR and PPA-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 4-

7). 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 PFOA 

based on increased liver weight and increased liver triglycerides and cholesterol concentrations. 

The NOAEL for mPPARα mice was 0.1 mg/kg PFOA. The NOAEL for PPARα-null mice was 

0.3 mg/kg 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 PFOA, the highest dose tesed. However, a nonsignificant, dose related 

increase was observed in plasma cholesterol.

Minata et al. (2010) examined hepatobiliary injury in mice treated with PFOA. Male 

wild-type 129S4/SvlmJ mice (n=39) and PPARα-null (129S4/SvJae-Pparαtm1Gonz/J mice (n=40)

were orally dosed with 0, 12.5, 25, or 50 μmol/kg PFOA (~0, 5.4, 10.8, or 21.6 mg/kg 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 
Perfluorooctanoic Acid – February 2014 4-39 

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protein 2 (Mdr2) and tumor necrosis factor α (TNF-α) mRNA 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 multidrug resistance-associated protein 2 (MRPs). 

 

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 and in PPARα-null mice at 

50 μmol/kg compared to the concentrations of their respective controls. Plasma ALT was not 

different from control in the treated mice. In wild-type mice, total bilirubin was significantly 

decreased at 12.5 μmol/kg and significantly increased at 50 μmol/kg. In PPARα-null mice, total 

bilibrubin was significantly increased at 50 μmol/kg. Total bile acid was significantly increased 

at 50 μmol/kg in PPARα-null mice. Total cholesterol was significantly decreased in wild-type 

mice at 25 and 50 μmol/kg and total triglyceride was significantly increased at 12.5 and 25 

μmol/kg. Total cholesterol was significantly decreased at 12.5 and 25 μmol/kg and significantly 

increased at 50 μmol/kg in PPARα-null 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, or 50 

μmol/kg. 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. Increased apoptosis in hepatic cells, 

hepatic arterial walls, and bile duct epithelium was observed at 25 and 50 μmol/kg in wild-type 

mice. Ultrastructure examination of livers from PFOA treated wild-type mice showed decreased 

glycogen granules, degranulated or disrupted rough endoplasmic reticulum, nuclear vacuoles, 

extensive peroxisome proliferation, and slight mitochondria proliferation. 

In PPARα-null mice treated with 12.5, 25, or 50 μmol/kg PFOA, hypatocellular 

hypertrophy, cytoplasmic vacuolation, and increased microvesicular steatosis were observed. 

These observations are consistent with Wolf et al. (2008a). At 50 μmol/kg, 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. Increased apoptosis was observed in bile 

duct epithelium at 25 and 50 μmol/kg in PPARα-null mice. Ultrastructure examination of livers 

from PFOA treated PPARα-null mice showed decreased glycogen granules, degranulated or 

disrupted rough endoplasmic reticulum, increased cytoplasmic lipid accumulation, mitochondria 

proliferation, and mitochondrial changes including swelling and decreased matrix density. 

Peroxisome proliferation was not observed. Ultrastructure of bile duct showed degradation of 

cytoplasmic structure, vacuolization, disintegrations of nuclei and organelles, and periductal 

infiltration of fibroblasts and macrophages as well as fibrosis. 

The marker for oxidative damage, 8-hydroxydeoxyguanosine, and TNF-α were not 

elevated or upregulated in wild-type mice. In PPARα-null mice, 8-hydroxydeoxyguanosine was 

elevated in the liver at 50 μmol/kg and TNF-α mRNA was significantly increased at 25 and 50 

μmol/kg. The transporter Mdr2 moves biliary phospholipids from hepatocytes to bile and was 

significantly upregulated in wild-type mice at all doses, but only at 12.5 μmol/kg in PPARα-null 

mice. The BSEP transports bile acid from hepatocytes to bile and was significantly decreased in 

wild-type mice at 50 μmol/kg, significantly increased in PPARα-null mice at 12.5 μmol/kg, and 
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significantly decreased at 50 μmol/kg. The transporter MRP2 also transports bile acid and was 

significantly decreased at 50 μmol/kg in both groups of rats. 

Under the conditions of the study, the LOAEL for male wild-type and PPARα-null mice 

was 12.5 μmol/kg PFOA (~5.4 mg/kg) based on increased liver weight. A NOAEL was not 

established.

Li et al. (2011) investigated the involvement of mouse and human PPARα in PFOAinduce testicular toxicity. Wild-type, PPARα-null, and humanized PPARα male 129/Sv mice 

were given PFOA daily by gavage at doses of 0, 1, or 5 mg/kg/d for six weeks. Body weight and 

testis weight were not affected by treated in any group. Absolute and relative-to-body weights of 

the epididymis and seminal vesicle plus prostate gland were decreased only in high-dose wildtype 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 wildtyp 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 

wild-type 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 significane 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; peripherial 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 of 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 was decreased in both groups of wild-type and high-dose humanized 

PPARα mice and superoxide dismutase 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.

Inhalation Exposure

No data on the effects of short term inhalation exposures to PFOA were identified in the 

literature.
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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%, or 1.5% PFOA (equivalent to 0, 0.25, 2.5, 6.25, 12.5, 25, or 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. 

4.2.3 Subchronic Studies 

Oral Exposure 

Monkey. Goldenthal (1978b) administered rhesus monkeys (2/sex/group) doses of 0, 3, 10, 30 

or 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 which 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/kgday 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 

slight to 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 vs. 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 one and three 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. 
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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 

to moderate 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 based on decreased heart and brain weight, and the NOAEL was 3 mg/kg. 

Rat. Goldenthal (1978a) administered ChR-CD rats (5/sex/group) dietary levels of 0, 10, 30, 

100, 300, and 1000 ppm PFOA for 90 days. These dose levels are equivalent to 0, 0.56, 1.72, 

5.64, 17.9, and 63.5 mg/kg-day in males, and 0.74, 2.3, 7.7, 22.36 and 76.47 mg/kg-day in 

females. Animals were observed twice daily and body weight and food consumption were 

recorded weekly. Blood and urine samples were collected during the pretest period and at 1 and 

3 months of the study for hematology, clinical chemistry and urinalysis. At necropsy, the organs 

from the control, 100, 300, and 1000 ppm groups were weighed and examined for 

histopathologic lesions; livers from the 10 and 30 ppm groups were also examined 

microscopically.

There were no treatment-related changes in behavior or appearance and no treatment 

related deaths. There was a decrease in body weight gain for male rats at the 300 and 1000 ppm 

dose levels. At 13 weeks, mean body weight of males in the 1000 ppm group was significantly 

less than that of controls. There were no treatment related effects on the hematologic, 

biochemical or urinary parameters.

Relative kidney weights were significantly increased in males in the 100, 300, and 1000 

ppm groups. However, absolute kidney weights were comparable among groups, and there were 

no histopathological lesions. Absolute liver weights were significantly increased in males in the 

30, 300 and 1000 ppm groups and in females in the 1000 ppm group. The mean absolute liver 

weight for the 0, 10, 30, 100, 300, and 1000 ppm groups was 13.4, 14.3, 19.1, 18.6, 20.1, and 

19.2 g, respectively. Relative liver weights were significantly increased in males in the 300 and 

1000 ppm groups and in females in the 1000 ppm group. Discoloration on the surface of the liver 

was observed in male rats in the 1000 ppm group. Hepatocellular hypertrophy (focal to 

multifocal in the centrilobular to midzonal regions) was observed in 4/5, 5/5, and 5/5 males in 

the 100, 300, and 1000 ppm groups, respectively. Hepatocyte necrosis was observed in 2/5, 2/5, 

1/5, and 2/5 males in the 30, 100, 300, and 1000 ppm groups, respectively. Under the conditions 

of this study, the LOAEL for males is 30 ppm (1.72 mg/kg-day) based on liver effects and the 

NOAEL is 10 ppm (0.56 mg/kg-day); the LOAEL for females is 1000 ppm (76.5 mg/kg-day) 

based on increased liver weight, and the NOAEL is 300 ppm (22.4 mg/kg-day). 

In a dietary study, reported in Palazzolo (1993) and Perkins et al. (2004), male ChR-CD 

rats (45-55 per group) were administered concentrations of 1, 10, 30, or 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 control group pair-fed to the 100 ppm dose 

group); both fed the basal diet. Following the 13-week exposure period, 10 animals per group 
Perfluorooctanoic Acid – February 2014 4-43 

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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, or 13 weeks of treatment; 

10 animals per group were sacrificed after 13 weeks of treatment and following the 8 week 

recovery period. Serum samples collected from 10 animals per group at each scheduled sacrifice 

during treatment and from 5 animals per group during recovery were analyzed for estradiol, total 

testosterone, luteinizing hormone, and PFOA. The level of palmitoyl CoA oxidase was analyzed 

from a section of liver that was obtained from 5 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 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 in groups 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 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, when compared to 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 

appeared to be some indication of elevated estradiol for the 100 ppm group at week 5. 

Significant increases in absolute and relative liver weights and hepatocellular 

hypertrophy were observed at weeks 4, 7, or 13 in the 10, 30 and 100 ppm groups (Table 4-9). 

There was no evidence of any 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. During recovery, 

however, no liver effects were observed, indicating that the treatment-related liver effects were 

reversible. Under the conditions of this study, the LOAEL is 10 ppm (0.64 mg/kg-day) based on 

increases in absolute and relative liver weights with hepatocellular hypertrophy. The NOAEL is 

1.0 ppm (0.06 mg/kg-day).
Perfluorooctanoic Acid – February 2014 4-44 

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TABLE 4-9. 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

5± 0.4

7 ± 1.5

5± 1.1

9± 1.7

7± 0.8

8 ±1.9

14± 3.8c

18 ±5.5

10 ±2.1

24 ± 11.4c

32 ± 12.2c 

14 ± 3.4c 

37± 14.8cd

54± 35.3cd

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

2/15

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 

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

3.23±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±0.59c 

5.49±0.84c 

From Palazzolo 1993; Perkins et al. 2004

Mean ±SD; NA= not available

a non-pair-fed controls; b pair-fed controls; c statistically significant (p < 0.05) with non-pair-fed control; d statistically 

significant (p < 0.05) with pair-fed control 

Inhalation and Dermal Exposure 

No data on the effects of subchronic inhalation or dermal exposures to PFOA were 

identified in the literature.

4.2.4 Neurotoxicity 

Johansson et al. (2008) gave male neonatal Naval Medical Research Institute mice 

(NMRI, 3-4 litters, ~5-6 male pups/litter) a single gavage dose of 0, 0.58, or 8.7 mg PFOA/kg in 

a lecithin/peanut oil emulsion on postnatal day 10, the approximate peak time of rapid brain 

growth in mice. Spontaneous behavior (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 minute periods. The habituation 

ratio was determined by dividing the activity for the third 20 minute 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 minutes and comparatively higher activity for the last 20 minutes. 

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 when compared to 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 on rearing behavior. 

 

At 4 months of age, mice were tested for nicotine-induced behavior and behavior in the 
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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 were also 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. There were significant 

differences in locomotion and total activity at 4 months in the low dose animals which support 

identifying the 0.58 mg/kg dose as a marginal LOAEL. However, the data at the low dose are 

less compelling as those for 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/litter) a single gavage dose of 0 or 8.7 mg PFOA/kg on 

postnatal day 10. Protein levels of calcium/calmodulin-dependent protein kinase II (CaMKII), 

growth-associated 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 not changed 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 not changed in the cerebral cortex. Synaptophysin 

levels in mice exposed to PFOA were significantly increased in the hippocampus (52%) and 

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 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/group) to 0 or 0.3 

mg PFOA/kg/day in the diet from GD1-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 minutes was registered in 5 minute 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 (3 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 (4 trials) was measured by how long the mice remained on a rotating drum as 

a rotarod accelerated from 4 to 40 rpm over 5 minutes. 
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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 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 Day1 through the dark phase Day 2. Both male and female offspring 

exposed to PFOA had significantly (p<0.05) less inactive periods 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 4 trials compared to control females, but no 

effect of treatment was observed in male offspring. The authors concluded that prenatal 

exposure to PFOA resulted in sex-related alterations in offspring behavior and motor function. 

4.2.5 Developmental/Reproductive Toxicity 

Reproductive Effects

A standard oral two-generation reproductive toxicity study of PFOA in Sprague-Dawley 

rats was conducted (York, 2002; Butenhoff et al., 2004a; York et al., 2010). Five groups of male 

and female SD rats (30/sex/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/sex/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/sex/group) were dosed similarly, beginning at weaning. The F2 generation 

rats were maintained through lactation day 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 were also determined. Food 

consumption, body-weight gain, selected organ weights, gross pathology, and appropriate 

histopathology were evaluated. 

F0 Males. One F0 male rat in the 30 mg/kg-day dose group was sacrificed on day 45 of the study 

due to adverse clinical signs. Statistically significant increases in clinical signs were also 

observed in male rats in the high-dose group that included dehydration, urine-stained abdominal 

fur, and ungroomed coat. Significant reductions in body weight and body weight gain were 

reported for most of the dosage period and continuing until termination of the study in the 3, 10, 

and 30 mg/kg-day dose groups. Absolute food consumption values were also significantly 

reduced during these periods in the 30 mg/kg-day dose group, while significant increases in 

relative food consumption values were observed in the 3, 10, and 30 mg/kg-day within those 

same periods.

No treatment-related effects were reported at any dose level for any of the reproductive 

parameters assessed. No treatment-related effects were seen at necropsy or upon microscopic 

examination of the reproductive organs. 

At necropsy, statistically significant reductions in terminal body weights were seen at 3, 

10, and 30 mg/kg-day (6%, 11%, and 25% decrease from controls, respectively; p≤0.05). The 

absolute weight of the liver was statistically significantly increased in all dose-groups (Table 4-

10). Absolute kidney weights were statistically significantly increased in the 1, 3, and 10 mg/kg-
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day dose groups, but significantly decreased in the 30 mg/kg-day group. Organ weight-toterminal body weight ratios for the liver and for the left and right kidney were statistically 

significantly increased in all treated groups. Organ weight-to-terminal body weight ratios for the 

brain were statistically significantly increased for the 3, 10, and 30 mg/kg-day dose groups; 

organ weight-to-brain weight ratios were significantly reduced for some organs at the high dose 

group, and significantly increased for other organs among all treated groups; these changes 

occurred in a pattern typically associated with decrements in body weight. 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 group and 7/10 males in 

the 30 mg/kg-day dose group. 

TABLE 4-10. Organ weight data from F0 males

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*

From Butenhoff et al. 2004a.

Mean±S.D.; n = 29-30; significantly different from control: *p<0.05, **p<0.01.

Under the conditions of the study, the LOAEL for F0 parental males is 1 mg/kg-day, the 

lowest dose tested, based on significant increases in absolute and relative liver and kidney 

weight. A NOAEL for the F0 parental males could not be determined. 

F0 Females. 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 or upon 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 (sexes combined) was significantly 

reduced (p≤0.01) by 8-10% throughout the first two weeks of lactation in the 30 mg/kg-day 

group; at weaning the mean body weights were reduced 4.5%, but the difference was not 

statistically significance. 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 lactation days 6-8. No other effects were noted, and 

there were no treatment-related findings for the pups necropsied at weaning. 

F1 Males. Significant increases in treatment-related deaths (5/60 animals) were reported in F1 
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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 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. 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 (p≤ 0.01) in the average day of preputial separation 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 were noted.

The absolute and relative weights of the liver were statistically significantly increased in 

all treated groups (p ≤ 0.01). The absolute weights of the left and/or right kidneys were 

statistically significantly increased in the 1 and 3 mg/kg-day dose groups and statistically 

significantly decreased in the 30 mg/kg-day dose group (Table 4-11). Organ weight-to-terminal 

body weight and brain weight ratios for the kidney were statistically significantly increased in all 

treated groups. Histopathology was not conducted on the kidney. All other organ weight changes 

observed (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 (cytoplasmic hypertrophy and vacuolation of the cells of the adrenal cortex) and in the 

liver of animals treated with 3, 10, and 30 mg/kg-day (hepatocellular hypertrophy). 
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TABLE 4-11. Organ weight data from F1 males

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

From Butenhoff et al. 2004a.

Mean±S.D.; n = 29-30; significantly different from control: *p<0.05, **p<0.01.

For F1 males, the LOAEL for developmental/reproductive toxicity was 10 mg/kg-day 

based on significantly reduced body weight and body weight gain during the lactation period, 

and the NOAEL was 3 mg/kg-day. 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 

(observed prior to and during cohabitation and during the entire dosing period), and in terminal 

body weights; and significant changes in absolute and relative liver weights. A NOAEL for adult 

systemic toxicity in the F1 males could not be determined. 

F1 Females. 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 weights were observed in high-dose animals 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 (p≤0.01) delays 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 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. 

Both 3M (2002) and the U.S. EPA (2002b) conducted a more detailed analysis of the 

estrus cycle data. The 3M analysis of the data concluded that there were no differences in the 
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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 the U.S. EPA (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 or 4 days of cornified smears), possibly pseudopregnant, (defined as 6-

greater days of leukocytes) or persistent estrus (defined as 5-or greater days of cornified smears). 

The two groups were not different in any of the parameters measured. Thus, the increase in the 

number of estrous stages per 21 days that was 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 

females. 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 did not reveal any 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. There was a statistically significant increase (p< 0.01) in the number of pups 

found dead on lactation day 1 in the 3 and 10 mg/kg-day groups. An independent statistical 

analysis was conducted by U.S. EPA (2002c), and no significant differences were observed 

between dose groups and the response did not have any trend in dose. There were no treatmentrelated effects on any of the developmental parameters assessed, and at necropsy no treatmentrelated effects were noted. The NOAEL for developmental/reproductive toxicity in the F2 

offspring was 30 mg/kg-day. 

Developmental Effects

Prenatal Development. Prenatal developmental toxicity studies of PFOA have been conducted 

in rats, and mice and are summarized below. A greater number of studies have been conducted 

using mice as experimental animals than rats. Most studies have used the oral route of exposure.

Rat. Pregnant Sprague-Dawley rats were gavage dosed with 0, 3, 10 or 30 mg/kg-day PFOA

during days 4-10, 4-15, or 4-21 of gestation, or from gestation day 4 to lactation day 21 

(Hinderliter et al., 2005; Mylchreest, 2003). Clinical observations and body weights were 

recorded daily. On gestation days 10, 15, and 21, 5 rats/group/time-point were sacrificed and the 
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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 minutes post-dose. The remaining 5 rats 

per group were allowed to deliver. On lactation days 0, 3, 7, 14, and 21, the pups were counted, 

weighed (sexes separate), and examined for abnormal appearance and behavior. Randomly 

selected pups were sacrificed and blood samples were collected. On lactation days 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.

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 the control group during 

gestation, and mean body weights were approximately 4% lower than 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 (litter size of 3-6 pups as 

compared to 12-19 pups/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 pup 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 in order to 

ascertain whether there was a sex difference in the bioaccumulation of PFOA in the mouse, and 

to evaluate the effects of PFOA on pre- 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, or 40 mg/kg PFOA daily by oral gavage on GD1-17. Maternal weight was monitored 

during gestation. Dams were divided into two groups. In the first group, dams were sacrificed on 

GD18 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 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 GD18. Dams were allowed to give birth on GD19. 

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 four 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 sex. The time to sexual maturity was determined by monitoring vaginal 

opening and preputial separation beginning on PND 24. Two to four pups per sex 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. Benchmark dose values (BMD5 and BMDL5) were 
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calculated for the maternal and developmental toxic endpoints. 

Signs of maternal toxicity were observed following exposure to PFOA during pregnancy. 

Statistically significant (p < 0.05) dose-related increases in maternal liver weight were also 

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 

(p < 0.05) increases 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 decreased at all doses except 5 

mg/kg. Ossification of hindlimb proximal phalanges was significantly decreased 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 (p < 0.05) increases 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 (p < 0.05), increases 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 at 10 and 20 mg/kg/day dose group. Postnatal survival and 

viability at the 1 and 3 mg/kg/day dose groups was 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 were also 

manifested by significant (p < 0.05) delays in eye opening at doses > 5 mg/kg/day. Slight delays 

(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 for males was indicated based on accelerated pubertal development, and a 

NOAEL was not established. For females, the developmental LOAEL was 3 mg/kg based on 

growth retardation, and the NOAEL was 1 mg/kg. 

For maternal toxicity, BMD5 and BMDL5 estimates for decreases in maternal weight gain 

during pregnancy were 6.76 and 3.58 mg/kg, respectively. For increases in maternal liver weight 

at term, BMD5 and BMDL5 estimates were 0.20 mg/kg and 0.17 mg/kg, 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, respectively. Significant alterations in postnatal growth 
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and development were observed at 1 and 3 mg/kg/day, with BMD5 and BMDL5 estimates of 1.07 

and 0.86 mg/kg, respectively, for decreased pup weight at weaning; and 2.64 and 2.10 mg/kg, 

respectively, for delays in eye opening. The BMD5 and BMDL5 estimates for reduced phalangeal 

ossification were < 1 mg/kg. BMD5 and BMDL5 estimates for reduced fetal weight at term were 

estimated to be 10.3 and 4.3 mg/kg, 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 (WT)] and PPARα-null mice were orally dosed from GD1-17 with 0, 

0.1, 0.3, 0.6, 1, 3, 5, 10, or 20 mg PFOA/kg. Heterozygous (HET) litters were also produced by 

mating WT and PPARα-null males with WT and PPARα-null dams to determine if genetic 

background affected survival. The HET litters were sacrificed on PND15. Survival at birth was 

recorded and live offspring counted and weighed by sex. Litters were counted and offspring 

weighed on PND1-10, 14, 17, and 22. Weaning occurred on PND 22, and all dams and one 

pup/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 

non-pregnant females and those with full litter resorptions), number of implants, or pup weight at 

birth. WT 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 WT dams and 20 mg/kg/day in PPARα-null dams, 100% litter loss occurred. 

Relative liver weight was significantly increased in WT adult females dosed with ≥1 mg/kg and 

in PPARα-null adult females dosed with ≥3 mg/kg. 

Body weight in WT offspring born of dams dosed with 1.0 mg/kg was significantly 

reduced (p<0.05) compared to control offspring body weight gain on PND9, 10, and 22 (males) 

and PND7 to10 and PND22 (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 WT litters exposed to ≥0.6 mg/kg, but was not 

affected in PPARα-null litters. Survival was significantly decreased (p<0.05) for WT and HET 

pups born to WT dams dosed with 1 mg/kg and for HET pups born to PPARα-null dams dosed 

with 3 mg/kg. Offspring born of WT dams showed a dose-related trend for 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 PPARα-null dams. At weaning, 

relative liver weight was significantly increased (p<0.05) in WT offspring gestationally exposed 

to ≥0.1 mg/kg 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 may 

also contribute. Under the conditions of the study, the maternal/reproductive LOAEL for WT 

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 WT 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. 
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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 or 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 GD18 (n = 5-8/group) or allowed to give birth and then sacrificed, 

along with their litters (n = 8-14), on PND20. Livers from dams, fetuses, and pups were weighed 

and collected for histopathological evaluation and RNA analysis. Gene expression results are 

given in section 4.3.4. Mammary gland whole mounts were prepared from female pups on 

PND20 for quantification of ductal length and number of terminal end buds. 

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 compared with 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 the other genotypes. 

Maternal liver weight was significantly increased in the treated groups of all genotypes 

on GD18 and in wild-type animals on PND20. Maternal liver weight was not affected on 

PND20 in the PPARα-null or PPARα-humanized mice. Relative fetal liver weight on GD18 was 

significantly increased in fetuses from treated wild-type and humanized dams. On PND20, 

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 PFOA 

treated groups on GD18 and PND20, with decreased incidence and severity by PND20. On 

GD18 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 4.4.1. Only wildtype fetuses and pups from treated dams showed similar liver lesions.

The average length of mammary gland ducts and the average number of terminal end 

buds per mammary gland per litter were not different between control and treatment groups from 

all genotypes. 

Yahia et al. (2010) gavage dosed pregnant ICR mice (n=5/group) with 0, 1, 5, or 10 mg 

PFOA/kg from GD0-17 or 18. The dams dosed from GD0-17 were sacrificed on GD18, and the 

fetal skeletal morphology was evaluated. Dams dosed from GD0-18 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. Maternal absolute liver weight was 

significantly increased (p<0.05) at doses ≥ 5 mg/kg 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 
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medullar and proximal tubule were slightly hypertrophic at all doses. Treatment with 10 mg/kg 

caused a significant increase in AST, ALT, GGT, and ALP and a significant decrease in total 

serum protein, albumin, globulin, triglycerides, phospholipids, total cholesterol, and free fatty 

acids. At 5 mg/kg, total serum protein and globulin were significantly decreased, and 

phosphorus was increased. At 1 mg/kg, BUN and phosphorus were significantly increased. The 

maternal LOAEL was 1 mg/kg 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, 

increased incidence of cleft sternum, delayed phalanges ossification, and delayed eruption of 

incisors was observed. Delayed parturition was observed in dams treated with 10 mg/kg, 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 was significantly reduced 

compared to control pup body weight. By PND4, 16% of offspring born to dams dosed with 5 

mg/kg had died. No pathological differences were observed in the lungs or brains of treated and 

control offspring. The developmental LOAEL was 5 mg/kg based on decreased body weight and 

decreased survival rate, and the NOAEL was 1 mg/kg. 

Suh et al. (2011) examined placental prolactin-family hormone and fetal growth 

retardation in mice treated with PFOA. Pregnant CD-1 mice (n=10/group) were treated with 0, 

2, 10, or 25 mg/kg PFOA from GD11 to GD16. Dams were sacrificed on GD16 and uteri were 

removed and examined. Three placentas/group were analyzed using histochemistry and the 

numbers of glycogen tropholblast cell (GlyT) in the junctional zone and sinusoidal trophoblast 

giant cells (S-TGC) in the labyrinth zone were counted and compared. Trophoblast cells express 

prolactin-family genes. mRNA from three placentas/group were analyzed using situ 

hybridization, northern blot hybridization, and RT-PCR for mouse placental lactogen (mPL)-II,

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 GD13-16 in dams 

treated with 25 mg/kg PFOA compared to controls. At ≥2 mg/kg PFOA, placental weight was

significantly decreased and the number of resorptions and dead fetuses was significantly 

increased. At ≥10 mg/kg 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 

post-implantation loss was 3.87, 8.83, 30.98, and 55.41% for the 0, 2, 10, and 25 mg/kg PFOA 

groups, respectively. 

The placentas of dams dosed with ≥10 mg/kg PFOA displayed necrotic changes. Parietal 

and sinusoidal TGC and GlyT cell frequency in the placental junctional and labyrinth zones was 

significantly decreased (>0.05) in a dose-dependent manner in treated dams. At 25 mg/kg 

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 express may be secondary to insufficient 
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trophoblast cell differentiation and increased cell necrosis. These effects reduced placental 

efficiency and contributed in part to fetal growth retardation. 

Other Developmental Studies/Mammary Gland Development. The following studies used 

study designs and/or examined endpoints not usually included in developmental studies. The 

studies were conducted to determine critical periods of exposure focusing 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 direct exposure of 

peripubertal animals starting about the time of weaning. 

Indirect gestational and/or lactational exposures 

An overview of experiments designed to assess the developmental effects of PFOA 

exposure to CD-1 mice during gestation is given in Table 4-12. Additional details of the studies 

are described in the text below. All studies included measures of maternal and offspring survival 

and body weight. 

TABLE 4-12. Overview of studies in which pregnant CD-1 mice were administered PFOA

Dose

(mg PFOA 

/kg/day)

Timing Endpoints Reference

0, 5 GD 1-17, 8-17, or 12-

17

Body weight; mammary gland GD 18 (dams) 

and PND 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 of 

female pups up to 18 months

Wolf et al., 

2007; White et 

al., 2009

0, 3, 5 GD 1-17

Cross-fostered at birth

Body weight; developmental landmarks and 

growth to PND 245; mammary gland of 

female pups up to 18 months

Wolf et al., 

2007; White et 

al., 2009

0, 5 GD 8-17

Cross-fostered at birth

Mammary gland of dams and female pups on 

PND 1, 3, 5, 10

White et al., 

2009

0, 0.3, 1.0, 3 GD 1-17 Liver weight; mammary gland of female 

pups PND 7, 14, 21, 28, 42, 63, 84

Macon et al., 

0, 0.01, 0.1, 2011

1

GD 10-17

0, 1, 5 GD 1-17 Body weight; reproductive parameters; 

mammary gland of F0, F1, and F2 females

White et al., 

0, 1 2011

+ 5 ppb in 

drinking 

water

GD 1-17; drinking 

water started GD 7 and 

continued to F2 

generation

White et al. (2007) orally dosed pregnant CD-1 mice with 0 or 5 mg PFOA/kg on 

gestation days (GD) 1-17 (n=14), 8-17 (n=16), or 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 dams and litters were sacrificed on PND10 
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and remaining dams and litters were sacrificed on PND20. 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. On GD18, stunted alveolar development was 

observed in the mammary gland of dams treated with PFOA on GD1-17 compared to the 

mammary glands of the control dams which were saturated with milk-filled alveoli. Dams 

treated with PFOA on GD 1-7 or 8-17 exhibited significant mammary gland epithelial 

differentiation delays on PND10 as evidenced by an excessive amount of adipose tissue. In 

comparison, mammary glands from control dams on PND10 were well differentiated, full of 

alveoli filled with milk, and contained few apoptotic bodies and little adipose tissue. At PND20, 

the mammary glands of all PFOA-treated dams were similar in appearance to the mammary 

glands of control dams at PND10. Mammary gland differentiation in dams treated on GD12-17 

was not statistically different from control dams. 

The pups were also impacted by their in utero PFOA exposure. Their mammary glands 

exhibited significantly stunted epithelial branching and longitudinal growth at PND10 and 20. 

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, effects were observed on the dam 

and pup mammary gland, decreased pup body weights and decreased survival for the pups 

exposed diring GD 1-17. 

In the study by Wolf et al. (2007), CD-1 mice were orally dosed with 0 or 5 mg PFOA/kg 

on GD7-17 (n=14), 10-17 (n=14), 13-17 (n=12), and 15-17 (n=12) or with 20 mg/kg on GD15-

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 PND22, 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 4th and 5th inguinal mammary glands were removed from female 

offspring and analyzed at various intervals up to 18 months of age (White et al., 2009). The 

remaining offspring were observed until PND189. 

Maternal weight gain was increased in dams exposed to PFOA beginning on GD7 and 

10, but there was no affect 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 on GD7-17 or 10-17 or with 20 mg/kg on GD15-17. By PND78, all male offspring 

had recovered to control body weight levels. On PND161, the offspring of dams dosed during 

GD13-17 weighed significantly more than control. Litters exposed to 20 mg/kg on GD15-17 

experienced decreased survival (p<0.01) during postnatal days 1 through 22. Maternal relative 

liver weight was significantly increased in all PFOA-treated dams except those treated during 

GD15-17. 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 GD7-17 and 10-

17 (Wolf et al., 2007). 
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Mammary gland developmental scores for all offspring of dams exposed to PFOA were 

significantly lower compared to scores from offspring of control dams at PND29 and 32. 

Delayed ductal elongation and branching and delayed appearance of terminal end buds were 

characteristics of delayed mammary gland development at PND32. At 18 months of age, 

increases in epithelial density and in density of darkly staining foci were observed in offspring of 

PFOA-exposed dams. Based on histopathologic observation the darkly staining foci were 

hypothesized to be hyperplasia of ductal epithelium, inflammatory cell influx into ductal regions, 

increased stromal density, or altered differentiation of ductal epithelia. The average number of 

foci per mammary gland was 1.5, 29.8, 17.9, 32.8, and 25.5 for the control, GD15-17, GD13-17, 

GD10-17, and GD7-17 groups, respectively (White et al., 2009). The 5 mg/kg 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 dose was 

associated with decrease postnatal pup survival.

The objective of a second component of the Wolf et al. (2007)/White et al. (2009) study 

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), or 5 (n=36) mg PFOA/kg on GD1 through 17 and their offspring crossfostered at birth to create 7 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 PND22, 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 4th and 5th inguinal mammary glands were removed from female offspring 

and analyzed at various intervals up to 18 months of age (White et al., 2009). The remaining 

offspring were observed until PND245. 

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, but no differences in 

the number of implantation sites were observed between the treated and control mice. Absolute 

and relative livers 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 (PND22). No difference in the number of live pups born per litter were found 

between treated and control mice, but male and female pup birth weight was reduced (p<0.01) in 

dams receiving 5 mg/kg (Wolf et al., 2007). The 3 mg/kg dose was a LOAEL for increases in 

liver weight in the dams while 5 mg/kg 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 from maternal 

grooming behavior which 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 PND2 and 1, respectively, and remained reduced throughout the lactation period. 
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Body weight recovery to control levels was reached by male offspring within two weeks of 

weaning, but recovery in female offspring in the 5U and 5U+L groups did not occur until after 

PND85.

Postnatal survival in 5U+L pups was significantly decreased compared to control survival 

beginning at PND4 and continuing throughout lactation. Survival in the other groups was not 

different from 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 PND22 except for females in the 3L group. At PND42, mammary gland 

scores from females in the 3U+L group were the only ones not statistically different from control 

scores. This may have been due to inter-individual variance and multiple criteria used to 

calculate mammary gland development scores. Despite the fact that the mammary gland scores 

from the 3U+L group were not significantly different from scores of the control animals, they did 

exhibit reduced branching and mammary fat pads with reduced parenchymal density. All 

offspring of dams exposed to PFOA exhibited delayed mammary gland development at PND63 

including those exposed only through lactation (3L and 5L). A 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) or 5 (n=56) mg PFOA/kg from GD8 through 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 4 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 PND1, 3, 5, and 10. Blood and liver samples were collected 

for PFOA analysis. The 4th and 5th inguinal mammary glands were collected from dams and 

female offspring and analyzed. 

Maternal weight gain in treated dams was significantly higher compared to control 

weight gain, but there were no effects of treatment on litter size or pup birth weight or weight at 

PND1. Significantly decreased body weight occurred in the pups of the 5U+L group on PND3 

and in all PFOA-exposed pups on PND 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 PND1, 

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 PND5; serum PFOA levels were approximately 

15,000 ng/mL (White et al., 2009). 

On PND1, the mammary glands of PFOA-exposed dams were similar to glands seen in 

late pregnancy, prior to parturition. In control dams nursing offspring from PFOA-exposed 

dams, reduced alveolar filling was noted as early as PND3 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

PND10 (terminal sacrifice). Delayed mammary gland development was observed as early as 
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PND1 in all female offspring from PFOA-exposed dams including those exposed through 

lactation only (5L). Delayed mammary gland development persisted through the study duration 

(White et al., 2009). 

Macon et al. (2011) gavage dosed CD-1 mice with 0, 0.3, 1.0, or 3.0 mg PFOA/kg from 

GD1 to GD17 (n=13 dams/group). Offspring were sacrificed on PND 7, 14, 21, 28, 42, 63, and 

84, and blood, liver, brain, and the fourth and fifth mammary glands were collected from female 

pups. 

Body weight in male and female offspring was not affected through PND 84. Absolute 

liver weight was significantly increased at ≥0.3 mg/kg in females and at ≥1.0 mg/kg in males on 

PND 7, and at 3.0 mg/kg in females at PND14. Relative liver weight was significantly increased 

at ≥0.3 mg/kg in males and females on PND7, at ≥1.0 mg/kg in females on PND14, and at 3.0 

mg/kg in males and females on PNDs 14, 21, and 28. No dose-related differences were observed 

absolute and relative brain weights.

Delayed mammary gland development of female pups was observed as early as PND7 at 

≥1.0 mg/kg and PND14 at ≥0.3 mg/kg and persisted until the end of the study. The delayed 

development was characterized by reduced epithelial growth and the presence of numerous 

terminal end buds at PND 63 and 84. The LOAEL was 0.30 mg/kg based on significantly 

increase liver weight and delayed mammary gland development. No NOAEL was established.

Macon et al. (2011) also gavage dosed CD-1 mice with 0, 0.01, 0.1, or 1.0 mg PFOA/kg 

from GD10 to GD17 (n=5-8 dams/group) to examine the effects of low doses of PFOA on 

mammary gland development. Female offspring were sacrificed on PND1, 4, 7, 14 and 21, and 

blood, liver, and the fourth and fifth mammary glands were collected. 

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 was also significantly increased at PND14. Mammary gland 

development was delayed by exposure to PFOA, especially longitudinal epithelial growth. At 

PND21, 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. 

The LOAEL was 0.01 mg PFOA/kg based on delayed mammary gland development. No 

NOAEL was established.

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/group) were gavage dosed with 0, 1, or 5 mg PFOA/kg from GD1-17. A separate 

group of pregnant mice (n=7-10 dams/group) were gavage dosed with either 0 or 1 mg PFOA/kg 

from GD1-17 and received drinking water containing 5 ppb PFOA beginning on GD7. 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 breeding and early gestation, to 

simulate a chronic low-dose exposure. F0 females were sacrificed on PND22. F1 offspring 

were weaned on PND22 and bred at 7-8 weeks of age. F2 litters were maintained through 

PND63. Groups of F1 and F2 offspring (n=1-2 offspring/litter from 5-7 litters/group) were 

sacrificed on PND 22, 42, and 63. A group of F2 offspring (n=6-10/group) was also sacrificed 
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on PND10. A lactational challenge experiment was performed on PND10 with F1 dams and F2 

offspring. 

F0 Females. Gestational exposure to 5 mg PFOA/kg significantly increased prenatal loss in F0 

mice and significantly decreased the number of live offspring and the postnatal survival of the 

viable pups. Weight gain and number of implants did not differ among the groups. At PND 22, 

control F0 dams displayed weaning-induced mammary involution. At PND22, 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 (~PND10). 

F1 Generation. Offspring body weight at PND42 was significantly reduced for those whose 

dams received 5 mg/kg; at PND63, body weight was significantly reduced at 1 mg/kg + 5 ppb in 

the drinking water. Liver-to-body weight ratios were significantly increased at 1 mg/kg on 

PND22 and at 5 mg/kg on PNDs 22 and 42. 

F1 Females. There was no indication of toxicity in F1 females. Exposure to PFOA did not 

affect prenatal loss or postnatal survival, although F1 females that had been gestationally 

exposed to 5 mg/kg, had significantly fewer number of implants. On PNDs 22, 42, and 63, the 

mammary glands of all mice exposed to PFOA were significantly delayed in development. F1 

females were bred and examined on LD10 and 22. On LD10, delayed mammary gland 

development (lactation morphology) was observed in all groups. Although, morphological 

differences were observed, there were no differences among the groups in the lactational 

challenge experiment. 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 minutes. 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. By LD22, only F1 dams in the 1 

mg/kg + 5 ppb PFOA and 5 mg/kg groups displayed delayed mammary gland development. 

F2 Generation. A significant reduction in body weight was observed in control + 5 ppb drinking 

water PFOA offspring on PND 1, but there was no difference by PND 3. Offspring from the 1 

mg/kg + 5 ppb drinking water PFOA group had a significant increase in weight compared to 

controls on PNDs 14, 17, and 22. Liver-to-body weight ratios were not different among the 

groups. Mammary gland development in groups with 5 ppb PFOA in the drinking water was 

significantly delayed at PND42. Mammary glands in the remaining groups displayed delayed 

development, but control F2 females had unusually low scores at PNDs 10 and 22 which may 

have reduced the ability to detect statistically significant effects in other treatment groups.

In summary, this study showed delay mammary gland development in dams and female 

pups exposed continuously to a concentration as low as 5 ppb in the drinking water for three 

generations. Direct dosing of dams with 1 or 5 mg/kg during gestation also resulted in delayed 

mammary gland development in dams and their offspring at both doses as well as reduced pup 

survival and decreased body weight at 5 mg/kg. 

Direct peripubertal exposures

Yang et al. (2009) gavage dosed 21 day old female BALB/c mice (5/group) with 0, 1, 5, 

or 10 mg PFOA/kg for 5 days/week for 4 weeks to determine the effects of peripubertal PFOA 
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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. The mammary glands of female BALB/c mice 

treated with 5 or 10 mg/kg 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 (not tested at 10 

mg/kg). 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, and did not occur at 5 or 10 mg/kg. Under the conditions of this study, the 

LOAEL was 1 mg/kg/day based on delayed vaginal opening, increased liver weight and 

decreased uterine weight, and no NOAEL was established.

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 causing 

significantly reduced body weight. At 5 mg/kg, 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 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 and significantly decreased in C57BL/6 mice dosed with 10 mg/kg. 

There was no difference in uterine weights between mice treated with 5 mg/kg and control mice. 

Vaginal opening was delayed in C57BL/6 mice dosed with 5 mg/kg and did not occur in mice 

dosed with 10 mg/kg. Under the conditions of this study, the LOAEL was 1 mg/kg/day based on 

and increased liver and uterine weights, and no NOAEL was established.

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 Yang et al. (2009). In experiments to determine if 

PFOA has a hormonal effect on mammary gland development, C57BL/6 mice (n=10/group) 

were ovariectomized (OVX) at 3 weeks of age, allowed one week to recover, and treated with 0 

or 5 mg PFOA/kg bw for 4 weeks. Abdominal and inguinal mammary glands were collected at 

sacrifice, prepared as wholemounts, 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 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 PFOA for 4 weeks starting at 

21 days of age. After the last dose, the mice were ovariectomized, allowed to recover for 1 

week, and injected s.c. for 5 days with 17β-estradiol (E, 1 μg/0.2 ml/mouse), progesterone (P, 1 

mg/0.2 ml/mouse), or both (E+P, 1 μg+1 mg/0.2 ml/mouse). The mice were sacrificed 24 hours 

after the last hormone injection. Abdominal and inguinal mammary glands were collected at 
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sacrifice, prepared as whole mounts, and scored for growth and development. In the mammary 

glands of mice treated with PFOA and E, stimulated terminal ducts were observed, and in 

PFOA-treated 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 progesterone.

In experiments to determine if PFOA-induced mammary gland development stimulation 

is related to PPARα expression and to determine the impact of PFOA on steroid hormones and 

growth factors, female C57BL/6 and PPARα-null C57BL/6 mice (n=5-10 mice/group) were 

gavage dosed with 0 or 5 mg/kg bw PFOA 5 days/weeks for 4 weeks starting at 21 days of age 

(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 gland were used in 

western blot analysis of several enzymes, local growth factors, and receptors. Among those were 

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 

progesterone and androstenediole to testosterone. Growth factors critically involved in 

mammary gland development including amphiregulin (Areg), insulin like growth factor I (IGFI), and hepatocyte growth factor (HGFα), and markers of cell proliferation including cyclin D1 

and proliferating cell nuclear antigen (PCNA) were analyzed by western blot. Areg mediates 

estrogen receptor α (ERα) function and is a ligand for epidermal growth factor receptor (EGFR). 

These receptors were also 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 

PFOA-treated 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 

progesterone 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 sex hormone 

binding globulin and albumin levels were not significantly changed by treatment with PFOA. 

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 
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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 may increase serum steroid hormone levels in the ovaries.

Due to the increased progesterone levels observed in PFOA-treated mice, the expression 

of liver metabolic enzymes was analyzed. Liver metabolic function may 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 but not limited to CYP1a1, CYP1a2, and HSD17β2, was significantly down-regulated 

(p<0.05) in C57BL/6 mice treated with PFOA, but expression in PFOA-treated PPARα-null 

mice was comparable to 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 may 

not contribute to PFOA-induced mammay gland development stimulation. 

Inhalation Exposure

There were two components (inhalation and oral) to the Staples et al. (1984) 

developmental toxicity study of PFOA in Sprague-Dawley rats. The oral component was 

presented earlier. The study design consisted of whole-body dust inhalation exposure for 6 

hours/day, on GD 6-15. The MMAD ranged from 1.4 to 3.4 μm and the geometric standard 

deviation (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 GD21 prior to 

parturition, and in experiment 2, the dams were allowed to litter and were sacrificed on PND23; 

offspring were sacrificed on PND35. In the first trial (both experiment), dams (n=12) were 

exposed to 0, 0.1, 1, or 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/group and two additional groups (6 dams/group) 

were added and were pair-fed to the 10 and 25 mg/m3 groups. In experiment 2 of trial 2, only 6 

control and 6 dams dosed at 10 mg/m3 were allowed to litter.

In the first experiment, the dams were weighed on GD 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 as well as macro- and microscopic evaluation 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 GD 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 were also observed at these concentrations, with statistical significance at the high-
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concentration only. Likewise, statistically significant increases in mean liver weights (p < 0.05) 

were seen at the high-concentration group. Under the conditions of the study, a NOAEL and 

LOAEL for maternal toxicity of 1 and 10 mg/m3

, respectively, were indicated. 

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. Under the conditions 

of the study, a NOAEL and LOAEL for developmental toxicity of 10 and 25 mg/m3, 

respectively, were indicated.

In the second experiment, in which the dams were allowed to litter, the procedure was the 

same as that for the first experiment 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 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. 

In the second experiment, clinical signs of maternal toxicity seen at 10 and 25 mg/m3

were similar in type and incidence to those described for trial one. Maternal body weight gain 

during treatment at 25 mg/m3 was less than controls, although the difference was not statistically 

significant. In addition, 2 out of 12 dams died during treatment at 25 mg/m3

. No other treatmentrelated effects were reported, nor were any adverse effects noted for any of the measurements of 

reproductive performance. Under the conditions of the study, a NOAEL and LOAEL for 

maternal toxicity of 1 and 10 mg/m3, respectively, were indicated. 

 

Signs of developmental toxicity in this group consisted of statistically significant 

reductions in pup body weight on PND1 (6.1 g at 25 mg/m3 vs. 6.8 g in controls, p=0.02). On 

PNDs 4 and 22, pup body weights 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. Under the 

conditions of the study, a NOAEL and LOAEL for developmental toxicity of 10 and 25 mg/m3

, 

respectively, were indicated.

Dermal Exposure

No data on the developmental effects of dermal exposures to PFOA were identified in the 

literature. 

4.2.6 Chronic Toxicity 

Oral Exposure 

Monkey. Male Cynomolgus monkeys (n=4 or 6/dose) were administered PFOA by oral capsule 

containing 0, 3, 10, or 30/20 mg/kg/day for 26 weeks (Thomford, 2001b, Butenhoff et al., 2002). 
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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 

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 (estrone, estradiol, estriol, testosterone, TSH, free T4, total T4 

and CCK) were not significantly altered throughout the study. However, free 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 4-13. The 

cause of the increase in liver weights was suggested to be due to 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. 

TABLE 4-13. 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*

From Butenhoff et al. 2002

*Significantly different from control, p<0.01.

Because administration of PFOA to rats has been shown to result in liver, Leydig cell and 

pancreatic acinar cell tumors, Butenhoff et al. (2002) analyzed markers of tumor formation in the 

monkey study just described. In the liver, a two-fold 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 has also been proposed 

that changes associated with the pancreatic acinar cell tumors in rats include increased serum 
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CCK concentrations and indications of cholestasis, including increases in alkaline phosphatase, 

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 is associated with Leydig cell tumors in rats. There were no changes in 

replicative DNA synthesis in the pancreas or testes. 

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. Under the conditions of the study, the LOAEL was 3 mg/kgday based on increased liver weight and a NOAEL was not established.

Rat. The chronic toxicity of PFOA was investigated in a 2 year study in rats by Butenhoff et al. 

(2012). Sprague-Dawley (Crl:CD BR) rats (50/sex) were fed diets containing 0, 30 or 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 sex were fed 0 or 300 ppm PFOA and evaluated at the one year interim 

sacrifice. All animals were observed daily throughout the two 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 

one and two years of dosing. Organ weights were determined after each sacrifice and the tissues 

subjected to histological examination.

There were dose-related decreases in body weight gains in male and female rats as 

compared to the controls; statistical significance was attained for only the high-dose group (both 

sexes). The body weight changes were considered treatment related since feed consumption 

increased (rather than decreased) during the study. There were no significant differences in 

mortality between the treated and untreated groups. Significant decreases in red blood cell 

counts, hemoglobin concentrations and hematocrit values were observed in the high-dose male 

and female rats as compared to control values. Clinical chemistry changes included slight (less 

than 2-fold) but significant increases in ALT, AST, and AP in both treated male groups from 3-

18 months, but only in the high-dose males at 24 months. No dose- or treatment-related 

differences in absolute and relative organ weights were found between the treated and control 

groups at 2 years. 

Significantly increased incidences of lesions in the liver were observed in the high-dose 

male group. At 1 year, diffuse hepatomegalocytosis, portal mononuclear cell infiltration and 

hepatocellular necrosis were seen. At 2-years, significant increases in megalocytosis 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, and 

hyperplastic nodules, a localized proliferation of hepatic parenchymal cells, were also 

significantly increased in high-dose males.

Among the high dose males histological changes were noted in tissues other than the 

liver. A significant increase in aspermatogenesis and vascular mineralization of the testes and the 

incidence of pulmonary hemorrhage (44%) and alveolar macrophates (62%) were reported. 

The LOAEL for male rats is 300 ppm (14.2 mg/kg/day) based on a decrease in body 
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weight gain and histological changes in the liver, testes, and lungs. The LOAEL for female rats 

is 300 ppm (16.1 mg/kg/day) based on decreased body weight gain and hematologic effects. The 

NOAEL for both sexes is 30 ppm (1.3 mg/kg/day 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/group) were fed a diet containing 0 or 300 ppm PFOA (0 or 13.6 mg/kg/day). 

Interim sacrifices were conducted every three 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.

 

Body weight was significantly decreased from days 8 to 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% vs. 14% control group) 

Pancreatic aceinar 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, 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. 

4.2.7 Carcinogenicity 

Oral Exposure 

Tissues from the animals in the Butenhoff et al. (2012) study were evaluated for 

neoplastic and preneoplastic formations. Hepatocelluar carcinomas were observed at 6, 2, and 

10% 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% incidence was observed for female rats in the highdose group. The differences between the treated and control groups were not significantly 

different. No liver adenomas were observed. 

At the one-year sacrifice, testicular masses were found in 6/50 high-dose and 1/50 lowdose 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 Leydig cell tumor incidence in the control, low-, and high-dose groups was 0/50 (0%), 2/50 

(4%), and 7/50 (14%), respectively. The increase was also 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). The spontaneous incidence 

of Leydig cell tumors in 2-year old Sprague-Dawley rats in other studies was reported to be 

approximately 5% (cited in Clegg et al., 1997). 

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 
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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. (2012) study were re-evaluated 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 4-14. 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 incidences in the control group than in either of the treated groups. 

Results of this follow-up evaluation indicated that rats sacrificed at the one-year interim 

sacrifice, as well as rats that died prior to the interim sacrifice were not considered at risk for 

tumor development. 

 

TABLE 4-14. 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

From Mann and Frame, 2004

A significant increase (p<0.05) in the incidence of mammary fibroadenomas in the low 

and high dose groups of female rats was noted in the Butenhoff et al. (2012) study. The 

incidence of mammary fibroadenomas was 22% (11/50), 38% (19/50) and 42% (21/50) in the 

control, 30, and 300 ppm groups, respectively (Hardisty, 2010). The increase was also 

statistically significant when compared to the historical control incidence of 19.0% observed in 

1,329 Sprague-Dawley control female rats used in 17 carcinogenicity studies as reported by 

Chandra et al. (1992). Butenhoff et al. (2012) did not consider the mammary fibroadenomas to 

be treatment related on the basis of the historical control incidence (24%) from a study of 181 

female rats terminally sacrificed at 18 months (Prejean et al., 1973). However, this is an 

inappropriate historical reference because the study duration was 18 months instead of 2 years. 

When the mammary fibroadenoma incidences were compared to the historical control incidence 

(37%) in 947 female rats in the Haskell Laboratory, however, there did not appear to be any 

compound related effect (Sykes, 1987). The incidence of other mammary gland tumor types is 

provided in Table 4-15. 

The mammary gland findings were re-examined by a Pathology Working Group 

(Hardisty et al., 2005, 2010; Table 4-15) using the diagnostic criteria and nomenclature of the 

Society of Toxicological Pathologists (Mann et al., 1996). The Pathology Working Group 

(PWG) concluded that there were no statistically significant differences in the incidence of 

fibroadenoma, adenocarcinoma, total benign neoplasms (lobular hyperplasia, fibroadenoma, and 
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adenoma) or total malignant neoplasms (adenocarcinoma) of the mammary glands between 

control and treated animals using Fischer’s Exact Test for pair-wise comparision (see Table 4-

15). There was also no significant difference in combined benign and malignant neoplasms 

between control and treated groups. 

The primary difference between the original reported findings and the PWG results 

involved findings initially reported as lobular hyperplasia being reclassified as fibroadenomas, 

mostly in the control group. With the reclassification, the incidence of single mammary 

fibroadenoma in the control, low- and high-dose groups were: 32% (16/50), 32% (16/50), and 

40% (20/50), respectively. The evidence for mammary fibroadenomas in the female rat is 

equivocal since the incidences were comparable to some historical background incidences, so the 

PWG concluded that there were no statistically significant differences in the incidence of 

fibroadenoma or other neoplasms of the mammary gland between control and treated animals 

(Hardisty et al., 2005, 2010). 

TABLE 4-15. 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%)

Fibroadenoma* 

(%)

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%)

From Hardisty et al., 2005, 2010 (Compares Butenhoff et al., 2012 with Hardisty et al., 2005, 2010) 

*Includes fibroadenoma, multiple counts 

A 2-year mechanistic study in male Crl:CD BR (CD) rats (Cook et al., 1994; Biegel et al., 

2001) resulted in Leydig cell tumors supporting the observations from the Butenhoff et al. (2012) 

study. The rats (n=156/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. 

 

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). The incidence in ad 

libitum control rats was 0% (0/80). In addition, the treated group had a significant increase in the 

incidences of liver adenomas and pancreatic acinar cell adenomas when compared to the pair-fed 

and ad libitum control groups. 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. 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. The incidence for the pancreatic 

acinar cell adenomas was 0% (0/80), 1% (1/79), and 9% (7/76). 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 
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0% (0/80) in control rats.

 

In the first carcinogenicity study (Butenhoff et al., 2012), there was no reported increase 

in the incidence of pancreatic acinar cell tumors, and 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 incidences 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 qualitiative; 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. (2001) study compared to the Butenhoff et al. (2012) study. 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 to be due most likely to difference in the diets used in the two 

laboratories (Frame and McConnell, 2003).

Inhalation Exposure

 No data on the tumorigenic effects of chronic inhalation exposures to PFOA were 

identified in the literature.

Dermal Exposure

No data on the tumorigenic effects of chronic dermal exposures to PFOA were identified 

in the literature.

4.3 Other Key Data 

4.3.1 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 4-16. 

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 & 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, 1996c; NOTOX, 

2000). Sadhu (2002) reported that PFOA did not induce gene mutation when tested with or 
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without metabolic activation in the K-1 line of Chinese hamster ovary (CHO) cells in culture.

Murli (1996b,d) 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, and TA104 (Freire et al., 2008). 

 

TABLE 4-16. Genotoxicity of PFOA In Vitro

Test System End-point With Activation Without Activation Reference

C3H10T1⁄2 mouse 

embryo fibroblasts

Cell Transformation NA - Garry & Nelson, 

1981

C3H 10T1⁄2 mouse 

embryo fibroblasts

Cytotoxicity NA - Garry & Nelson, 

1981

S. typhimurium

TA1537

Gene Mutation - + 

(not reproducible)

Lawlor, 1995; 1996

E. coli Gene Mutation - - Lawlor, 1995; 1996

Chinese hamster 

ovary cells

Chromosomal 

Aberrations +, + +, - Murli, 1996b, 1996d

Chinese hamster 

ovary cells

Polyploidy +, + +, - Murli, 1996b, 1996d

Human lymphocytes Chromosomal 

Aberrations - - Murli, 1996c; NOTOX, 2000

K-1 Chinese hamster 

ovary cells

Gene Mutation - - Sadhu, 2002

S. typhimurium

TA98, TA100, 

TA102, TA104

Gene Mutation

- - Freire et al., 2008

NA= Not applicable

The data summarized above 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. 

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, 1996a). 

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 is encodes for expression of 

the human cell surface protein CD59. AL and mitochondria-deficient AL cells were incubated 

with 0, 1, 10, 100, or 200 μM 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 
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oxygen species, nitric oxide, and superoxide anion production were measured in the cells, and 

the effects of reactive oxygen species/reactive nitrogen species quenchers [0.5% dimethyl 

sulfoxide (DMSO) and 0.2 mM NG-methyl-L-arginine, respectively] on mutagenicity and 

caspace acitivities were determined. At 100 and 200 μM 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 μM PFOA. There was no increase 

in mutations in mitochondria-deficient AL cells after incubation with 100 or 200 μM PFOA. 

Production of reactive oxygen species, nitric oxide, and superoxide anion was 

significantly increased at 100 and 200 μM PFOA after incubation for 1, 4, and 16 days. 

Inhibition of reactive oxygen species through incubation with DMSO significantly decreased the 

CD59 mutation frequency caused by 200 μM PFOA after a 16 day incubation. Incubation with 

100 or 200 μM PFOA did not increase reactive oxygen species or superoxide anion production in 

mitochondria-deficient AL cells. The highest dose 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-Larginine significantly decreased the increased caspace activity induced by 200 μM PFOA. The 

results led the authors to suggest that mitochondrial-dependent reactive oxygen species may play 

an important role in PFOA-induced mutagenicity and induction of caspase activities may be 

mediated by reactive oxygen and nitrogen species. 

4.3.2 Immunotoxicity 

Yang et al. (2000, 2001, 2002a,b) 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, or 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 two. Following five days of administration, significant decreases in 

thymus and spleen weight were noted. 

A second component of the Yang et al. (2000) study 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 

must 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 down regulation 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 numbers 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+ 
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surface protein marking them as T-cells for exportation to the spleen. There are several classes 

of the 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. (2000) study, 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 (50, 100, or 200 μM) of PFOA. However, under the exposure 

conditions PFOA did not appear to significantly alter the cell cycle. 

 In a later publication Yang et al. (2001) reported on their examination of the 

immunosuppressive effects of PFOA. As was the case with 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 to 10 days. One group of animals was exposed to PFOA each 

day until day of sacrifice on days 2, 5, 7, or 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 to 10 days in order to monitor 

recovery from the effects of exposure. In the recovery group, animals were sacrificed after 2, 5, 

or 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 two days of 

the study and decreased continuously for the remainder of the exposure period. The activity of 

palmitoyl-CoA and lauroyl-CoA, biomarkers for PPARα activation and and peroxisome 

proliferation, were also increased significantly and increasingly across the exposure period. The 

impact of PFOA exposure was similar to that observed in the Yang et al. (2000) study. 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 
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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. (2001) study, 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 dose-related increase in liver weights accompanied by a corresponding 

increase of pamitoyl Co-A and lauroyl Co-A activity. Enzyme activity was significantly 

increased for all doses. Spleen and thymus weights were significantly decreased at doses ≥ 

0.01% and higher but not for the lower doses; the increases in liver weights were significantly 

increased for the 0.02% and 0.05% doses. 

 A Yang et al. (2002a) study was designed to examine 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.

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 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 PPARα-null control mice, but it was not as dramatic as that in the PFOA-exposed 

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. 

 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 
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10 days. The 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 were also performed. The results 

showed that mice fed normal chow responded to challenge with HRBCs with a strong humoral 

response, 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 ConA 

and LPS with lymphocyte proliferation, as measured by thymidine incorporation. However, 

treating mice with PFOA (0.02% for 7 days) attenuated the proliferation. In a set of in vitro 

experiments, PFOA (1- 200 μM) 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 and cellular immunity. The first component of 

their publication had many similarities with the Yang et al. (2001) study. Adult female 

C57BL/6J mice (n=40/endpoint and 8/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 Day 15 (constant 

group) while the other half received distilled water from Day 11 through Day 15 (recovery

group). On Day 11, the mice were immunized with sheep red blood cells (SRBC). Sacrifices 

took place on Day 16 (1 day post-exposure period) or Day 31 (15 days post-exposure period). 

Vehicle and cage controls were also 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 bovine serum album (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. 

(2001) study. 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 

PFOA-treated 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 post exposure day 1 in both the recovery and constant exposure groups. 

There were no significant differences from controls for IgG levels and DTH foot-pad responses 

to the BSA challenge. 
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 The C57BL/6 mice used for the continuous-dosing vs. recovery component of the DeWitt 

et al. (2008) study 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, or 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, or 7.5 mg PFOA/kg/day administered for 15 days in the drinking water. The 

immunological sensitization and post-dose monitoring were identical to that used in the constantdosing vs. 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 was significantly reduced 

at ≥3.75 mg/kg PFOA in a direct dose-related manner. The IgG response 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 

levels on Day 16, and increased IgG levels on Day 31. The immunological NOAEL was 1.88 

mg/kg/day. Benchmark dose analysis of IgM serum titer data gave a lower bound 95% 

confidence limit of 1.75 mg/kg/day on a benchmark dose (one standard deviation) 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. 

 

 Loveless et al. (2008) administered 0, 0.3, 1, 10, or 30 mg linear PFOA/kg by oral gavage 

to groups of male CD rats (n=10/group) and CD-1 mice (n=20/group) for 29 days. Both species 

received a dose of SRBC on day 23 (rat) or 24 (mice). A separate group of high dose rats and 

mice were injected with water instead of SRBC. Rat body weight was recorded on days 0, 3, and 

6-28, while mice were weighed daily. At necropsy blood was collected for immune parameters. 

Cell counts were determined for the thymus and 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 popliteal lymph nodes revealed no effects in treated rats 

resulting from PFOA exposure. There was no difference observed in IgM titers between treated 

and control rats. The immunological NOAEL was 30 mg/kg/day. 

In mice, 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 was also 
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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 

corticosterone was significantly increased at 10 mg/kg/day and elevated (not statistically 

significant) at 30 mg/kg/day. When IgM and corticosterone were plotted against each other, a 

negative correlation coefficient suggested that increasing corticosterone 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. 

 

 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 corticosterone 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, or 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 post 

dosing. 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 

(AP, AST, ALT, LDH, GGT, and SDH), serum lipids (cholesterol and triglycerides) and 

corticosterone. A baseline measure of corticosterone 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 corticosterone and IgM were 

determined. 

 Body weight in the sham-operated mice declined during dosing in the highest dose group 

but recovered by 5 days post dosing. 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 

post dosing. 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 4-17).

TABLE 4-17. 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*

From DeWitt et al., 2009

*=p<0.05 vs corresponding sham control or ADX control group

Serum levels of triglycerides significantly decreased compared to controls with all doses 

for the sham operated mice on day five 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. 
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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 4-18). 

TABLE 4-18. 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

From 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 post dosing the corticosterone 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 stressrelated corticosterone production did not have a major impact on the IgM response to the SRBC 

inoculation. 

 Son et al. (2009) administered 0, 2, 10, 50, or 250 ppm PFOA (0, 0.49, 2.64, 17.63, or 

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 interleukin-6, and the protooncogene 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 250 ppm in the spleen. Interleukin-1β expression was also elevated at 50 ppm in the 

spleen. In the thymus, c-myc expression was significantly elevated by treatment with 50 and 25 

ppm PFOA.
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 The splenic and thymic lymphocyte population was altered by PFOA treatment as shown 

in Table 4-19. A 50% decrease was observed in splenic CD8+

 lymphocytes at all PFOA doses, 

and an increase 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 by PFOA. 

TABLE 4-19. 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+ - - ↓ ↓

From Son et al., 2009

↑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% or 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 a lipopolysaccharide (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 may 

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.

In a pilot study, Brieger et al. (2011) examined the effects of PFOA on human 
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leukocytes. Peripheral blood mononuclear cells (PBMC) were obtained from the blood of 11 

voluntary donors (n=6 females, 5 males). PBMC were incubated with varying concentrations of 

PFOA followed by assays for cell viability, proliferation, and natural killer (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 PBMC were incubated with 0-500 μg/mL for 24, 48, or 

72 hours, and HL-60 cells were incubated with 0-125 μg/mL PFOA for 24 hours. 

• In the proliferation assay, the PBMC were incubated with 0-100 μg/mL PFOA for 24 

hours, labeled with 6-carboxyfluorescein succinimidyl ester (CFSE), stimulated with 

concanavalinA, a T-cell mitogen (ConA, 5 μg/mL to half of all samples), and 

incubated for an additional 72 hours. 

• For the NK assays, PBMC cells 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) labled 

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 nM 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. Lipopolysaccharide 

(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 

9 samples used, cells from 6 donors had slightly increased proliferation and 3 had no response. 

In cells incubated with ConA and 100 μg/mL PFOA, a non-significant 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 priod 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 
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concentrations may show an impact on human immunity with a larger population. 

 

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/group) at doses 

of 0, 0.01%, 0.1%, 0.25%, 0.5%, 1.0%, or 1.5% PFOA (equivalent to 0, 0.25, 2.5, 6.25, 12.5, 25 

or 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/kd 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. (2007) study, groups of 5-15 animals were dosed 

dermally on the ears for 4 days with doses of 0, 0.5%, 0.75%, 1.0%, or 1.5% PFOA (equivalent 

to 0, 12.5, 18.75, 25, or 50 mg/kg/day). On days 1 and 10 they were injected ip with either 2.0 

mg alum or 7.5 μg ovalbumin (OVA) and 2.0 mg alum in 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 (ELISA). 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 37.5 

mg/kg/day highest dose groups was practically indistinguishable from the OVA control. 

The dermal LOAEL was 12.5 mg/kg/day based on increased liver weight and decreased 

spleen weight and spleen cellularity. No NOAEL was established. 
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In part three of the Fairley et al., (2007) study, 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 spleenocytes carrying the B220+ marker (expressed on B 

cells, activated B cells, subsets of T and 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 nebuilized methacholine for 

three minutes followed by two minutes of fresh air. The PenH measures were recorded every 30 

seconds over the next 5 hours. The area under the 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 were also 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. 

4.3.3 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, free 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%), free T4 (~↓80%), and total T3 (~↓25-48%). 

Serum testosterone was significantly decreased (p<0.05, ~↓70%) following a 3 day and 5 day 
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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. 

Reproductive Hormones. Cook et al. (1992) gavage dosed male CD rats (n=15/group) for 14 

days with 0, 1, 10, 25, or 50 mg PFOA/kg to examine the possibility than an endocrine-related 

mechanism may explain Leydig cell adenomas observed in rats. A separate control group was 

pair fed to the 50 mg/kg group. Blood and testicular interstitial fluid were collected at necropsy 

for hormone analysis including testosterone, estradiol, and LH. A separate group of rats was 

dosed with 0 or 50 mg PFOA/kg for 14 days and challenged with 100 IU of human corionic 

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 IU hCG was also anlayzed for progesterone, 17 α-hydroxyprogesterone, and 

androstenedione. 

The relative liver weight at 10 mg PFOA/kg was significantly increased (p<0.05). The 

relative liver weight was significantly increased (p<0.05) and the accessory sex organ unit 

relative weight was significantly decreased (p<0.05) at 25 and 50 mg PFOA/kg compared to 

those weights in control rats. The results at the highest dose were also different when compared 

to the pair fed controls. The relative weight of the ventral prostate was 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 IU 

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 progesterone or 17 α-hydroxyprogesterone between control and 

treated rats. The authors suggested that the decreased serum testosterone levels observed may be 

due to decreased conversion of 17 α-hydroxyprogesterone to androstenedione attributable to the 

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 ± 2 IU 

hCG (final 3 hours) and 0, 100, 200, 250, 500, 700, or 1000 μM PFOA for a total of 5 hours and 

analyzed for testosterone concentration. Leydig cells were also incubated ±500 μM PFOA and 
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anlayzed for testosterone and estradiol at 0, 0.5, 1, 3, 6, 12, 24, and 48 hours. 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. 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. 

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 μM PFOA + hCG, the testosterone concentration 

was significantly increased compared to cells treated with only 100 μM PFOA. Cytotoxicity 

occurred at ≥ 750 μM PFOA. In the time course experiment, 500 μM 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. 

In the in vivo study, 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 PFOAtreated rats compared to the pair fed control rats. In 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 may directly affect Leydig cell function. 

Liu et al. (1996) treated adult male Crl:CD BR (CD) rats (n=15/group) with 0, 0 pair fed, 

0.2, 2, 20, or 40 mg PFOA/kg for 14 days by oral gavage to determine the impact of PFOA on 

aromatase activity. A separate study examined the effect of PFOA on aromatase activity in 

cultured hepatocytes and is discussed below. Aromatase is a cytochrome P450 enzyme localized 

to the endoplasmic reticulum that catalyzes the conversion of androgens to estrogens. 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. 

 In the in vitro study, hepatocytes isolated from control male CD rats were incubated with

0-1000 μM PFOA and the aromatase activity was evaluated after 18, 42, and 66 hours (Liu et al. 

1996). 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 the in vivo study, the body weight of rats treated with ≥ 20 mg PFOA/kg was 

significantly decreased (p<0.5) 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 rat dosed with 40 mg/kg PFOA. Absolute and relative 

liver weight were significantly increased (p<0.5) 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 remaning 
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analysis, the pair-fed control group were similar to the ab 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 2-fold 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 μM. The authors 

estimated the EC50 values for the outcomes and they are shown in Table 4-20. Liu et al. (1996) 

concluded that the PFOA-increased protein yields suggested induction of the endoplasmic 

reticulum 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 2-fold. 

TABLE 4-20. 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

Relatvie liver weight 1.56

Serum estradiol 3.24

Terminal body weight 11.65

From Liu et al. 1996

EC50= half-maximum response 

In a 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/group) were orally dosed with 0, 0 pair fed, 0.2, 2, 20, or 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 were 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 may 

play in animals exposed during gestation compared to during their adult years. Timed-pregnant 

CD-1 mice were gavage dosed in two blocks on gestation days 1-17 but not thereafter. Block 1 

animals were dosed with 0, 1, 3, or 5 mg PFOA/kg, and Block 2 animals were dosed with 0, 

0.01, 0.1, 0.3, 1, or 5 mg PFOA/kg/day. At birth, pups were pooled within each block and dose 

group and randomly redistributed among the dams (10 pups/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 ovariectomized (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 4-21. A group of naive 8-week old adult mice were dosed for 17 

days with 0, 1, or 5 mg PFOA/kg/day to compare the impact of exposure in adult animals to 
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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 4-21. 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

From Hines et al., 2009

 Body weight of offspring born to dams exposed to 5 mg PFOA/kg was significantly 

decreased (p<0.05) on PND1 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 

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 has also been associated with excess 

abdominal fat. Serum leptin levels were also determined as increased leptin levels have been 

associated with a leptin-resistance mechanism of action 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 gestionally-exposed mice. Serum estradiol 

was measured to determine if PFOA impacted hormone levels at 18 months in intact control and 

intact gestionally-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 

consumption between intact control and intact gestionally-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/kd/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. 
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 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 gestationallyexposed 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, and at high doses, decreased 

weight in early and late life, decreased white fat, increased brown fat, and decreased spleen 

weight were observed. 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. 

Adrenal Hormones. Thottassery et al. (1992) exposed intact or adrenalectomized (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), corticosterone (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, vs. 30.5 

pmole CO2/hr/mg protein, p <0.001). Ornithine decarboxylase activity was not different 

between intact rats treated with PFOA and intact rats treated with the vehicle. 
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 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 activity was not increased in rats 

of 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. 

4.3.4 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, or 

10 mg PFOA/kg/day on GD1-17. Dams were sacrificed on GD18, 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, glucose metabolism, proteosome activation, and ketogenesis. Although PPARα is 

known to regulate (at least partly) the expression of genes for the pathways or biological 

functions involved in lipid homeostasis, PFOA may also 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/group) and PPARα-null mice (129S4/SvJae-PPARαtm1Gonz/J, 6-8/group) 

dosed with 0, 1, or 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. Real-time (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 was less at 1 mg/kg than at 3 mg/kg PFOA. However, 85% of the altered genes at 1 

mg/kg PFOA were alterd at 3 mg/kg PFOA. 

 PPARα target genes including 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 
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mice exposed to Wyeth 14,643. Principle components analysis showed that PFOA-treated 

PPARα-null mice shared similarity with PFOA-treated wild-type mice at 3 mg PFOA/kg. 

 

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 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 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 up-regulated in PFOA-treated wild-type 

mice but not in PFOA-treated PPARα-null mice in microarray analysis. Expression of Ehhadh 

was up-regulated 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 PPARαnull mice (strain SV129/C57BL/6) gavage dosed or fed diets containing Wyeth (WY)-14,643 

(PPARα agonist); 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 

PPARα-null 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 WY-14,643 for 12 hours, or 3 or 7 days were compared. Four groups of altered genes 

were identified based on their expression behavior in wild-type and PPARα-null PFOA-exposed 

mice compared to genes from WY-14,643 treated mice. The first group consisted of genes (397) 

regulated by both PFOA and WY-14,643 in wild-type mice. They had a common direction and 

magnitude of change and were characterized as being lipid homeostasis, inflammation, cell 

proliferation, or proteome maintenance genes. Group II consisted of genes in wild-type mice 

(51) regulated solely by PFOA, and 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 

wild-type mice, and many were involved in lipid metabolism. Regulation of these genes was 
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also observed in WY-14,643 wild type mice. Group IV genes (19) were altered by PFOA only 

in PPARα-null mice, and most were xenobiotic metabolizing enzymes. 

 By comparing the gene expression patterns between PFOA and WY-14,643, the authors 

concluded that:

• PPARα is required for a majority of the transcriptional changes observed in the 

mouse liver following PFOA or WY-14,643 exposure 

• PFOA regulates genes that are also regulated by other peroxisome proliferators in a 

PPARα-dependent manner, and 

• PFOA also regulates PPARα-independent genes. 

The PPARα-independent genes included those involved in lipid homeostasis (upregulated), 

amino acid metabolism (down regulated), 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 

PPARα-null 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-thione 

exposed mice (≤0.06 Pearson’s correlation) and PFOA exposed PPARα-null mice. These results 

suggested that some genes altered by PFOA exposure in PPARα-null mice are regulated by CAR 

but not 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, or 200 μM PFOA for 24 

hours. Expression of acyl-CoA oxidase (Acox), Cyp4a1 (rat), Cyp4a11 (human), acyl-CoA 

thioesterase (Cte-rat, Acot1-human), and DNA damage inducible transcript (Ddit3) were 

determined by quantitative RT-PCR. The aforementioned genes are inducible by peroxisome 

proliferators except Ddit3 which is induced in the presence of direct or indirect DNA damage. 

Exposure to ≥ 100 μM 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 μM 

PFOA, and Cte/Acot1 was significantly induced at ≥20 μM PFOA in rat hepatocytes only. 

Expression of Cyp4a1/11 was significantly induced in rat hepatocytes at 5-50 μM and in human 

hepatocytes at 20-50 μM. 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 4.2.2. Male, 8-week old wild-type 

(mPPARα) mice, PPARα-null mice, and hPPARα mice were gavage dosed with 0, 0.1, or 0.3 

mg/kg PFOA (n=4-6/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 
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bifunctional protein [PH], peroxisomal thiolase [PT], very long chain acyl-CoA dehydrogenase 

[VLCAD], medium chain acyl-CoA dehydrogenase [MCAD], and CYP4A10). RXRα forms a 

heterodimer with PPARα controls 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 4-22. 

TABLE 4-22. 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 - +↑ +↑ - -*↓ -*↓ -*↑ -*↓ -*↓

From Nakamura et al., 2009

- 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

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. 

 As expected, mRNA and protein expression of PPARα was absent in PPARα-null mice. 

Treatment with 0.1 or 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 of 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 was also 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 
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significantly higher (p<0.05) in all hPPARα mice compared to mPPARα mice given the same 

concentration of PFOA. Treatment of hPPARα mice with 0.1 or 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. 

 

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 

may 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 may be needed to cause 

activiation 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 or 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 4.2.5. Females were either sacrificed on GD18 (n = 5-8/group) or 

allowed to give birth and then sacrificed, along with their litters (n = 8-14), on PND20. Livers 

from dams, fetuses, and pups were collected for measurement of mRNAs encoding the PPARα 

target genes cytochrome P4504a10 (Cyp4a10) and acyl-CoA oxidase 1 (Acox1), the constitutive 

androstane receptor (CAR) target gene (Cyp2b10), and the pregnane X receptor (PXR) target 

gene (Cyp3a11). 

On GD18, 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 PND20, 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. 

For fetuses on GD18, liver samples from treated groups showed increased expression of 

Acox1 and Cyp4a10 in wild-type and humanized mice. Expression of Cyp2b10 was 

unchangeded following maternal PFOA administration in all three genotypes, while expression 

of Cyp3a11 was increased in humanized fetal liver. On PND20, 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 Pgc-1α –mediated pathway of mitochondrial 

biogenesis: peroxisome proliferator-activated receptor gamma coactivator 1α (Pgc-1α), estrogenrelated receptor α (Errα), nuclear respiratory factors 1 and 2 (Nrf1, Nrf2), transcription factor A 

(Tfam), cytochrome c oxidase subunit II and IV (Cox II, Cox IV), NADH dehydrogenase 2 
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(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 were significantly increased 2 to 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 transcriptioinal level by 

activation of the Pgc-1α pathway. 

 

 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 to PFOA. Treatment with Wyeth 

14,643 caused the induction of CYP4A1 only. 

PPAR Receptor 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, β/δ, or γ 

reporter plasmids and exposed to 0.5-100 μM PFOA or 0.5-100 μM PFOA and MK-886 

(PPARα antagonist) or GW9662 (PPARγ antagonist). An antagonist for PPARβ/δ was not 

available. The three types, PPARα, β/δ or γ, are encoded by different genes, expressed in many 

tissues and have specific roles during development as well as in the adult. The results are shown 

in Table 4-23. PFOA activated mouse and human PPARα in a dose-dependent manner with a 

significant increase in activity observed at 10, 20, 30, and 40 μM in the mouse and 30 and 40 μM 

in the human 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. Activity of mouse PPARβ/δ was significantly increased after exposure to 40-

80 μM 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. 
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TABLE 4-23. 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 - - From Takacs and Abbott, 2007

+ Significant increase in activity between treated and control

- No difference in activity between treated and control

Biomarkers for Peroxisome Proliferation. Pastoor et al. (1987) dosed male Crl:CD (SD) BR 

rats for 1, 3, or 7 days with 0 or 50 mg PFOA/kg. Hepatic DNA content, cytochrome P450 

content, UDP-glucuronyltransferase, glutathione S-transferase, benzphetamine N-demethylase 

activity (marker for smooth endoplasmic reticulum proliferation), and ethoxyresorufin Odeethylase 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 carnitine aceyltransferase (CAT) and carnitine palmitoyltransferase (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 total cholesterol 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 Odeethylase 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 2-fold (p<0.05) 

after a 3-day treatment with 50 mg PFOA/kg. No differences were observed among the groups 

for the other enzymes. No differences were observed between rats treated for 7 days and control 

rats in plasma total cholesterol 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 

endoplasmic reticulum, and peroxisome proliferation pointed to increased liver weight due to 

hepatocyte hypertrophy. 

Cook et al. (1992) examined peroxisome β-oxidation activity in male CD rats gavage 

dosed with for 14 days with 0, 1, 10, 25, or 50 mg PFOA/kg. Peroxisome β-oxidation activity 

was significantly increased at ≥10 mg PFOA/kg compared to the control. At the highest dose, β-
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oxidation activity was significantly different from the pair fed controls. 

Sohlenius et al. (1992) fed male and female Wistar rats diets containing 0 or 0.02% 

PFOA for 7 days and examined hepatic peroxisome proliferation and related parameters. At 

necropsy, liver samples were collected for analysis of catalase, peroxisomal lauroyl-CoA oxidase 

activity, palmitoyl-CoA oxidation, DT-diaphorase, cytosolic epoxide, glutathione transferase 

activity, ω-hydroxylation of lauric acid, and protein concentration. In males, an increase was 

observed in liver weight, liver somatic index, mitochondrial protein content, and palmitoyl CoAoxidation. No effects were observed in the females.

Sohlenius et al. (1992) fed male and female C57Bl/6 mice diets containing 0 or 0.02% 

PFOA for 7 days to examine possible sex differences in hepatic response to PFOA. Two 

additional groups of female mice were fed a diet containing 0 or 0.05% PFOA for 5 or 10 days. 

At necropsy, liver samples were collected for analysis of catalase, peroxisomal lauroyl-CoA 

oxidase activity, palmitoyl-CoA oxidation, DT-diaphorase, cytosolic epoxide hydrolase, 

glutathione transferase activity, ω-hydroxylation of lauric acid, and protein concentration. 

 Liver weight, liver somatic index, mitochondrial protein, palmitoyl-CoA oxidation, and 

lauroyl-CoA oxidase activity were significantly increased (p<0.05) in all treated mice. 

Microsomal protein was significantly increased (p<0.05) in male mice fed diets of 0.02% PFOA 

for 7 days. In male and female mice fed diets of 0.02% PFOA, catalase, DT-diaphorase, ωhydroxylation of lauric acid, cytosolic epoxide hydrolase, and glutathione transferase activities 

were significantly increased (p<0.05) compared to controls. The authors concluded that no sex 

differences were apparent hepatic peroxisome proliferation and related parameters in mice fed 

diets containing PFOA.

Elcombe et al. (2010) examined peroxisome proliferation in the livers of male SpragueDawley rats exposed to 300 ppm PFOA in the diet for 1, 7, or 28 days (described in 7.2.2). 

Analysis of palmitoyl-CoA oxidation showed significant increases (p<0.05) in β-oxidation at all 

time points in PFOA diet-fed rats compared to β-oxidation in control rats. 

Cellular Differentiation. 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, or 250 μM 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, and differentiating cells were 

evaluated after 4-6 days of exposure to PFOA. The results are shown in Table 4-24. 

 Significant inhibition (p<0.0001) of DNA synthesis occurred in the undifferentiated cells 

after exposure to 250 μM PFOA with no change in DNA content. Lipid peroxidation was 

significantly increased (p<0.02) after exposure to 10 μM PFOA, and cell viability was 

significantly decreased (p<0.03) after 24 hour exposure to 100 or 250 μM PFOA. 
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 In differentiating PC12 cells, exposure to 250 μM 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 (p<0.007) in lipid 

peroxidation, but no effect was observed in cell viability. Tyrosine hydroxylase activity was 

decreased (p<0.05) after exposure to 10 and 250 μM PFOA, and the TH/ChAT ratio was 

decreased (p<0.05) at 10 μM PFOA. The results suggest that PFOA exposure caused the 

differentiating cells to shift (slightly) to favor the acetylcholine phenotype. 

 

TABLE 4-24. Impact of PFOA exposure on PC12 cells

Undifferentiated PC12 

Cells

PFOA (μM)

0 10 50 100 250

DNA Synthesis - - - - ↓

DNA Content - - - - -

Lipid Peroxidation - ↑ - - -

Cell Viability - - - ↓ ↓

Differentiating PC12 

Cells

Lipid Peroxidation - ↑ - - ↑

Cell Viability - - - - -

DNA Content - - - - ↓

Total protein/ DNA content 

ratio

- - - - -

Membrane/Total protein - - - - -

TH activity - ↓ - - ↓

ChAT activity - - - - -

TH/ChAT activity - ↓ - - -

From Slotkin et al., 2008

↑,↓ Significantly different from control (p<0.05) 

- Not different from control

 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 μM PFOA from day 0 to day 6 or 8 to 

determine the impact on phenotype and cytokine secretion. Maturation stimulus (prostaglandin 

E2, tumor necrosis factor, interleukin 1β, and interleukin 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. HLADR is a cytokine which presents antigens to elicit T-cell response. CD25, 80, 83, and 86 are cell 

receptors that act as co-receptors in T-cell activation; and interleukin 12p40 and 10 stimulate Tcells. Mature cytokine-activated dendritic cells secrete interleukin 12p40 and interleukin 10 as 

anti-inflammatory 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 
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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 μM 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 resulted 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. 

Intracellular Communication. Upham et al. (1998, 2009) examined the effects of 

perfluorinated fatty acids on gap junctional intercellular communication 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 gap junctional intercellular communication in the 

livers of rats after treatment for 1 week. In WB-F344 cells, gap junctional intercellular 

communication was inhibited by perfluorinated fatty acids with 7 to10 carbons within 15 

minutes of incubation. The inhibition was reversible with full recovery occurring within 30 

minutes of PFOA removal from media. Extracellular receptor kinase was activated by PFOA 

within 5 minutes of incubation in the cells. Preincubation of cells with the phosphatidylcholinespecific phospholipase C inhibitor D609 partially prevented gap junction intercellular 

communication inhibition by PFOA. The authors concluded that PFOA, having an 8 carbon 

chain, inhibited gap junction intercellular communication by activation of extracellular receptor 

kinase and phosphatidylcholine-specific phospholipase C, but noted that other mechanisms may 

be involved. 

Production of Reactive Oxygen Species. Takagi et al. (1991) fed male Fischer 344 rats diets 

containing 0, 10 or 20 mg PFOA/kg for two weeks to determine the formation of 8-

hydroxydeoxyguanosine (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 were also exposed to a mixture of PFOA and PFOS to determine 

antagonistic or synergic effects. The cells were exposed to 0, 50,100, 150, or 200 μM PFOA or 

to 0, 50, 150, or 200 μM each of PFOA and PFOS. A group of cells were also exposed to 0, 

50,100, 150, or 200 μM PFOS. The cells were cultured for 24, 48, or 72 hours. Cell viability, 

apoptosis, reactive oxygen species, mitochondrial membrane potential, antioxidant enzymes, 
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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 non-significant dose-dependent increase in apoptosis was observed in the cells cultured 

with PFOA. However, the combination of both PFOA and PFOS showed a significant dosedependent increase (p<0.05) in apoptosis. Intracellular reactive oxygen species were 

significantly increased (p<0.05) in cells cultured with 100, 150, or 200 μM PFOA or PFOS. 

HepG2 cells exposed to the mixture of 100 or 200 μM 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 μM PFOA 

significantly decreased (p<0.05) glutathione concentration and glutathione peroxidase activity, 

and 150 μM PFOA significantly increased (p<0.05) the activities of superoxide dismutase, 

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 glutathione-S-transferase 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 down-regulated and caspace-9 was up-regulated 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 reactive oxygen species, impacting mitochondria, and changing gene 

expression of apoptosis gene regulators.

 

 Eriksen et al. (2010) examined ability of PFOA to generate reactive oxygen species 

(ROS) and induce oxidative DNA damage in human HepG2 cells. Cell were incubated with 0, 

0.4, 4, 40, 200, 400, 1000, or 2000 μM PFOA and 2’,7’-dichlorofluorescein diacetate (DCFHDA). Hydrogen peroxide, H2O2, was used as a positive control. A fluorescence 

spectrophotometer was used to measure ROS production every 15 minutes 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 μM PFOA for 24 hours. Cytoxicity 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 for potential adverse effects of PFOA. 
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4.3.5 Structure-Activity Relationship 

 Bjork and Wallace (2009) compared the PPARα-dependent transcriptional activation 

potential of linear perfluorocarboxylid and sulfonic acids in rodent and human hepatic liver cells. 

The PFAAs tested included perfluorinated carboxylic acids with carbon chain lengths of 2 to 8 

and perfluorinated sulfonic acids with chain lengths of 4 to 8. Primary rat and human 

hepatocytes and HEPG2/C3A cells were incubated with 0 or 25 μM perfluorinated compounds 

for 24 hours. Expression of acyl-CoA oxidase (Acox), Cyp4a1 (rat), Cyp4a11 (human), acylCoA thioesterase (Cte-rat, Acot1-human), and DNA damage inducible (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-length 

dependent structure activity relationship. The study results suggested that the PPARα related 

gene expression by perfluorinated compounds was induced in primary rat hepatocytes, increased 

with carbon chain length and appeared to be greater in the carboxylic acids (i.e. PFOA) when 

compared to 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 may 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% dimethyl sulfoxide [DMSO]), 

a positive control (WY14,643) or PFOS at 0.5-100 μM. 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) was 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 μM in the mouse and 5 μM in humans; the LOEC was 1 μM 

(0.43 μg/mL) in the mouse and 10 μM (4.3 μg/mL) in humans. 

A similar study as conducted above (Wolf et al., 2008b) was repeated but included 

additional PFAAs (Wolf et al., 2012). Transfected cells were incubated with PFAAs at 

concentrations of 0.5 to 100 μM, vehicle (water or 0.1% DMSO as negative control) or with 10 

μM WY14,643 (WY, positive control) on each plate. Assays were performed with 3 identical 

plates per compound per species with 9 concentrations/plate and 8 wells/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 human and mouse PPARα was 

induced and was significant in all the PFAAs. Again, similar findings were identified with the 

carboxylates and the mouse PPARα being more reactive. The study also provided the C20max
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which was the concentration at which the PFAA produced 20% of the maximal response elicited 

by the most active PFAA. For PFOA, this was 6 μM in mouse PPARα and 7 μM in human 

PPARα. For comparison, PFOS was 94 μM and 262 μM, respectively.

4.4 Hazard Characterization 

4.4.1 Synthesis and Evaluation of Major Noncancer Effects 

Serum Lipids. Some epidemiological studies have noted an association between PFOA 

exposure and increased serum lipid levels, a risk factor for cardiovascular disease in humans. 

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. Steenland et al. (2009) and Frisbee et al. (2010) examined 

the relationship between serum PFOA and total cholesterol, high density lipoprotein (HDL) 

cholesterol, and low density lipoprotein (LDL) cholesterol in participants of the C8 Health 

project. Both studies reported positive associations between PFOA and total cholesterol. 

Steenland et al. (2009) found significant increasing trends with increasing serum PFOA decile 

for LDL and triglycerides but not HDL. Frisbee et al. (2010) found an association between 

serum PFOA and total cholesterol and LDL in children and adolescents in the C8 study. Nelson 

et al. (2010) found no associations between PFOA and LDL or VLDL in the general population. 

A study of residents in the Little Hocking water district by Emmett et al. (2006) found no 

association between PFOA and cholesterol. Occupational studies (Olsen et al., 2001a,b, 2003; 

Sakr et al., 2007a,b; Costa et al., 2009) found positive significant associations between serum 

PFOA and total cholesterol. These studies also found positive associations between serum 

PFOA and VLDL and LDL (Sakr et al., 2007a), and triglycerides (Olsen et al., 2001 a,b, 2003; 

Olsen and Zobel, 2007). 

Taken together, the above studies show a positive association between serum PFOA and 

cholesterol in both worker and general populations. Although not all studies showed an 

association, positive results were reported in both cross-sectional and longitudinal studies. 

Limited evidence also showed an increase in VLDL and LDL with higher serum PFOA 

concentrations. The C8 Science Panel (2012) concluded that there is a probable link between 

exposure to PFOA and diagnosed high cholesterol which is consistent with the data reviewed by 

EPA. 

High levels of serum lipids are a risk factor for cardiovascular disease in humans 

including ischemic heart disease (IHD), a condition in which blood flow to the heart is decreased 

through the development of atherosclerotic plaque or clots in the cardiac arteries. Sakr et al. 

(2009) conducted a mortality study examining workers at the Washington Works plant and IHD. 

The amount of time spent in relative job categories was used to estimate cumulative exposure. 

There were over 4,000 employees in the mostly male cohort and 239 deaths from IHD. Mortality 

from IHD was not associated with PFOA exposure. These results were similar to Leonard et al. 

(2008) who found a non-significant elevation for IHD mortality when compared to the regional 

worker population. Leonard et al. (2008) used 3 reference populations to derive his mortality 

rates for comparison; the US population, the West Virginia state population, and an 8-state 

regional employee population from the same company. Lundin et al. (2009) found no 

association between serum PFOA and risk of dying from IHD in workers of the Cottage Grove, 

Minnesota, plant compared to the general population of Minnesota.
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Information on serum lipids from animal studies have 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. Activation of the 

PPARα receptor leads to lowered serum lipids because of its induction of peroxisome 

proliferation. Peroxisomes are subcellular organelles that increased beta oxidation of fatty acids. 

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.

Cholesterol and/or triglycerides have been monitored in only a few of the animal studies. 

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 

total cholesterol was significantly decreased (LOAEL= 10.8 mg/kg/day) and total triglycerides 

significantly increased (LOAEL= 5.4 mg/kg/day) in wild type mice. In the PPARα null mice, 

the total cholesterol 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. 

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, and Loveless et al. (2008) 

reported decreased total cholesterol, 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. In male CD1 mice, total cholesterol, HDL, and triglycerides were decreased at 10 and 30 mg/kg/day 

(Loveless et al., 2008). In pregnant female ICR mice, triglyceride, total cholesterol, 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. 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.

 

Clinical Chemistry. Epidemiological studies examining clinical chemistry parameters have 

reported an association between PFOA exposure and increased liver enzymes and decreased 

bilirubin. With the exception of Olsen et al. (2000, 2001a,b, 2003), the occupational studies 

reported positive associations between PFOA and increased GGT (Sakr et al., 2007a; Decatur 

workers only, Olsen and Zobel, 2007; Costa et al., 2009), AST (Sakr et al., 2007b), ALP 

(Decatur workers only, Olsen and Zobel, 2007), or ALT (Decatur workers only, Olsen and 

Zobel, 2007, Costa et al., 2009). A negative association between serum PFOA and bilirubin was 

also observed in some workers (Sakr et al., 2007b; Olsen and Zobel, 2007; Costa et al., 2009). 

Examining the general adult population in the US, Lin et al. (2010) and Gallo et al. (2012) found 

a slight, but significant association with ALT and GGT in some population groups and in 

individuals with a higher BMI. 
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Changes in renal enzymes and hematology were not associated with PFOA exposure in 

worker populations (Olsen et al., 2000, 2001a,b, 2003; Sakr et al., 2007; Costa et al., 2009). 

However, one study found the odds ratio for chronic kidney disease was positively correlated 

with serum PFOA levels in the general population (Shankar et al., 2011). It is noted that a low 

glomerular filtration rate would diminish the removal of PFOA from serum for excretion by the 

kidney. 

In animal studies, serum levels of ALT and/or AST were significantly increased 

indicating apoptotic or necrotic damage to liver cells (Butenhoff et al., 2012; Son et al., 2008; 

Minata et al., 2010). 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 PFOA-exposed (10 mg/kg) pregnant ICR mice. Total protein, albumin, and 

globulin were significantly decreased in the same mice. 

Both the human and animal studies suggest effects on the liver as indicated by increases 

in liver enzymes. The data from animal studies for increases in ALTand AST are more 

definitive than the findings in human epidemiology studies. Concurrent with the evidence in 

animals of damage to liver cells, levels of certain transport proteins have been shown to be 

altered. In mice, increased expression of MRP3 and MRP4 (Maher et al., 2008) and decreased 

expression of OATps (Cheng and Klaassen, 2008) favors excretion of PFOA into the bile but 

may interfere with normal transport of biliary components. 

Hepatic Hypertrophy. In most PFOA animal studies (monkeys, rats, mice), short term 

and chronic exposure caused an increase in liver weight as at least one of the critical effects 

(Christopher and Marias, 1977; Merrick and Marias, 1977; Goldenthal, 1978a; Pastoor et al., 

1987; Butenhoff et al., 2012; Palazzolo, 1993; Butenhoff et al., 2002, 2004a; York, 2002; 

Perkins et al., 2004; Son et al., 2008; Wolf et al., 2008a; DeWitt et al., 2009; Elcombe et al., 

2010; Minata et al., 2010). Increased liver weights were observed in mice that are both active 

and null for PPARα activation (Wolf et al., 2008a; Minata et al., 2010; Albrecht et al., 2013). 

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 PPARα 

null mice were distinctly different from their respective controls (Wolf et al., 2008a; Minata et 

al., 2010). In the case of the wild-type PFOA-exposed mice, there was less endoplasmic 

reticulum than in controls and more lipid-like vacuoles scattered throughout the cytoplasm. The 

PFOA-exposed PPARα null mice had limited endoplasmic reticulum 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 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 
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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. 

Minata et al. (2010) reported degenerative histological changes in the bile ducts of 

PPARα null mice at doses ≥10.6 mg/kg/day. PPARα null mice had an increased hepatocyte 

proliferating cell nuclear antigen 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 effects on the liver that appear to be independent of PPARα activation. 

Other Organs. In general, effects on organs other than the liver tend to occur at doses higher 

than those that affect the liver. Some studies have observed kidney and lung effects including 

tubular epithelium swelling and pulmonary congestion in male Sprague Dawley rats (LOAEL = 

20 mg/kg/day; Cui et al., 2009). In a chronic rat study, increased relative kidney weight was 

observed at 5.64 mg/kg/day (Goldenthanl, 1978a). In a chronic monkey study (Goldenthal, 

1978b), the critical effects were increased pituitary weight in males (LOAEL= 3 mg/kg/day) and 

increased brain and decreased heart weight in females (LOAEL=10 mg/kg/day). The chronic 

study by Butenhoff et al. (2012) noted hematological effects (decreased red blood cells counts 

and hemoglobin concentrations) in treated male and female Sprague Dawley rats at a LOAEL of 

24 mg/kg/day. Increased thickness and prominence of the zona glomerulosa and vacuolation 

were observed in the cells of the adrenal cortex of male rats fed 10 mg/kg/day for approximately 

56 days (York, 2002; Butenhoff et al., 2004a). 

 

Hyperglycemia. Several human epidemiology studies have examined the association between 

PFOA and hyperglycemia both in the general population or a subset of individuals with type II 

diabetes. Leonard et al. (2008) examined mortality associated with diabetes in workers of the 

Washington Works plant and found an association between the Washington Works employees 

and the regional plant worker population (SMR=1.97, 95% CI 1.23, 2.98). Lundin et al. (2009) 

found diabetes mortality to be associated with moderate PFOA exposure (Hazard Ratio 3.7, 95% 

CI: 1.4-10.1) among Cottage Grove workers whose cause of death had been identified as 

diabetes. No deaths from diabetes had occurred among workers with high exposure, based on job 

classification. MacNeil et al. (2009) examined type 2 diabetes among adults from the C8 health 

project and found no association found between serum PFOA concentration and elevated fasting 

serum glucose levels or type II diabetes among members of the community in the vicinity of the 

Washington Works plant. There are limitations in mortality analyses for disorders such as 

diabetes where adherence to diet and other treatment modalities has a major impact on longevity. 

Occupational studies examining glucose levels or related endpoints found no relationship to 

PFOA concentration (Olsen et al., 2000, 2001a,b, 2003; Sakr et al., 2007a; Costa et al., 2009). 

Lin et al. (2009) and Nelson et al. (2010) found no association between PFOA and metabolic 
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syndrome, glucose concentration, insulin levels, and/or insulin resistance in the general US 

population. 

Hines et al. (2009) found no differences in glucose tolerance tests at 15-16 weeks and at 

17 months of age in 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. 

Hypertension. Serum uric acid levels were measured in two occupational populations and in the 

C8 Health Project population as a biomarker for hypertension. Hypertension is a complex 

condition related to a number of physiological conditions. Cholesterol buildup in the arteries is 

one factor that can lead to elevated blood pressure as are kidney function and dietary sodium. 

Impaired circulation can impact glomerular and tubular removal of PFOA and other substrates 

excreted utilizing the common transporters in the renal tubules. Both occupational studies (Sakr 

et al., 2007a, Costa et al., 2009) found an increased association with serum PFOA and uric acid 

levels. The odds ratio for high serum uric acid levels increased from the lowest to the highest 

quintile of PFOA serum concentrations in the C8 Health Project population (Steenland et al., 

2010). The study showed that higher PFOA concentrations were associated with higher 

incidence of high serum uric acid, but uric acid levels remained within the normal range. A 

positive association between chronic kidney disease and serum PFOA levels found in the general 

population (Shankar et al., 2011) was independent of body mass index, hypertension and serum 

cholesterol levels. No epidemiology data were identified that have examined the relationship 

between serum PFOA levels and resting blood pressure, and none of the animal studies have 

examined this end point. 

 

Neurotoxicity. There are limited data on the neurotoxicity of PFOA. Fei et al. (2008b) found no 

association between maternal PFOA concentration and fine motor skills, gross motor skills, and 

cognitive abilities at 6 and 18 months of age. An epidemiology study of children derived from 

the NHANES population found a weak statistical association between serum PFOA with parental 

reports of ADHD (OR=1.12, 95% CI 1.01-1.23). However, one animal study (Johansson et al., 

2009) suggests a potential effect on habituation and activity patterns in NMRI mice treated on 

PND10 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 sex-related exploratory behavior differences. 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 than control males 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 or 100 μM PFOA increased spontaneous synaptic 

current and had an equivocal impact on neurite growth (Liao et al., 2009a, b). These data 

suggest a need for additional studies on the effects of PFAs, including PFOA, on the brain. 

Thyroid Effects. A few epidemiological studies have examined the association between 

presence of PFOA and altered thyroid homeostasis in humans. Thyroid-related measures among 
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fluorochemical production workers were assessed by Olsen et al. (2001a, 2003), Sakr et al. 

(2007a), Olsen and Zobel (2007), and Costa et al. (2009) relative to serum PFOA levels. Most of 

these studies showed no associations between PFOA and any thyroid hormones. Olsen and 

Zobel (2007) reported a negative association for free T4 and a positive association for T3; 

however, the hormone levels were well within the normal reference range. The authors 

concluded that there was no consistent relationship between altered thyroid hormone status and 

serum PFOA. 

Emmett et al. (2006) found no association between serum PFOA and thyroid hormones 

and/or disease in residents served by the Little Hocking water district. Results from a study by 

Bloom et al. (2010) did not identify an association between thyroid function and nonoccupational PFOA exposure (mean serum concentration of 1.3 ng/ml; 95% CI: 1.15-1.53). 

TSH and free T4 levels were used as markers for thyroid function. Only one subject (3.2%) 

exceeded the laboratory reference range for normal (0.40-5.00 μIU/mL) and no subjects 

exceeded the laboratory reference range for normal free T4 (0.80-1.80 ng/dL). Pirali et al. 

(2009) reported no relationship between intrathyroidal concentrations of PFOA and underlying 

thyroid disease. 

Chan et al. (2011) found no association between serum PFOA concentrations and 

hypothyroxinemia in pregnant women. More recently, Melzer et al. (2010) examined the records 

of the NHANES data for thyroid disease and serum PFOA levels in the U.S. general population 

and found that women with PFOA ≥ 5.7 ng/ml were more likely to report current thyroid disease 

compared to women with PFOA ≤ 4.0 ng/ml. A near significant and similar trend was observed 

in men. However, lack of segregating the thyroid disorders into those associated with 

hyperactivity versus hypoactivity was a limitation in their analysis. The physiological 

significance of these cross-sectional analyses remains to be determined.

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 free T3 

in circulation were reduced significantly in the 30/20 mg/kg/day treatment group, an effect seen 

as early as 5 weeks after initiation of treatment. Recovery of T3 deficits was noted upon 

cessation of chemical treatment after serum level of PFOA returned to baseline 90 days later. 

Serum total T4, free T4 and TSH were not altered throughout the study. The preferential effects 

of PFOA on serum T3 and a lack of TSH compensatory response are similar to those observed 

with PFOS.

Martin et al. (2007) showed that serum total T4 and free T4 were markedly and abruptly 

(1 day after oral gavage treatment) depressed (~ 80%) by PFOA (20 mg/kg) in adult male rats; 

while serum T3 was also reduced (25%) though 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). Serum total T4 and 

free T4 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 and PFOA has 

yet to be evaluated and the lowest effective dose of the chemical remains unknown. 
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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 may be 

relevant to humans because of the long PFOA human half-life. 

Displacement of T4 from binding to the hormone transport protein transthyretin (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.

Reproductive and Developmental Effects. There is a wealth of data on the impact of PFOA 

exposure on reproductive and/or developmental effects in both humans and animals. A series of 

epidemiological studies have been published examining a variety of endpoint: fertility 

(fecundity), growth, and developmental biomarkers or milestones. The animal data have focused 

more extensively on developmental endpoints using traditional and non-traditional approaches.

Joensen et al. (2009) found that there was no association between serum levels of PFOA 

and testicular hormones among 105 Danish men who provided blood and semen for the military 

draft in 2003. A nonsignificant negative association was observed between serum PFOA 

concentration and semen volume, sperm concentration, sperm count, sperm motility, or sperm 

morphology. Fei et al. (2009) aimed to measure fecundity by using longer time to pregnancy as 

the outcome. More than 1200 women were included in a phone survey. Infertility was defined 

as having > 12months time to pregnancy (TTP); pregnancies were separated into: planned, 

partially planned, and not planned. The fecundity odds ratios for the three highest PFOA 

concentration quartiles were 0.72, 0.73, and 0.60. The likelihood ratio test for trend was 

significant (p< .0001). The results provide suggestive evidence that PFOA concentration may 

reduce fecundity, however, the results should be viewed with caution, due to potential selection 

bias and/or confounding biological and behavioral factors. 

Several studies have looked at the association between PFOA concentration and fetal 

growth (birth weight, birth length, head/chest circumference, etc); whereas some researchers 

found no significant associations, the results for others were positive. Monroy et al. (2008) 

found no association between umbilical cord or serum PFOA concentration and birth weight in 

105 mothers in the Family Study conducted in Canada. These findings were supported by Nolan 

et al. (2009) who examined mothers and babies who were serviced by the Little Hocking Water 

Association (LHWA). Many researchers used log-transformed values and adjusted for possible 

confounders, but still found no significant association between PFOA and birth weight, birth 

length, head, or chest circumference (Hamm et al., 2009, Washino et al., 2009). Conflicting 

results were seen in the early studies from Fei et al (2007, 2008a) where there was a statistically 

significant inverse association between birth weight and plasma PFOA concentration among 

“normal-weight” women. A statistically significant decrease in birth length was also observed 

when the second and fourth quartile measures of PFOA exposure were compared to those for the 

first quartile. 
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In studies that examined pregnancy outcome among women believed to be exposed to 

PFOA, there were no association seen between congenital anomalies and complications of labor 

and delivery such as meconium, prolonged labor, breach, cord prolapse, or fetal distress (Nolan 

et al., 2010). There was also no association between PFOA and birth outcomes including 

miscarriages and livebirths (Stein et al., 2009). Knox et al. (2011) suggested that PFOA may be 

associated with endocrine disruption in women after finding that the odds of having experience 

menopause were significantly higher in the highest PFOA quintile group relative to the lowest 

PFOA group. 

Fei et al. (2008b) found no association between maternal PFOA concentration and fine 

motor skills, gross motor skills, cognitive skills, at 6 and 18 months of age. The age at which 

children reached developmental milestones did not show any relationship to maternal plasma 

PFOA concentration. Maternal plasma PFOA concentration was also not associated with child 

risk of hospitalization for infectious diseases (Fei et al., 2010), but was associated with decreased 

antibody titers in children following vaccination (Grandjean et al., 2012). In the general 

adolescent population in the US, a positive association was observed between serum PFOA and 

parent-reported ADHD (Hoffman et al., 2010). Lopez-Espinosa et al. (2011) reported reduced 

odds of having reached puberty in girls included in the C8 Health project. 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. 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, but some effects were observed in mice. Both species, however, showed 

some indications of the potential developmental toxicity of PFOA. Doses which elicited a 

response were much higher in rats compared with those in mice. The species differences in 

response and dose rate are likely related to half-life differences of hours for the female rat and 

days-to-weeks for the female mouse. 

Body weight gain was decreased in parental male animals in a two-generation study after 

dietary exposure to 30 mg/kg/day PFOA (York, 2002; Butenhoff et al., 2004a, York et al., 2010). 

Maternal body weight gain was reduced at ≥30 mg/kg/day and death occurred at ≥100 mg/kg/day 

in rats given PFOA via gavage during gestation (Staples et al., 1984; Hinderliter et al., 2005; 

Mylchreest, 2003). In mouse gavage studies, increased time to parturition was observed at ≥10 

mg/kg/day, and decreased maternal weight gain was observed at ≥20 mg/kg/day (Lau et al., 

2006). Mouse maternal mammary glands displayed stunted alveolar development (GD18) and 

delayed epithelial differentiation (PND10) (White et al., 2009). 

 

 Decreased fetal and offspring body weight was observed in rats following an inhalation 

exposure to 25 mg/m3

 for 6 hours/day on GD6-15 (Staples et al., 1984). 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 (York, 2002; 

Butenhoff et al., 2004a). In mouse gavage studies, decreased body weight and decreased 

neonatal survival were observed at ≥ 1 mg/kg/day and increased full litter resorptions and 

increased stillbirths were observed at ≥ 5 mg/kg/day for exposures lasting from GD1 to 17 (Lau 

et al., 2006; Abbott et al., 2007; White et al., 2007; Wolf et al., 2007). 
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Delayed mammary gland development of female pups was also observed following 

maternal doses ≥ 0.01 mg PFOA/kg in one study (Macon et al., 2011) and ≥3 mg/kg in another 

study (White et al., 2009). However, Albrecht et al. (2013) found no 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 following a maternal dose of 3 mg/kg. 

Two developmental studies with compared wild-type mice with PPARα null mice but 

results are inconclsive. 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 is involved in the 

etiology of PFOA-induced neonatal mortality. 

Cross-fostering studies showed that gestational PFOA exposure was required to produce 

decreased body weight, delayed eye opening, delayed body hair growth, and decreased survival 

in the offspring. However, persistent (PND1-PND63) mammary gland development delays were 

observed in offspring that had been gestationally or lactationally exposed as well as in offspring 

who had been gestationally and lactationally exposed (White et al., 2009). Restricted exposure 

studies showed that gestational exposure to PFOA during a specific time period led to differeing 

offspring effects (Wolf et al., 2007). Delayed eye opening and body hair growth were observed 

at 5 mg/kg/day in offspring exposed GD7-17 or 10-17, but decreased postnatal survival was

observed at the same dose in offspring exposed GD15-17. Delayed mammary gland 

development was observed irrespective of exposure duration, and delayed lactational 

differentiation was seen in the dams (White et al., 2009). In a three generation study, PFOA 

induced delays in mammary gland development and/or lactational differentiation. A chronic 

dose of 5 ppm PFOA in the drinking water also altered mammary morphology in mice (White et 

al., 2011). The mode of action for PFOA-induced delayed mammary gland development is 

unknown. 

 Delayed vaginal opening was observed in BALB/c mice who received 1 mg/kg/day by 

gavage for four weeks starting on PND21 (Yang et al., 2009). At higher doses (5 and 10 

mg/kg/day), vaginal opening did not occur and mammary gland development was delayed. In 

C57BL/6 mice, vaginal opening was delayed at 5 mg/kg/day and did not occur at 10 mg/kg/day. 

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. 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. 

 

Immunotoxicity. An occupational study of the immunotoxicity of PFOA found no association 

between the immunoglobins IgA, IgG, and IgM and PFOA in male workers (Costa et al., 2009). 

No associations were found between maternal PFOA concentrations and hospitalization for 

infectious diseases in early childhood (Fei et al., 2010b) or infant allergies (Okada et al., 2012). 
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In contrast, higher PFOA concentrations in children were associated with lower antibody levels 

to diphtheria and tetnus (Grandjean et al., 2012) and a greater risk of having asthma (Dong et al., 

2013). In a pilot study using volunteer donated blood, Brieger et al. (2011) found that PFOA 

altered peripheral blood mononuclear cell viability and reported a significant association was 

observed between PFOA concentration and the release of LPS-induced TNF-α and IL-6 by 

peripheral monocytes. 

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 splenocyte and thymocyte populations at 

various stages of differentiation. Recovery usually occurred within several days of cessation of 

PFOA dosing. 

 

 Activation of the PPARα receptor appears to be related to at least a part of the thymus 

response to PFOA. When Yang (2002a) dosed PPARα null mice with the same dose that caused 

an effect in the wild-type mice, no significant changes in spleen weight or cellularity were 

observed, but there was a small and significant decrease in thymus weight and cellularity 

compared to controls. 

 In one component of the Yang et al. (2002b) study, the functional impact of changes in 

spleen and thymus were evaluated through the response of treated mice to horse red blood cells 

(HRBC). 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 serum from PFOA treated mice was 

evaluated post treatment, 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. 

 DeWitt et al. (2008) used different functionality assays in their study in C57Bl/6 mice. 

The IgM response to sheep red blood cells (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 four 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.

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 at the NOAEL of 1.88 mg/kg/day. 

 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, 
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however, thymus and spleen weights, thymus and spleen cell counts, and IgM titers were 

decreased at ≥10 mg PFOA/kg/day. Corticosterone was also increased in mice at the same 

doses.

 The data collected from the immunotoxicity studies support a mode of action 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 adrenalectomized mice treated with PFOA were 

compared to sham operated controls did not support the Loveless et al. (2008) hypothesis. 

 

 Data from PPARα null mice suggest that rodents may 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 null mouse strain indicate the potential 

for an inadequate humoral response in exposed populations. 

4.4.2 Synthesis and Evaluation of Carcinogenic Effects 

A number of human epidemiology studies were conducted on worker populations to 

examine an association between PFOA exposure and cancer. Lundin et al. (2009) saw no 

association between cancer mortality and exposure to PFOA when comparing workers of the 

Cottage Grove facility to the Minnesota general population. The mortality data for liver cancer, 

pancreatic cancer, prostate cancer, and testicular cancer all showed no association to exposure. 

The Dupont (2003) report on the West Virginia Washington Works Plant found statistically 

significant elevations in cancer of the bladder (SIR 1.9; 95% CI: 1.15-3.07) and kidney and 

urinary organs (SIR 2.3; 95%CI: 1.36-3.65). This same report found non-statistically significant 

elevations in leukemia incidence.

 

Two analyses of leukemia incidence were conducted prior to the Dupont (2003) report. 

Walrath and Burke (1989) reported statistically significant elevated Odds Ratio of 2.1 for 

leukemia incidence in male employees. Karns and Fayerweather (1991) conducted a follow up 

case-control study at the same Washington Works facility where they matched employees (four 

controls for each case). The matched Odds Ratios were significantly elevated for employees 

who had previously worked as custodians and engineers, 8.0 (90% CI, 1.1-60.0) and 7.9 (90% 

CI, 1.0-76.0), respectively. The level of significance chosen by the authors for each p=.10 were 

not comparable in stringency to those used in many of the other papers; also the CIs were quite 

wide even at the 90%level.
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Two human epidemiology studies focused on general populations, one of which was 

exposed to PFOA through contaminatd water supplies. First, Eriksen et al. (2009) looked at risk 

of cancer in the Danish general population and plasma PFOA concentration. The Incidence Rate 

Ratios (IRRs) crude and adjusted showed no association between plasma PFOA concentration 

and prostate, bladder, pancreatic, and liver cancer. Vieira et al. (2013) showed a positive 

association between PFOA levels and kidney and testicular cancers in residents living in an area 

with contaminated water. Positive associations were found between kidney cancer and the very 

high and high serum 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) and between for testicular cancer with the very high exposure 

category (OR: 2.8, 95% CI: 0.8, 9.2; n=6). No associations for either cancer were found in the 

lower exposure categories. Ovarian cancer, non-Hodgkin’s lymphoma, and prostrate cancer 

were also positively associated with the very high exposure category, but showed weaker or 

negative associations for the other exposure categories.

Two animal carcinogenicity studies have shown that PFOA exposure can lead to liver 

adenomas or carcinomas (Butenhoff et al., 2012; Biegel et al., 2001), Leydig cell adenomas 

(Butenhoff et al., 2012; Biegel et al., 2001), and pancreatic acinar cell tumors (PACT) (Biegel et 

al., 2001) in male Sprague-Dawley rats. Liver adenomas were observed in the Biegel et al. 

(2001) study 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 both the male controls and exposed male and female rats in the 

high dose group and exposed males in the low-dose group. The differences from control were 

not significant in either case but the incidence among the high dose males (10/50) was similar to 

that for the adenomas in the Biegel et al. (2001) study. Liver lesions were identified in the males 

and females at the one- and two-year sacrifices (Butenhoff et al., 2012). Increased incidence of 

diffuse hepatomegalocytosis and hepatocellular necrosis were occurred at 20 mg/kg/day. At the 

2-year sacrifice, hepatic cystoid degeneration was observed at 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% vs. 0% in control rats). 

Testicular Leydig cell tumors (LCT) 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. (2001) study included one dose group (20 mg/kg/day), and 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. PACTs were only observed in the Biegel et al. (2001) 

study at an incidence of 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; 2/34 (6%) and 1/43 (2%), 

respectively. Re-examination 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, and the lesions in the Biegel et al. (2001) study had progressed 

to adenomas. 

The initial findings from the Butenhoff et al. (2012) study were equivocal for mammary 

fibroadenomas in female rats. However, a re-examination of the tissues by a PWG found no 

statistically significant differences in the incidence of fibroadenomas or other neoplasms of the 
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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 were also observed in female rats (Butenhoff 

et al., 2012). Mann and Frame (2004) re-examined 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 (Lawlor, 1995, 1996; Friere et 

al., 2008) 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. In clastogenicity studies in CHO 

by Murli (1996b, 1996d), the results were positive with activation for chromosomal 

abnormalities and polyploidy and equivocal in the absence of activation. Micronucleus assays 

by Murli (1995, 1996a) 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 reactive oxygen 

species, impacting mitochondria, and changing expression of apoptosis gene regulators. Erikson 

et al. (2010) observed a PFOA-induced increase in reactive oxygen species production in HepG2 

cells, but no PFOA-induced oxidative DNA damage or cytotoxicity. 

4.4.3 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 non-genotoxic 

mechanism involving receptors and perturbations of the endocrine system. 

Rat Liver Tumors. PPARα agonism is the proposed mechanism of action (MOA) for the liver 

carcinomas and adenomas in rats following chronic PFOA exposure (Maloney and Waxman, 

1999; Klaunig et al., 2003). In the PPARα agonism MOA (Figure 4-1), binding of PFOA to the 

PPARα receptor results in increased peroxisome proliferation and cell replication. Peroxisome 

are single-membrane organelles found in a number of plant and animal cells that have the 

capacity to carry out oxidation of long chain fatty acids and other substrates through hydrogen 

peroxide generating pathways (Goodrich and Sul, 2000). Peroxisome beta oxidation produces 

acetyl CoA that can be utilized for anabolic reactions but is not linked to ATP production. 

Peroxisomes metabolize the long chain fatty acids via both beta and omega oxidation pathways 

(Fielding, 2000). 
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 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).

 The key events in the MOA of PPARα-agonist induced liver tumors involve four causal 

key events (Klaunig, 2003). The first key event is activation of PPARα. Other key events are 

cell proliferation, apoptosis, and clonal expansion. Of these key events, PPARα activation is 

highly specific for this MOA whereas cell proliferation, apoptosis, and clonal expansion are 

common for other MOAs. 

From U.S. EPA, 2005b 

FIGURE 4-1. PPARα-Agonism Mode of Action for Liver Tumors

The primary data that demonstrate PFOA activation of the PPARα receptor are those 

from Rosen et al (2008a,b) 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, and evidence that PFOA also acts through PPARα independent 

mechanisms associated with CAR and PXR receptors. Martin et al. (2007) examined the 
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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. 

Few studies examined the occurrence of steps in the PPARα agonism MOA beyond 

activation of PPARα. The one study that looked at apoptosis from a MOA perspective (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 PFOA (comparable to the high dose in both cancer 

studies). The data in mice were not consistent. Minata et al. (2010) observed increased apoptosis 

in hepatocytes, arterial walls, and bile duct epithelium of wild type 129S4/SvlmJ mice and in the 

bile duct epithelium of PPARα-null mice at 10.8 or 21.6 mg/kg PFOA. On the other hand, 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. 

There are studies in male rats that examined cell proliferation. 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 dietary dose. The highest 

increase was observed after 7 days of treatment (5-fold increase) which had declined to a 2-fold 

increase after 28 days of dosing. The liver results from the Biegel et al. (2001) mechanistic 

study were negative for cell proliferation in male Sprague-Dawley rats exposed to the same 

dietary concentration (20 mg/kg/day) and sacrificed at 1, 3, 6, 9, 12, or 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. (2012) study identified 

hyperplastic nodules in 6/50 high dose males and 4/50 high dose females at 20 mg/kg/day; 10/50 

males and 2/50 females had hepatocellular carcinomas.

The study by Wolf et al. (2008) 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 but only at 10 mg/kg/day, the highest 

dose, in PPARα-null mice. This suggests PPARα involvement in increased cell division of wildtype mice at lower doses.

There were no studies identified that focused specifically on preneoplastic foci and clonal 

expansion of altered cells after PPAR exposures. 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. 

There are several “associative” events that are markers of PPARα agonism but are not 

directly involved in the etiology of liver tumors. These include peroxisome proliferation and 

expression of genes for enzymes associated with the peroxisomes. These associative events 

include (Klaunig et al., 2003): 

• In vivo evidence of an increase in number and size of peroxisomes,

• Increases in the activity of acyl CoA oxidases such as palmitoyl CoA oxidase activity, 

• Hepatocyte hypertrophy, increase in liver weights, 

• Hepatic CYP4A1 (Lauric acids monooxygenase) induction, and 
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• Peroxisomal gene expression. 

There is ample evidence that PFOA is a potent peroxisome proliferator inducing peroxisome 

formation in the livers of rats and mice (Ikeda et al, 1985; Pastoor et al., 1987; Sohlenius et al., 

1992; Yang et al., 2001; Wof et al., 2008; Minata et al., 2010; Elcombe et al., 2010). PFOA was 

also found to activate mouse and human PPARα using a transient transfection cell assay (Takacs 

and Abbott (2007). Maloney and Waxman (1999) demonstrated that 5-10 μM PFOA activated 

mouse PPARα using COS1 cells (kidney fibroblast derived cells) transfected with a luciferase 

reporter gene. 

The temporal and dose-response relationship of measures of peroxisome proliferation, 

hepatocellular hypertrophy, liver weight, and liver histopathology have been 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 (Palazzolo, 1993; Perkins et al., 2004). There was no evidence of 

peroxisome proliferation, hepatocellular hypertrophy or liver weight increases at 0.06 mg/kgday. At doses ranging from 0.64 to 6.5 mg/kg-day there is a clear relationship between 

peroxisome proliferation (indicated by increased palmitoyl CoA oxidase activity), hepatocellular 

hypertrophy and increases in liver weight at all time points. 

One can conclude, based on the available data, that the liver tumors in the PFOA 

Sprague-Dawley cancer bioassays can be attributed to its peroxisome proliferative impact since 

there are data that supports each of the steps in the proposed mode of action and thus are not 

relevant to humans (Klaunig et al., 2003). 

Leydig Cell Tumors (LCT). The mode of carcinogenic action for PFOA-induced LCTs 

identified in Biegel et al. (2001) has not been fully elucidated. A large number of non-genotoxic 

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 (Carpino et al., 2007). The consensus 

conclusions of a workshop on the MOA for LCT were reported by Clegg et al. (1997) and 

classified the chemicals that caused LCT in animal studies into seven MOA categories. The 

postulated MOAs support the following hormonal steps to the process: 

• A xenobiotic chemical inhibits the production of testosterone leading to low serum levels 

• Low serum testosterone levels signal the hypothalamus to produce gonadotropin 

releasing hormone (GnRH) 

• GnRH signals the pituitary to release luteinizing hormone (LH)

• LH signals the Leydig cells to produce testosterone 

• LH causes Leydig cell proliferation in testosterone production 

The sustained increase in circulating LH and chronic stimulation of Leydig cells by 

growth-stimulating mediators including IGF-1, TGF-α, leukotrienes and various free radicals 

could lead to DNA replication errors and LCT development (Clegg et al., 1997). TGF-α, which 

binds to the epidermal growth factor receptor and stimulated cell proliferation has been detected 

in Leydig cells (Teerds et al., 1990). 

A series of mechanistic studies have been conducted by Cook and colleagues (Cook et 

al., 1992; Biegel et al., 1995; Liu et al., 1996) to investigate the mechanism of LCT formation in 
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male Sprague-Dawley rats exposed to PFOA. Administration of PFOA to adult male rats by 

gavage for 14 days was shown to decrease testosterone levels and increase serum estradiol levels 

(Cook et al., 2002). These endocrine changes correlated with its potency to induce LCTs in rats 

and were hypothesized to play a role in the PFOA-induction of LCTs (Biegel et al., 2001). 

Subsequent experiments have shown 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 has also been shown to directly inhibit 

testosterone production when incubated with isolated Leydig cells and ex vivo studies 

demonstrated that this inhibition was reversible (Biegel et al., 1995). This inhibition of 

testosterone biosynthesis appeared to be mediated by PPARα (Gazouli et al., 2002) and may 

contribute to the development of LCTs through disruption of the hypothalamus-pituitarytesticular axis. 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 were tested. 

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 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 shown to occur in wild-type and humanized 

PPARα mice. These effects were not seen in PPARα-null mice.

The induction of LCTs by PFOA can be attributed to a hormonal mechanism whereby 

PFOA either inhibits testosterone biosynthesis and/or lowers testosterone by increasing its 

conversion to estradiol by upregulating aromatase activity in the liver. Both of these 

mechanisms appear to be mediated by PPARα. However, the data are not currently sufficient to 

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. 

Pancreatic Acinar Cell Tumors (PACT). As with LCT, the MOA for PACTs is not 

understood. These tumors are most commonly identified in rats but do occur in other animal 

species (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 

(Obourn et al., 1997; Klaunig et al., 2003): 

• Events that stimulate increases of growth factors such as cholecystokinin (CCK) 

and/or gastrin activate a feedback loop resulting in proliferation of the secretory 

pancreatic acinar cells. A high fat diet, trypsin inhibition, cholestasis, or changes in 

bile composition are proposed initiators of this sequence of events. 

• Increased levels of testosterone support the growth of acinar cell preneoplastic foci 

and carcinomas and estrogens inhibit the cellular proliferation.

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 
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vitro assays and found that PFOA was not an agonist for the CCKA receptor that activated CCK 

release. PFOA also had no inhibitory action on trypsin at levels 1000 times greater (0.31 μg/mL) 

than the positive control. 

The Obourn et al. (1997) study 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 ppmWyeth-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. 

Plummer et al. (2007) reported on gene expression changes induced in pancreatic acinar cells 

isolated from Sprague-Dawley rats fed diets containing 300 ppm (~20 mg/kg/day) PFOA for 28 

days. Expression of genes regulated by PPARα, γ, δ in pancreatic acinar cells was directly 

opposite of the expression of those same genes in liver tissue. 

As indicated in the discussion of the LCTs above, PFOA appears to suppress testosterone 

production through the induction of aromatase. Accordingly, the second proposed MOA for 

PACTs is not applicable to PFOA. At the present time, there are insufficient data to demonstrate 

a MOA that can account for the PACTs identified in the chronic study by Biegel et al. (2001). 

Other Potential Modes of Action. There are other potential MOAs of actions that may 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.

Gap junction intercellular communication (GJIC), a process by which cells exchange 

ions, second messages, and other small molecules, is important for normal growth, development, 

and differentiation, as well as 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 may 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. Berthiaume and Wallace 

(2002) noted an increase in mitochondrial biogenesis in liver following treatment of rats with 

PFOA. PFOA has also been 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 slightly decreases 

the membrane potential. The observed effects are believed to be attributed to a slight increase in 

nonselective permeability of the mitochondria membranes caused by PFOA’s surface-active 

properties (Starkov and Wallace, 2002). 

 

Estrogen has been shown to promote hepatocarcinogenesis in rats (Yager and Yager, 

1980; Cameron et al., 1982). Thus, the increase in estrogen levels after PFOA exposure may 

play a role in hepatocarcinogenesis in rats as well as influencing LCTs. More research is needed 
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to support the involvement of this MOA to the hepatocarcinogenesis of PFOA. 

4.4.4 Weight of Evidence Evaluation for Carcinogenicity 

The findings for cancer in humans are equivocal; some epidemiological studies show a 

positive association with exposures and others are negative. The Dupont (2003b) report on the 

West Virginia Washington Works Plant found statistically significant elevations in cancer of the 

bladder, kidney, and urinary organs and a nonstatistically significant elevation in leukemia 

incidence. Two earlier worker studies had statistically significant findings for leukemia but both 

had methodological limitations (Walrath and Burke, 1989; Karns and Fayerweather, 1991). The 

mortality data for liver cancer, pancreatic cancer, prostate cancer, and testicular cancer all 

showed no association to exposure. In contrast, the highest serum PFOA levels were associated 

with testicular, kidney, prostate, and ovarian cancers and non-Hodgkin’s lymphoma among the 

general population exposed through contaminated water (Vieira et al., 2013). These findings are 

limited by small numbers of cases or a positive association with the very high exposure category, 

but weaker or negative associations for the other exposure categories. 

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 may also play a role in the other tumor types observed, but more 

data that will 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 fact that PFOA is not metabolized and would not be activated by the microsomal enzymes, 

the clastogenic effects were likely the result of an indirect mechanism. 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 conditions perturbed by PFOA rather than 

PFOA alone. A compound that is not metabolized will not be able to covalently introduce any of 

its atoms into the structure of DNA.

Despite the limitations in the data for the Leydig cell and pancreatic tumors, a finding of 

Suggested evidence of carcinogenicity is justified and quantification of the dose response is not 

recommended. The tumor dose-response data are not indicative of a high potency, and the doseresponse information from the noncancer studies indicate that protecting for several of the 

noncancer endpoints (e.g. increases in liver weights, hepatocyte cellular abnormalities, 

immunotoxicity, developmental effects and delays) will result in an RfD that will be protective 

for any tumorigenic effects. 

4.4.5 Potentially Sensitive Populations

 Human biomonitoring studies do not suggest major differences between serum PFOA 

levels in males and females. However, the worker populations that would be those most likely to 

demonstrate such differences because of their higher exposures were predominantly male. 
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Some animal models have gender differences which 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 (Hundly et al., 2006). Studies of the transporters involved 

in the toxicokentics of PFOA demonstrate that they are differentially impacted by the presence of 

male and female sex hormones (Kudo et al., 2002; Cheng et al., 2006) influencing tissue 

persistence. 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 half 

lives (increased excretion). The gender differences in mice are not as pronounced as those in 

rats. Work by Cheng et al. (2002) and Cheng and Klaassen et al. (2009) demonstrated that 

hormones impact transporters in liver and kidney. 

In studies in which male and female rats were used, the males were more sensitive to 

toxicity than female rats (Goldenthal, 1978a; York, 2002; Butenhoff et al., 2004a; Sohlenius et 

al., 1992). Monkeys and mice displayed similar sensitivities following PFOA exposure 

(Goldenthal, 1978b; Christopher and Marias, 1977; Sohlenius et al, 1992). In the monkey 

studies the number of animas/sex/dose group was too small reveal a difference related to gender.

Unfortunately much work remains to be done in order 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 may be more like the male rat rather than the female rat. There is a 

broad range of half-lives in human epidemiology studies suggesting a variability in human 

transport capabilities resulting from genetic variations in 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. 

 An epidemiological study examining postnatal outcomes including developmental 

milestones at 6 and 18 months of age did not find any differences in fine motor, gross motor, or 

cognitive skills relative to maternal PFOA concentration (Fei et al., 2008b).

Several animal studies have examined potential modes of action for developmental 

effects following prenatal exposure to PFOA. PFOA-exposure during development in rabbits, 

rats, and mice resulted in increased resorptions (mouse) increased fetal skeletal variation (rabbits, 

rats, 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), delayed mammary gland development (mouse dam and 

offspring) (York, 2002; Butenhoff et al., 2004a; Lau et al., 2006; Abbott et al., 2007; Wolf et al., 

2007; White et al., 2007, 2009, 2011; Macon et al., 2011). 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 mechanisms of action for 

these developmental effects are unknown, but several potential modes of action have been 

investigated. 
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 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 that resulted in the 

observed developmental effects. Decreased postnatal survival was observed at the highest dose 

used (20 mg/kg/day). Eye opening and body hair growth were delayed in offspring exposed for 

the longest periods of time (GD7-17 and GD10-17) and may 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 were observed in offspring exposed 

via lactation only indicating that PFOA alters growth regulation in the developing fetus which 

persist as growth continues postnatally. 

 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 (GD1-17) PFOA exposure (≥5 mg/kg/day) indicating the mechanism of PFOAinduced resorptions was independent of PPARα expression. These resorptions may 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 PPARα 

independent factors may also 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 

PFOA-exposed offspring. White et al. (2007) found that dams dosed with PFOA GD1-17 and 

GD8-17 had significantly delayed mammary gland development (full of alveoli, visible adipose 

tissue, not well differentiated) at PND10 which is the peak of lactation in rodents. The delayed 

dam mammary gland development may play a role in the observed reduced offspring body 

weight gain if the quantity or quality of the milk is altered by PFOA (Lau et al., 2006; Abbott et 

al., 2007; Wolf et al., 2007; White et al., 2007). Restricted gestational exposure and crossfostering studies showed that delayed offspring mammary gland development, observed PND1-

63, occurred regardless of exposure duration or timing (gestation vs. lactation exposure). The 

developmental delays persisted even as the internal PFOA dose decreased (White et al., 2007, 

2009, 2011; Macon et al., 2011). More studies need to be conducted to elucidate the MOA for 

dam and offspring mammary gland effects. 

Mammary gland development was also affect by peripubertal exposure to PFOA (Yang et 

al., 2009, 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. Experiments examining the mechanism for accelerated 

mammary gland development showed that PFOA promotes steroid hormone production in the 
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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, may place PFOA 

into the environmental endocrine disruptor obesogen category similar to diethylstilbestrol (DES) 

(Newbold et al., 2007). In humans, increased leptin levels are associated with increased body fat 

and suggestive of a leptin-resistance mechanism of action for overweight (Considine et al., 

1996). A similar relationship may occur in prenatally PFOA-exposed mice. Studies determining 

MOAs should be conducted to determine relevance to human health. 
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5.0 DOSE-RESPONSE ASSESSMENT 

5.1 Dose-Response for Noncancer Effects 

A Reference Dose (RfD) or Reference Concentration (RfC) is used as a benchmark for 

the prevention of long-term 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, etc. The RfD is expressed in terms of milligrams per kilogram 

per day (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.

5.1.1 RfD Determination 

The risk assessment for PFOA is fraught with a number of challenges including:

• Its toxicokinetic complexity 

• The variability in half-lives between and within species

• The wealth of inconsistent data from epidemiological studies 

• The robust but diverse database of animal studies 

• Inconsistencies between some of the associations from the epidemiology studies 

and the effects from the animal studies

• The metabolic inertness of PFOA in living systems

• The number of animal studies that lack a NOAEL. 

 

A subset of the challenges has an impact on dose-response, especially the toxicokinetic 

features that lead to differences in half-lives across species and in the case of rats, genders. 

Toxicokinetics also influence intraindividual and lifestage variability. Each of these 

toxicokinetic features highlights the importance of internal dose, and supports the utilization of a 

pharmakolinetic model as a component of the dose-response assessment.

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. Endpoints monitored included standard clinical chemistry 

parameters, measures of cardiovascular risk, signs of organ damage, standard haematological 

endpoints, and diabetic or prediabetic conditions. Reporductive and developmental parameters 

were also evaluated in some epidemiology studies. 

In reviewing the epidemiological monitoring data for PFOA it is important to remember 

that many individuals were coexposed to other fluorocarbons, and, although only serum PFOA 

measurements are often given, many individuals had other PFCs present in their serum. In no 

study was the actual exposure concentration or dose to the individuals known; thus, a 

quantitative exposure-response assessment based on the human data is not feasible.

 

In many C8 project studies, drinking water was identified as the primary contributor to 

exposure. However, indoor air and dietary exposures were possible and quantitative exposure 
Perfluorooctanoic Acid – February 2014 5-2 

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data are lacking for all routes. Similarly, occupational epidemiology studies cannot quantify 

exposures occurring away from the workplace. In all of the epidemiology studies the wide 

ranges of serum levels are likely a reflection, not only of intrahuman toxicokinetic variability, 

but also of diversity in external exposure sources. This is especially evident in the units used to 

express PFOA serum levels in occupational populations, μg/mL, versus in the general 

population, ng/mL. From the studies discussed in Section 4.1.1.1, mean serum levels in the 

occupational cohorts ranged from 1.02 – 4.63 μg/mL while those for the general population 

ranged from 1.05 – 354 ng/mL. The highest level seen in the general population is from 

residents serviced by a contaminated water district. For both the positive and negative findings 

summarized in the paragraphs below, mean or median serum PFOA levels fall within these 

ranges for worker and general populations. 

The epidemiological studies have found a positive association between serum PFOA 

concentration and total cholesterol in the general population (Steenland et al., 2009; Frisbee et 

al., 2010; Nelson et al., 2010) and in worker populations (Olsen et al., 2001a,b, 2003; Sakr et al., 

2007a,b; Costa et al., 2009) but no consistent trends in LDL, VLDL, or HDL levels. A positive 

association has been shown between serum PFOA concentration and increased liver enzymes 

and/or decreased bilirubin in both worker and general populations (Sakr et al., 2007,a,b; Olsen 

and Zobel, 2007; Costa et al., 2009; Lin et al., 2010; Gallo et al., 2010), chronic kidney disease 

in the general population (Shankar et al., 2011), and the odds of experiencing early menopause 

(Knox et al., 2011). Maternal or child plasma levels of PFOA were positively associated with 

decreased antibody titers in children after vaccination (Grandjean et al., 2012), obesogenic 

effects in female children at 20 years of age (Halldorsson et al., 2012), delayed puberty in girls 

(Lopez-Espinosa et al., 2011), and parent reported ADHD (Hoffman et al., 2010). 

No association between serum PFOA and risk of dying from ischemic heart disease has 

been found in several worker populations (Sakr et al., 2009; Leonard et al., 2008; Lundin et al., 

2009). No consistent associations have been found between hyperglycemia, type II diabetes 

(MacNeil et al., 2009; Lin et al., 2009; Nelson et al., 2010), thyroid homeostasis (Olsen et al., 

2001a,2003; Sakr et al., 2007a; Olsen and Zobel, 2007; Costa et al., 2009; Emmett et al., 2006; 

Bloom et al., 2010), fertility, fecundity, and birth outcome (Fei et al., 2009; Monroy et al., 2008; 

Hamm et al., 2009; Washino et al., 2009; Nolan et al., 2010; Stein et al., 2009). Maternal or 

child plasma PFOA levels were not associated with developmental milestones in children (Fei et 

al., 2008b), child risk of hospitalization for infectious diseases (Fei et al., 2010), or attainment of 

puberty (Christensen et al., 2011). 

 

Animal Data – Long Term Studies. Effects in animal studies are most commonly those 

associated with activation of the PPARα receptor leading to peroxisome proliferation. These 

include increased liver weight, decreases in triglycerides, cholesterol, and lipoprotein, plus 

increases in ALT and/or AST. However, the mechanisms for other effects, such as decreased 

body weight, immunological effects, and developmental delays, are unknown, yet may be 

relevant to human health risk assessment.

 

As an initial step in the dose-response assessment, U.S. EPA identified a suite of animal 

studies with low dose NOAELs and/or LOAELs as potential candidates for development of a 

chronic Reference Dose (RfD). These studies are listed in Table 5-1. The candidate studies 

were selected based on their low dose NOAEL and/or LOAEL, a duration of ≥7 weeks, use of a 
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control, and two or more doses. Table 5-1 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 5-1. 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 Increased relative pituitary weight Goldenthal, 

1978b

Monkey

Female

90 days 3 10 Decreased absolute and relative heart 

weight

Goldenthal, 

1978b

Monkey

Male

26 weeks none 3 Increased absolute liver weight 

(hepatocellular hypertrophy) and mean 

liver-to-body weight percentages

Thomford, 

2001b, 

Butenhoff et al., 

2002

Rat

Male

90 days 0.56 1.72 Hepatocyte necrosis, increased absolute 

liver weight

Goldenthal, 

1978a

Rat

Male

13 weeks 0.06 0.64 Increased absolute and relative liver weight 

with hepatocellular hypertrophy

Palazzolo, 1993; 

Perkins et al., 

2004

Rat

Male

2 years 1.3 14.2 Decreased body weight gain, increased 

kidney weight, and hepatic cystoid 

degeneration, megalocytosis, decreased red 

blood cell count, hemoglobin 

concentration, and hematocrit values, 

testicular vascular mineralization, 

pulmonary hemorrhage

Butenhoff et al., 

2012

Rat Male

F0 

generation

84 days none 1 Increased absolute liver and kidney weight York, 2002; 

Butenhoff et al., 

2004a; York et 

al., 2010

Rat Female

F0 

generation

127 30 none York, 2002; 

Butenhoff et al., 

2004a; York et 

al., 2010

Rat Male

F1 

generation

10 weeks none 1 Decreased body weights and weight gains 

increased absolute and relative liver 

weights

York, 2002; 

Butenhoff et al., 

2004a; York et 

al., 2010

Rat

Female

90 days 22.4 76.5 Increased absolute and relative liver 

weight

Goldenthal, 

1978a

Rat 

Male and 

Female

2 years 1.3 (m)

1.6 (f)

14.2 (m)

16.1 (f)

M: Decreased body weight gain; histopath. 

lesions in liver, testes, and lungs.

F: Decreased body weight gain and 

decreased red blood cell count, 

hemoglobin, concentration, and hematocrit 

values

Butenhoff et al., 

2012

Rat Female

F1 

Generation

10 weeks 10 30 Female: Delay in sexual maturity, 

decreased body weight and weight gain 

York, 2002; 

Butenhoff et al., 

2004a; York et 

al., 2010

When examining the effects associated with the low dose exposures summarized in Table 

5-1, changes in relative and/or absolute liver weight appear to be the common denominator for 
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monkeys and rats (York, 2002; Butenhoff et al., 2004a; Goldenthal, 1978a; Palazzolo, 1993; 

Perkin et al., 2004; Thomford, 2001b, Butenhoff et al., 2002) with or without other hepatic 

indicators of adversity (Palazzolo, 1993; Perkin et al., 2004; Thomford, 2001b, Butenhoff et al., 

2002). Serum PFOA levels, where available, associated with increased liver weight were 81 and 

41 μg/mL for the male monkey and rat, respectively. Body weight effects were seen in several 

studies (York, 2002; Butenhoff et al., 2004a; Butenhoff et al., 2012). Testicular effects were 

observed by Butenhoff et al. (2012) and in the chronic one-dose study by Beigel et al. (2001). 

There were developmental delays for males and females in the two-generation study published 

by York (2002) and Butenhoff et al. (2004a). 

 

 Four of the studies in Table 5-1 lack a NOAEL and have LOAELs that range from 1 to 3 

mg/kg/day. The NOAELs for the remaining 7 studies range from 0.06 to 22.4 mg/kg/day. 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 5-

1 suggests that several data sets have the potential to be co-critical in the dose-response 

evaluation. 

 

Animal Data –Short Term Studies. A number of studies in mice have also identified adverse 

effects following low dose exposures. 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 5-2. 
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TABLE 5-2. Shorter term and Developmental Oral Exposure Studies

Species Study 

Duration

NOAEL

(mg/kg/day)

LOAEL

(mg/kg/day)

Critical Effect(s) Reference

Rat

Rat

Male

29 days 1 10 Increased absolute and relative 

liver weight, focal liver necrosis

Loveless et 

al., 2008

Mouse

Mouse

Male

29 days 0.3 1 Increased absolute and relative 

liver weight, decreased relative 

spleen weight, moderate-severe 

liver hypertrophy with single 

cell and focal necrosis

Loveless et 

al., 2008

Mouse

Male

29 days 1 0.3 Decreased spleen weight Loveless et 

al., 2008

Mouse 

Female

15 days none 0.94 Increased absolute and relative 

liver weight (1 and 15 days post 

dose)

DeWitt et al., 

2008

Mouse

Female

15 days 1.88 3.75 Decreased IgM (1 days post 

dose), increased IgG (15 days 

post dose), decreased absolute 

and relative spleen weight (1 day 

post dose)

DeWitt et al., 

2008

Mouse

Female

17 (pups) 

/18 (dams) 

days

none 1 Increased absolute maternal liver 

weight, reduced ossification 

(calvarin, enlarged fontanel), 

accelerated onset of puberty in 

male offspring

Lau et al., 

2006

Mouse

Female

17 days

GD1-17

none 3 Increased absolute and relative 

maternal liver weight, delayed 

offspring eye opening and body 

hair growth, increased offspring 

relative liver weight, decreased 

offspring body weight, delayed 

mammary gland development 

(female offspring)

Wolf et al., 

2007; White 

et al., 2009

Mouse 

Female

11 days

GD7-17

none 5 Increased maternal and pup 

relative liver weight, delayed 

offspring eye opening and hair 

growth, decreased male 

offspring body weight, delayed 

mammary gland development 

(female offspring)

Wolf et al., 

2007; White 

et al., 2009

As was the case with the longer-term studies, increased liver weight was a common effect 

among the shorter-term studies summarized in Table 5-2. Increases in absolute or relative liver 

weights were reported in six of the eight studies (Lau et al., 2006; Wolf et al., 2007; DeWitt et 

al., 2008; Loveless et al, 2008; Table 5-2). The co-occurring effects at the LOAEL were effects 

on spleen, thymus, liver, and/or developmental endpoints. Four of the eight studies lacked a 

NOAEL. For those studies with a NOAEL, the value ranged from 0.3 to 1.88 mg/kg/day while 

the LOAELs ranged from 0.3 to 10 mg/kg/day. Serum PFOA levels associated with effects in 

female mice ranged from 21.9 – 36.3 μg/mL. In both cases, 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 29 days, suggesting that physiological responses 
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to PFOA occur early in the exposure continuum and at doses comparable to those observed in the 

long term studies. 

5.1.1.1 Benchmark Dose Approach

 As a second step in the dose-response analysis, benchmark modeling of doseresponse for the liver weight changes from the studies wherein this was a critical effect was 

conducted. The data were modeled using benchmark dose software (BMDS) v. 2.1.2 for a 

continuous dataset with modeled variance and restricted slope where applicable. A 10% increase 

in absolute liver weight was chosen as the benchmark response (BMR) for the initial analysis. 

Although it is a biomarker for systemic exposure in rodents rather than a biomarker of adversity, 

it serves as a lowest common denominator for loss of homeostasis in a sensitive species. 

PFOA is not metabolically altered during absorption, distribution and excretion. 

Systemic homeostasis is lost when renal excretion is no longer balanced with absorption. There 

is clear evidence that PFOA has a high affinity for the proteins that compose the PPARα receptor 

of rodents resulting in hypertrophy and increased liver weight as an early signature of exposure. 

Thus, both liver responses serve as biomarkers for the loss of homeostasis and can be 

extrapolated to species other than PPARα sensitive-rodents. The studies of Nakamura et al. 

(2009) and Albrecht et al. (2013) of mice with the humanized PPAR-α receptor did not observe 

significant changes in liver weight. However, there were changes in humanized mice that 

occurred at doses that exhibited the liver weight effects in wild type mice and support its use as a 

biomarker for loss of homeostasis. Specifically, humanized mice exhibited changes in gene 

products that modulate lipid metabolism (Albrecht et al., 2013; Nanamura et al., 2009) at doses 

comparable to those associated with liver weight effects in wild-type mice.

Results of the benchmark analyses are shown in Table 5-3 where goodness-of-fit 

information [p value and Akaike’s Information Criterion (AIC)] was used to choose the best 

model for each data set. Liver weight data were available in the male rat (Butenhoff et al., 

2004a; Palazzolo, 1993; Perkins et al., 2004; York, 2002; and York et al., 2010), male mouse 

(Loveless et al., 2008), and female mouse (Lau et al., 2006). For each data set, the BMD for a 

10% increase in liver weight and the lower-bound confidence limit on the BMD (BMDL) are 

provided. 
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TABLE 5-3. Benchmark Dose Modeling for a 10% Increase in Liver Weight

Species Best fit 

Model

BMD10

mg/kg/day

BMDL10

mg/kg/day

Notes Reference

Rat

Rat

Male

Exponential: 

Model 4

0.798012 0.456224 Increased absolute and 

relative liver weight 

Palazzolo, 1993; 

Perkins et al., 

2004

Rat Male Exponential 

Model 4

0.241376 0.151833 Increased absolute and 

relative liver weight

Loveless et al., 

2008

Rat Male

F0 

generation

Exponential: 

Model 4

0.399087 0.274007 Increased absolute and 

relative liver weights

York, 2002; 

Butenhoff et al., 

2004a; York et 

al., 2010

Rat Male

F1 

generation

Exponential: 

Model 4

0.599409 0.3655 Significant increases in 

absolute and relative liver 

weights

York, 2002; 

Butenhoff et al., 

2004a; York et 

al., 2010

Mouse

Mouse 

Male

Exponential: 

Model 4

0.099268 0.0680743 Increased absolute and 

relative liver weight

Loveless et al., 

2008

Mouse 

Female

Exponential: 

Model 4

0.292244 0.246874 Increased maternal absolute 

liver weight

Lau et al., 2006

 

Data from Thomford (2001b) for monkeys and Goldenthal (1978) for the male rat were 

not presented because the low number of animals (n= 2 to 6) in each dose group resulted in 

highly variable lower bound estimates. An attempt was also made to use the BMDS to model 

liver-to-body weight ratio data from the studies where this information was available. However, 

all models failed with p<0.1 indicating that the fit was not reliable

Taken together, the analyses presented in Table 5-3 identify BMD values that vary by 

less than an order of magnitude (about 0.1 to 0.8 mg/kg/day) for a 10% increase in liver weight

despite differences in species. Likewise, the values for the calculated lower-bound estimates did 

not vary greatly (about 0.1 to 0.5 mg/kg/day) indicating similarity in response between species. 

It is interesting to compare the empirical results presented in Table 5-1 with the BMD model 

results in Table 5-3. For the six data sets modeled, a NOAEL was not identified in four and was 

0.06, 0.3, and 0.56 mg/kg/day in the remaining three studies. The LOAEL was 1 mg/kg/day in 

four studies and 0.64, 1.72, and 3.0 mg/kg/day in the others. The BMDL10 values all fall below 

the experimental LOAELS.

The model results shown in Table 5-3 were chosen based on goodness-of-fit criteria as 

described above. The modeling output for all studies is included in Appendix 1. A subset of the 

modeled data is used in the RfD derivation (section 3.1.1.3). Figure 5.1 provides the graphic 

results from those datasets which were selected as critical for the RfD derivation. Data from the 

F1 generation male rats (York, 2002; Butenhoff et al., 2004a; York et al., 2010) were not used 

because of uncertainties about the timing for the developmental changes in the renal tubular 

transporters that differentiate male rats from female rats (see Section 3.4). The developmental 

toxicity study in female mice (Lau et al., 2006) supports the endpoint of liver weight effects and 

is included in Figure 5-1 despite the limitation posed by its 17-day, relatively short exposure 

duration. 
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Male Rat (Palazzolo, 1993; Perkins et al., 2004) Male Rat (Loveless et al., 2008) 

F0 Male Rat (Butenhoff et al., 2004a; York, 2002; York et 

al., 2010) Male Mouse (Loveless et al., 2008)

Female mouse (Lau et al., 2006)

 FIGURE 5-1. BMDS graphic output from selected model runs.

 16

 18

 20

 22

 24

 26

 28

 30

 0 1 2 3 4 5 6

Mean Response

dose

Exponential Model 4 with 0.95 Confidence Level

07:37 08/18 2011

BMDL BMD

 

Exponential

 12

 14

 16

 18

 20

 22

 24

 0 5 10 15 20 25 30

Mean Response

dose

Exponential Model 4 with 0.95 Confidence Level

13:12 12/22 2011

BMDLBMD

 

Exponential

 20

 22

 24

 26

 28

 30

 0 5 10 15 20 25 30

Mean Response

dose

Exponential Model 4 with 0.95 Confidence Level

10:57 10/26 2011

BMDLBMD

 

Exponential

 2

 3

 4

 5

 6

 7

 0 5 10 15 20 25 30

Mean Response

dose

Exponential Model 4 with 0.95 Confidence Level

10:48 08/24 2011

BMDLBMD

 

Exponential

 2

 2.5

 3

 3.5

 4

 4.5

 5

 5.5

 6

 6.5

 0 5 10 15 20

Mean Response

dose

Exponential Model 4 with 0.95 Confidence Level

15:12 10/25 2011

BMDLBMD

 

Exponential
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5.1.1.2 Pharmacokinetic Model approach

 In linking chemical exposure to toxic endpoints, careful consideration of the 

pharmacokinetics is crucial. This is especially true for PFOA, where inter-species and gender 

variation in clearance half-life can vary by several orders of magnitude. If the toxicological 

endpoints are assumed to be driven by internal concentrations, then it is the internal exposure 

that must be calculated and considered across species. Differences in pharmacokinetics (e.g., 

male rats excrete PFOA less quickly than females) and differences across species produce 

differences in the external dose needed to achieve the same internal dose. 

 

PFOA has been found to have dose-dependent kinetics. Although repeated doses rapidly 

result in quasi-equilibrium blood concentrations, a single dose results in a much longer half-life 

than would be consistent with a rapid approach to quasi-equilibrium (Andersen et al. 2006, Lou 

et al. 2009). Using a simple, linear pharmacokinetic model (e.g. a one- or two-compartment 

model) to predict internal dose resulted in estimated exposures due to repeated doses that were 

greater than actually seem to occur (Butenhoff et al. 2004, Lou et al. 2009). 

 

The saturable renal resorption model (Andersen et al. 2006: Section 3.5.1) has been used 

to describe how PFOA can reach steady state faster than the elimination half-life would indicate. 

The mechanism for this phenomenon is suggested to be the result of saturable resorption from 

the kidney filtrate: Unbound compound in plasma is filtered at the kidney glomerulus and passes 

to the proximal tubules where it is reabsorbed by transporters and re-enters systemic circulation. 

If the amount of compound in the filtrate exceeds the ability of the transporters to reabsorb it, it 

is rapidly excreted so that additional doses do not affect the steady state serum levels. 

In the Lou et al. (2009) report, the one and two compartment models were able to 

reasonably predict the serum levels for single 1 and 10 mg/kg doses, but not at the high dose of 

60 mg/kg/day where excretion exceeded transporter resorption capability and serum values 

declined more rapidly than predicted based on the responses at 1 and 10 mg/kg/day (Section 

3.5.1). Elimination from the body was much faster at the high dose than it was for the lower 

doses. This behavior is consistent with the saturable resorption model: at low doses excretion 

becomes balanced with uptake. Thus, plasma levels remain at steady state as long there is no 

increase in transporter density in the membrane of the renal tubule and dosing is constant. When 

a dose was large enough to exceed resorption capacity, the excretion rate increases but simple 

one- and two-compartment models do not allow for excretion rate changes. In Lou et al. (2009) 

the addition of a saturable resorption component to the model greatly improved the fit for the 60 

mg/kg single dose and 20 mg/kg/day repeat doses, while still describing the 1 and 10 mg/kg 

doses. 

 

The biological basis for the saturable resorption model parameters is still uncertain (Lou 

et al., 2009), particularly with respect to the idea that glomerular filtration alone is responsible 

for the movement of PFOA into the renal filtrate given that PFOA is highly protein bound 

(Kerstner-Wood et al., 2004) and export transporters are located along the apical and basal 

membranes of the renale tubular cells. Protein binding kinetics will impact the levels available 

for loss through glomerular filtration because the amount of free PFOA in serum will vary with 

the saturation of serum protein-binding sites. Some loss of PFOA to the renal filtrate will also 
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occur by way of tubular export transporters. However, the qualitative behavior of the saturable 

resorption model is excellent (Andersen et al., 2006; Lou et al., 2009). 

The saturable resorption model is an empirical model similar to the commonly used oneand two-compartment models. Due to the empirical nature of the saturable resorption 

mechanism present in the Tan et al. (2008) and Loccisano et al. (2011, 2012a,b) models, no 

existing model for PFOA PK allows extrapolation between species. Resorption parameters 

cannot be directly linked to the properties of the relevant serum transport proteins or membrane 

transporters. Therefore the saturable resorption model parameters must be estimated using 

species-specific pharmacokinetic data. 

Pharmacokinetic data (serial blood concentrations following treatment with known 

quantities of PFOA) were collected for three species: cynomolgus monkey (Butenhoff et al., 

2004), Sprague Dawley rat (Kemper et al., 2003), and mouse. Data were available for two 

strains of mouse and were analyzed separately: CD1 (Lou et al., 2009) and C57BL6/N (Dewitt et 

al., unpublished). Due to the pronounced difference in the pharmacokinetics 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. 

 

The available pharmacokinetic studies were not designed with the expectation of nonlinear pharmacokinetics. Accordingly, many parameters associated with saturable resorption 

may be uncertain. For example, the Kemper et al. (2003) studies (Section 3.3) focused upon 

single doses, although the relatively large highest dose (25 mg/kg) may be sufficient to cause 

non-linearities such as that observed for a 60 mg/kg dose in CD1 mice (Lou et al., 2009) 

described above. In the absence of ideal data, the question of parameter values becomes one of 

what range of values is consistent with the data, rather than which single value is most 

consistent. For the monkey, some of the data came from an intravenous administration study in 

males and females (Butanoff et al., 2004b). Considering a range of possible values that might 

equally well explain the data can be addressed with Bayesian statistical analysis. In this case the 

Bayesian analysis was used to determine parameter distributions for all five data sets considered.

 

A non-hierarchical model for parameter values has been assumed; i.e. there is a single 

value shared by all individuals of the same species/strain and gender (except for rat, where there 

is a clear gender difference in renal excretion of PFOA). Body-weight and treatment (number 

and magnitude of doses) are the only parameters that may vary between individuals. However, 

with the exception of the Kemper et al. (2003) data, individual bodyweights were not available 

so that a default bodyweight was used for each species and gender. In model development it was 

assumed that animals were sacrificed at the end of the external exposure; i.e. there was no 

continued internal exposure after dosing had stopped. 

 

Bayesian analysis allows formal inclusion of prior knowledge in the form of set 

distributions on the parameters being estimated (Gelman et al., 2004). Given that empirical 

pharmacokinetic parameters can have a wide range of values, vague, bounded prior distributions 

are appropriate (Wambaugh et al., 2008). For all estimated parameters the prior knowledge 

values were assumed to be distributed log-normally. This constrained the parameters to positive 

values. The mean and variances assumed for the CD1 mice, male and female Sprague-Dawley 
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rats, and Cynomolgus monkeys are given in Table 5-4. For the C57BL/6N mice the available

data were insufficient to achieve a converged statistical analysis, so the posterior parameter 

distributions for the CD1 mouse were used, with the variances increased ten-fold. 

 

TABLE 5-4. Description of prior distributions used to analyze all but the C57BL/6N mice

Bounds

Parameter Mean Variance Lower Upper

ka 1 1000 1/6 105

Vcc 1 1000 10-10 105

k12 1 1000 10-10 105

RV2:V1 1 0.5 10-10 100

Tmaxc 1 1000 10-10 105

kT 1 1000 10-10 105

free 0.01 0.5 10-10 1

Qfilc 1 1000 0.01 105

Vfilc 1 1000 10-10 10

Parameters were log-normally distributed with the mean and variance listed. Bounds were used to reduce the time spent sampling 

in areas thought to have low probability, and were expanded when large posterior mass at the bounds indicated that the bounds 

were too narrow. For the C57BL/6N the posterior parameter distributions for the CD1 mouse was used, with the variances 

increased ten-fold. 

The deep tissue compartment of the Andersen et al. (2006) model is characterized in 

terms of the rates to and from that compartment (k12 and k21, respectively). This corresponds to a 

volume of distribution V2 = k12*V1/k21, so that the ratio of the volume of the second 

compartment to the first is equal to the ratio of k12 to k21 (Rv2:v1 = k12/k21). In order to enforce the 

assumption that the primary (serum) compartment contains a significant portion of the PFOA, it 

was assumed that the volume of the deep tissue compartment was constrained to be no more than 

100 times greater than the volume of distribution of the serum. For this reason the ratio of the 

two volumes is estimated, rather than the rate from the second compartment to the first 

compartment. The rate of flow from the deep tissue back to the serum was calculated as k21 = 

k12/Rv2:v1. 

Bayesian analysis was performed using Markov Chain Monte Carlo (Gelman et al. 2004). 

The crux of a Bayesian analysis is the assessment of whether the distributions of values in the 

Markov Chain reflect the posterior distribution of the statistical model given the priors and the 

data. The distribution of parameter values were considered to “burned in” (i.e., true draws from 

the posterior combination of the prior assumptions and the available data which were

independent of the assumed starting values) when they passed the Heidelberger and Welch 

Stationarity test (Heidelberger and Welch 1983) as implemented in the Coda Package (Best et al.

1995).

For all pharmacokinetic parameter values (Table 5-5) a mean and a 95% interval are 

reported, summarizing the distribution of parameter values found to be consistent with the 

available data. Parameters such as the flow to and volume of the filtrate compartment were found 

to be very uncertain as in Anderson et al. (2006) and Lou et al (2009). Between male and female 

rats parameters were largely similar; including the affinity of the putative resorption transporters 

but the maximum resorption transport rate was nearly 200 times greater for the males. The 

median fraction of blood flow to the filtrate (Qfilc) was physiologically reasonable (less than or 

equal to the fraction of blood flow to the kidney) in all cases. In contrast the Lou et al. (2009) 

model allowed an unreasonably high portion of cardiac output to pass through the kidney in 
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order to optimize the fit to the data. Although not totally based on biology, the form of the 

saturable resorption model seems to be an appropriate framework for the projection of internal 

PFOA serum dose. 

 

TABLE 5-5. Estimated and assumed pharmacokinetic parameters used in model development

Parameter Units CD1 Mouse

Female

(Lau et al., 

2009)

C56BL/6N 

Mouse

Female

(DeWitt et al. 

unpublished)

Sprague 

Dawley Rat

Female

(Kemper et 

al., 2003)

Sprague 

Dawley Rat 

Male

(Kemper et 

al., 2003)

Cynomolgus 

Monkey male 

and female

(Butenhoff et 

al., 2004)

Body 

Weighta 

kg 0.025 0.025 0.20 (0.16 –

0.23)a

0.24 (0.21 –

0.28)a

7 (m) 4.5 (f)

Cardiac 

outputb 

L/h/kg0.7

4

8.68 8.68 12.39 12.39 19.8

ka 1/h 290 (0.6 –

73000)

340 (0.53 –

69000)

1.7 (1.1 –

3.1)

1.1 (0.83 –

1.3)

230 (0.27 –

73000)

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 1/h 0.021 

(3.1x10-10 – 

3.8x104)

0.35 (0.058 –

52)

0.098 (0.039 

– 0.27)

0.028 

(0.0096 – 

0.08)

0.0011 

(2.4x10-10 – 

3.5x104)

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 M/h 4.91 (1.75 –

2.96)

2.7 (0.95 – 22) 1.1 (0.25 –

9.6)

190 (5.5 –

50000)

3.9 (0.65 –

9700)

kT M 0.037 (0.0057 

– 0.17)

0.12 (0.033 –

0.24)

1.1 (0.27 –

4.5)

0.092 

(3.4x10-4 – 

1.6)

0.043 (4.3x10-

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 L/h 0.077 (0.015 

– 0.58)

0.017 (0.010 – 

0.081)

0.039 (0.014 

– 0.13)

0.22 (0.011 

– 58)

0.15 (0.02 – 

24)

Vfilc L/kg 9.7x10-4

(3.34x10-9 – 

7.21)

7.6x10-5

(2.7x10-10 – 

6.4)

2.6x10-5

(2.9x10-10 – 

28)

0.0082 

(1.3x10-8 – 

7.6)

0.0021 

(3.3x10-9 – 

6.9)

Estimated and assumed pharmacokinetic parameters for the Andersen et al. (2006) model. Means and 95% interval from Baysian 

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). As part of the Bayesian formalism, it is the posterior 

distribution of parameter values, as opposed to the mean values, that is used to make the predictions.

a Estimated average bodyweight for species used except with Kemper study where individual rat weights were available and 

assumed to be constant.

b Cardiac outputs obtained from Davies and Morris (1993)

 

If there is large uncertainty for parameters that greatly influence the predicted value, then 

the distribution of predicted values will be wide. If the quantity being predicted is not sensitive 

to the uncertain parameter, then the distribution of the predicted value will be unaffected 

(Wambaugh et al. 2008). Assessing parameter uncertainty allows appropriate confidence in 

model predictions. 

 

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 8-6 for rats, 8-7 for mice, and 8-8 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 
Perfluorooctanoic Acid – February 2014 5-13 

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predicted final serum concentration is the estimate for serum concentration at the time of 

sacrifice. Generally these values were similar. They differed by a factor of four when strains 

were different and closer to a factor of two when predicting using parameters from the same 

strain. Because these predictions do not perfectly match the measured serum concentrations, 

there remains uncertainty about the exposure estimates that has not been fully characterized.

Numbers in parentheses indicate standard deviation

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.

TABLE 5-6. 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 

Palazzolo, 

1993; 

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)

7310 (188)

69900 (1640)

170000 (6770)

331000 (27100)

York, 

2002; 

Butenhoff 

et al., 

2004a; 

York et 

al., 2010

Rat (M)

SpragueDawley

F0M: 10 

wk pre 

matingmating

Oral 

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)

93500 (2460)

207000 (11000)

351000 (35100)

422000 (74600)
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TABLE 5-7. Predicted Final Serum Concentration and Time-integrated Serum Concentration (AUC) for 

Studies in Mice

Study Species/ 

Strain

Exposure 

Duration

Oral 

Doses

mg/kg/d

Measure

d Serum

Value 

μg/ml

Species/ 

Strain 

Used for 

Prediction

Predicted 

Final 

Serum 

Value 

μg/ml

Predicted 

AUC mg/L*h

Wolf et 

al., 2007; 

White et 

al., 2009

Mouse 

(F)

CD-1 

GD1-17a

Oral 

gavage

3

5 

NT

NT

Mouse (F)

CD-1 

25 (2.22)

25.6

(2.26) 

31700 (1740)

38600 (2050) 

GD7-17

Oral 

gavage

5 24.8 Mouse (F)

CD-1 

29 (2.55) 23300 (1160)

DeWitt et 

al., 2008

Mouse 

(F)

C57BL/6

N 

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)

7320 (557)

13800 (952)

22400 (1330)

30500 (1520)

40100 (4470)

55900 (12000)

Lau et 

al., 2006

Mouse 

(F)

CD-1 

GD1-17

Oral 

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)

16500 (1010) 

33600 (2690) 

40900 (2950) 

51100 (2810) 

65400 (6480) 

90400 (17500) 

Numbers in parentheses indicate standard deviation

NA = not applicable; could not be determined

F=female; GD = gestation day; s=serum; NT=not tested

a

Sacrificed on PND 22.

b

DeWitt (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. 
Perfluorooctanoic Acid – February 2014 5-15 

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TABLE 5-8. 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

Thomford, 

2001b, 

Butenhoff 

et al., 

2002, 

2004b

Monkey 

(M)

Cyanomolgus

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)

387000 (50800)

565000 (61100)

710000 

(144000)

Numbers in parentheses indicate standard deviation

M= male

The AUC for the LOAEL or NOAEL of each data set can be used to determine an 

average serum concentration by dividing it by the duration of the study in days with adjustment 

for the hours in a day. The following equation is used for the conversion: 

Average Serum Concentration = AUC (mg/L*h) x 1 day/24 hours ÷ exposure duration 

(days) 

For example, in a case where the AUC was 30,000 mg/L*h and the study duration was 90 

days, the Average Serum Concentration would be calculated as follows: 

Average Serum Concentration = 30,000 mg/L*h ÷ (90 days x 24 h/day) = 13.89 mg/L 

This calculation 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 5-9 provides the AUC from the model, the dosing duration from each of the 

modeled studies, and the resultant average serum concentration. The internal doses associated 

with the effect levels (LOAELs) differ by less than an order of magnitude (20.3 mg/L to 88.6 

mg/L) while the AUC values differ by over two orders of magnitude (7320 mg/L*h to 387000 

mg/L*h). Given the differences in external doses, the projected serum levels are proportionally 

quite similar.
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TABLE 5-9. Average Serum concentrations Derived from the AUC and the duration of Dosing

Study Dosing 

duration 

days

NOAEL 

(AUC

mg/L*h)

NOAEL 

(Av serum

mg/L)

LOAEL

(AUC

mg/L*h)

LOAEL

(Av serum 

mg/L)

Dewitt, 2008 liver 15 None None 0.94

(7320)

20.33

Dewitt, 2008 

Immunotox

15 1.88

(13800)

38.33 3.75

(22400)

62.22

Lau, 2006 liver 17 None None 1 

(16500)

40.44

Palazzolo, 1993 

Liver

91 0.06 (7310) 3.35 0.64 

(69900)

32.00

Thomford, 2001b; 

Butenhoff et al., 

2004 Liver

182 None None 3 (387000) 88.60

Wolf, 2007 gd 7-

17 

Liver

11 None None 5

(23300)

88.26

Wolf, 2007 gd 1-

17 

Liver

17 None None 3 

(31700)

77.70

York, 2002

liver

84 none none 1 

(93500)

46.38

Table 5-9 identifies 20.33 mg/L as the lowest concentration associated with an increase in 

liver weight in female mice (DeWitt et al, 2008). Average serum values for increased relative 

liver weight in male rats (Palazzolo et al., 1993) and female mice (Lau et al. 2006) are no more 

than 2-fold higher than the average serum value in DeWitt et al. (2008). Thus, it appears that the 

LOAELs are roughly consistent across gender, species, and treatment with respect average serum 

concentration. Assuming that mode of action and susceptibility to toxicity do not vary and that 

pharmacokinetics alone explains variation, it is reasonable to expect similar concentrations to 

cause similar effects in humans.

The predicted serum concentrations can be converted into an oral equivalent dose at 

steady state by recognizing that, at steady state, clearance from the body must equal dose to the 

body. Clearance can be calculated if the rate of elimination (derived from half-life) and the 

volume of distribution are both known. 

Measures of half-life in humans have been determined for both workers and the general 

population (Section 3.5.2). 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, WV after the drinking 

water concentrations decreased. The Office of Water has chosen to use the Bartell et al. (2010) 

half-life value because it is appears to be 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 volume of distribution of 0.17 L/kg bw (Section 3.5.3). 

The volume of distribution is defined as the total amount of PFOA in the body divided by the 

blood or serum concentration. The volume of distribution was calibrated using human serum 

concentrations and exposure data from NHANES and assumes that most PFOA intake came 
Perfluorooctanoic Acid – February 2014 5-17 

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from contaminated drinking water. The value for volume of distribution was calibrated so that 

the model prediction of elevated blood levels of PFOA was consistent with the values from 

NHANES. Lack of consideration of PFOA contributions from sources other than drinking water 

sources is a weakness of the volume of distribution value determined by Thompson et al. (2010), 

however this estimate is not radically different from the 0.198 L/kg bw determined for the 

monkey in the study by Buttanhoff et al (2004). 

The half life (t1⁄2) and volume of distribution (Vd) are utilized to calculate the clearance for 

PFOA according to the following equation assuming first order kinetics for clearance (CL) 

(Medinsky and Klaassen, 1996): 

CL = Vd x (ln 2 ÷ t1/2) = 0.17 L/kgbw/day 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 clearance of 1.4 x 10-4 L/kg BW/day. 

Scaling the derived average concentrations (in mg/L) for the NOAELs and LOAELs in 

Table 5-10 gives predicted oral human equivalent doses (HEDs) 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 human kinetics in 

contrast to the non-linear phenomena observed at high doses in animals. 

Thus, HED = average serum concentration (in mg/L) x CL 

Where:

 Average serum is from model output in Table 5-9 

 CL = 0.00014 L/kg bw/day 
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TABLE 5-10. Human Equivalent Doses 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, 2008 

liver

15 None - - 0.94 20.33 0.0028

Dewitt, 

Immunotox

15 1.88 38.33 0.0054 3.75 62.22 0.0087

Lau, 2006 liver 17 None - - 1 40.44 0.0057

Palazzolo, 1993 

Liver

91 0.06 3.35 0.00047 0.64 32.00 0.0045

Thomford, 

2001b liver

182 None - - 3 88.60 0.0124

Wolf, 2007 GD 

1-17

liver

17 None - - 3 77.70 0.0109

Wolf, 2007 gd 

7-17 

Liver

11 None - - 5 88.26 0.0124

York 2002 

male F0 liver

84 None - - 1 46.38 0.0065

5.1.1.3 RfD Quantification

There are several acceptable points of departure (PODs) that can be used in the process of 

identifying the POD for RfD development: 

 

• NOAEL or LOAEL values 

• Lower 95% confidence bounds on the BMD (BMDLs), and 

• Human Equivalent Doses (HED). 

Studies that have more than one POD for the same NOAEL or LAOEL are summarized 

in Table 5-11. There are very few NOAELs that have been identified for PFOA; LOAELs are 

the more common benchmark identified through classical toxicological studies. Some of the 

studies with NOAELs and/or LOAELs are chronic and others are short term. Modeling of dose 

response to identify a BMD and BMDL was successful for only a subset of studies. All 

benchmark models targeted a 10% increase in liver weight as a common metric for effect. This 

selection was not made based on toxicity but on the desire to find a common denominator against 

which to evaluate dose-response across studies and justified by the fact that other adverse effects 

accompanied the LOAEL for increased liver weight in some cases. 

The subset of studies that were amenable for derivation of HED based on average serum 

measurements from the pharmacokinetic model was limited because of the need to have dose and 

species-specific serum values for model input. Most entries in Table 5-11 provide at least two 

potential PODs for consideration; only one provides five values (Palazzolo et al. 1993). 
Perfluorooctanoic Acid – February 2014 5-19 

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TABLE 5-11. RfD Point of Departure Options (mg/kg/day) from the PFOA Animal Studies

Studies NOAEL LOAEL BMDL10 HED

NOAEL

HED

LOAEL

Notes

Dewitt, 2008 

female mouse

none 0.94 0.0028 Increased absolute and 

relative liver weight; 

Dewitt, 2008 

female mouse

1.88 3.75 0.0054 0.0087 Decreased IgM, increased 

IgG, decreased absolute 

and relative spleen 

weights; 

Lau, 2006, GD 

1-17, mouse 

none 1 0.247 0.0057 Increased maternal liver

weight

Loveless et al., 

2008 male rat

1 10 0.152 Increased liver weight, 

focal liver necrosis

Loveless et al., 

2008 mouse 

male

0.3 1 0.0681 Increased liver weight, 

liver hypertrophy with 

single cell and focal 

necrosis

Palazzolo et al, 

1993 male rat

0.06 0.64 0.456 0.00047 0.0045 Increased liver weight, 

hepatocellular 

hypertrophy

Wolf, 2007 gd 

7-17 

Liver

None 5 0.0109 Increased maternal 

relative liver weight 23 

days after last dose, 

reduced male pup body 

weight at birth

Wolf, 2007 gd 

1-17 

none 3 0.0124 Increased maternal 

absolute and relative liver 

weight 23 days after last 

dose

York 2002 

male F0

none 1 0.274 0.0065 Increased liver and kidney 

weight

 

As explained previously, human data have identified significant relationships between 

serum levels and specific indicators of adverse health effects but lack 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 must be accommodated 

in deriving the RfD. 

Each of the POD values represented in Table 5-11 requires a different quantification 

approach. Thus, EPA has systematically examined the impact of POD on outcome through three 

sets of calculations as follows:

• Derivation from 5 PODs: NOAEL, LOAEL BMDL10, HEDNOAEL, and HEDLOAEL

from the same study (Palazzolo et al., 1993) 

• Derivation from the BMDL10 values 

• Derivation from HED values derived from modeled average serum values. 

The Palazzolo et al. (1993) study is the only one among the studies selected for 

consideration as candidates for RfD determination that offers the opportunity to examine the 

impact of five PODs on outcome. It is the only study with a NOAEL, LOAEL, BMDL10, and 
Perfluorooctanoic Acid – February 2014 5-20 

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HED values. Thus, it is ideally suited for use in examining the impact of quantification approach 

on the RfD outcome. The critical effect was a significant increase in liver weight accompanied 

by hepatic hypertrophy and increased palmitoyl Co A oxidase activity. As described earlier, the 

increase in liver weight was selected as a low dose common biomarker for systemic loss of 

homeostasis across studies. In other studies, adverse reproductive (Lau et al., 2006) or 

immunological effects in mice (Loveless et al., 2008) co-occurred in combination with the 

increased liver weights at the LOAEL justifying selection of the liver as a common metric. 

Table 5-12 provides the potential RfDs derived from the five individual Palazzolo et al 

(1993) PODs.

 

TABLE 5-12. The Impact of Quantification Approach on the RfD outcome for the PODs from the Palazzolo 

et al (1993) Study

POD POD dose 

mg/kg/day

UFH UFA UFL UFS UFD UFtotal Potential 

RfD

mg/kg/day

NOAEL 0.06 10 219 - 10 - 21400 0.000003

LOAEL 0.64 10 219 10 10 - 214000 0.000003

BMDL10 0.456 10 219 - 10 - 21400 0.00002

HEDNOAEL 0.00047 10 3 - - - 30 0.00002

HEDLOAEL 0.0045 10 3 10 - - 300 0.00002

There is good agreement between the adjusted PODs from the HEDs from the modeled 

average serum levels and the BMDL10 value. The adjusted PODs derived from the NOAEL and 

LOAEL are about an order of magnitude more conservative than those from the BMDL10 and 

HEDs. The differences in outcome reflect the differences in the quantification approach rather 

than the experimental data utilized. The BMDL10 value more closely approximates the study 

LOAEL rather than the study NOAEL. 

Uncertainty Factor (UF) Application. 

UFH A ten-fold adjustment is assigned to account for intrahuman variability and applied for all 

PODs. 

UFA. Determination of the interspecies uncertainty factor for the NOAEL, LOAEL and 

BMDL10 requires application of the equation for first order kinetics in order to determine the 

pharmacokinetic adjustment associated with differences in half-life between humans and male 

rats. The equation that describes first order kinetics is as follows (Medinksy and Klaassen, 

1996): 

CL = Vd x (ln 2 ÷ t1/2) 

Where: 

 Vd = 0.17 L/kg 

 Ln 2 = 0.693 

Half-life = 11.5 days for rats and 839.5 days for humans 

CLrat = 0.17 L/kg x (0.693 ÷ 11.5 days) = 0.01024 L/kg/day 
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CLhuman = 0.17 L/kg ÷ (0.693 ÷ 839.5 days) = 0.00014 L/kg/day 

The ratio of CLrat to CLhuman (0.01024 L/kg/day ÷ 0.00014 L/kg/day = 73) is used as the 

pharmacokinetic adjustment for differences between species. The total UFA requires an 

additional 3-fold factor for species differences in pharmacodynamics (73 x 3 = 219). 

UFL A ten-fold UF was applied for the LOAEL and HEDLOAEL to account for using an effect 

level as the POD. A UF of 1 was used for the BMDL10 and the NOAEL-derived PODs 

following Agency policies. 

 

UFS A ten-fold factor was applied to the NOAEL, LOAEL, and BMDL10 PODs to account for 

less than lifetime exposure. The HEDs reflect steady state serum values from the 

pharmacokinetic model and reduce the UFS to 1 for the HED PODs. 

UFD In all cases the uncertainty factor for the strength of the database (UFD) is 1. The data base 

for oral PFOA exposure studies is essentially complete although mechanistic questions relative 

to the MOA have not yet been fully elucidated. 

BMDL10 PODs. The lowest, species-specific benchmark dose values in Table 5-11 are those 

from the Loveless et al. (2008) study in the male mouse and male rat. As was the case with the 

Palazzolo et al. (1993) study, the effect modeled was a 10% change in liver weight. In this case, 

the LOAELs from the Loveless et al. (2008) publication for both species included co-critical 

focal hepatic cellular necrosis as well as decreased spleen weight for the mouse. The BMDL10

for the mice was lower than that for the rats (Table 5-11).

 The data for the York et al. (2002) and Palazzolo et al. (1993) are included in Table 5-13 

for comparison with the Loveless et al. results in the rat. In York et al. (2002), the change in 

liver weight was accompanied by a change in kidney weight. A UFL of 10 was applied to 

account for the intraindividual variability. 

 

TABLE 5-13. The Impact of Quantification Approach on the RfD Outcome for PODs from the BMDLs

POD Dose

mg/kg/day

UFH UFA UFL UFS UFD UFtotal Potential 

RfD

mg/kg/day

BMDL10 rat Loveless 0.152 10 219 - 10 - 21900 0.000007

BMDL10 mouse Loveless 0.0681 10 150 - 10 - 15000 0.000005

BMDL10 rat Palazzolo 0.456 10 219 - 10 - 21900 0.00002

BMDL10 rat York 0.274 10 219 - 10 - 21900 0.00001

The BMDL10 values from the Loveless et al. (2008) 29-day studies are lower than those 

from the longer-term Palazzolo et al. (1993) and York et al. (2002) studies but differ by a 

maximum factor of 4 which is relatively small when considering the magnitude of the total 

uncertainty factors.

Uncertainty Value Application

The UFH, UFL, UDS and UFD values are assigned as described for the Palazzolo et al. (1993) data 

in Table 5-11. 
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The UFA for the rat component of the Loveless et al (2008) study is the same as that described 

above for the Palazzolo et al (1993) study. It is necessary to derive the pharmacokinetic 

component of UFA for the mouse and monkey using half-life and volume of distribution to 

calculate the clearance ratio as was done for the rat:

Mouse 

CLmouse = 0.17 L/kg x (0.693 ÷ 15.6 days) = 0.00697 L/kg/day 

CLhuman = 0.17 L/kg x (0.693 ÷ 839.5 days) = 0.00014 L/kg/day 

The ratio of CLmouse to CLhuman (0.00697 L/kg/day ÷ 0.00014 L/kg/day = 50) is used as the 

pharmacokinetic adjustment for differences between species. Thus, the total UFA adds a 3-fold 

factor for species differences in pharmacodynamics between mice and humans (50 x 3 = 150). 

HED PODs. The HEDs derived from Palazzolo et al. (1993), Dewitt et al. (2008), Lau et al. 

(2006), Thomford (2001), and York et al. (2002) were each examined as the potential basis for 

the RfD. Only two of these studies, Palazzolo et al. (1993) and Dewitt et al., (2008), contained a 

NOAEL from which the HED could be derived. Because the immunological effects which were 

the basis of the NOAEL in the Dewitt et al. study occurred at a dose about four-fold higher than 

the effects on liver weight, this value is not included in Table 5-14. The Dewitt et al. (2008) 

LOAEL HED based on increased liver weight has the lowest quantitative values from among 

these studies. LOAEL HEDs from the other studies are about twice the Dewitt et al. (2008) 

value. The Lau et al. (2006) LOAEL HED included reduced fetal ossification; accelerated 

puberty in male offspring as co-critical effects. The York et al. (2001) study included increased 

kidney weight as co-critical at the LOAEL and the Palazzolo et al. (1993) study reported hepatic 

hypertrophy and increased palmitoyl Co A oxidase activity. 

TABLE 5-14. The Impact of Quantification Approach on the RfD Outcomes for the HEDs from the 

Pharmacokinetic Model Average Serum Values

POD Value

mg/kg/day

UFH UFA UFL UFS UFD UFtotal Potential 

RfD

mg/kg/day

PK-HEDNOAEL Palazzolo 0.00047 10 3 - - - 30 0.00002

PK-HEDLOAEL Palazzolo 0.0045 10 3 10 - - 300 0.00002

PK-HEDLOAEL Dewitt 0.0028 10 3 10 - - 300 0.000009

PK-HEDLOAEL Lau 0.0057 10 3 10 - - 300 0.00002

PK-HEDLOAEL York 0.0065 10 3 10 - - 300 0.00002

PK-HEDLOAEL Thomford 0.0124 10 3 10 - 300 0.00004

The outcomes for potential RfD values as determined in Tables 5-11, 5-12, and 5-13 

differ by about an order of magnitude (0.000003 mg/kg/day to 0.00004 mg/kg/day). Of the 

values derived from the HED values one is 0.000009 mg/kg/day (Dewitt et al., 2008) and 4 are 

0.00002 (Palazzolo et al., 1993; Lau et al., 2006; York et al., 2002). These results demonstrate 

the ability of the model to normalize the animal data across species, strain, gender, and exposure 

duration. 
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 The Palazzolo et al. (1993) and York et al. (2002) studies were conducted in male 

Sprague Dawley rats with durations of 91 days via diet and 84 days via gavage, respectively. 

The Lau et al. (2006) study was conducted in pregnant female rats with an 18-day average 

exposure via gavage leading to increased liver weights in the dams and developmental effects in 

their male offspring. The Dewitt et al. (2008) drinking water study was conducted in female 

C57BL/6 mice with a 15-day exposure. The Thumford (2001b) study was conducted in male 

cynomolgus monkeys for 182 days; the PFOA was administered in capsules. 

Uncertainty Value Application

The UFH, UFL, UFS and UFD values are assigned as described for the HED LOAEL entry for the 

Palazzolo et al. (1993) data in Table 5-12. 

The UFA is 3 for each study because the HED was derived using the steady state serum values 

from the pharmacokinetic model. The 3-fold factor is applied to account for toxicodynamic 

differences between the animals and humans. 

 

RfD Selection Based on the consistency of the response and with recognition of the use of liver 

weight having been chosen as a common denominator in the RfD analysis for loss of 

homeostatisis and protection against co-occurring adverse effects, the 0.00002 mg/kg/day 

outcome is selected as the RfD for PFOA. This value is the outcome for all modeled rat and 

mouse serum values except for the Dewitt et al. (2008) 15-day study with an impact on liver 

weight but not the co-monitored immunological effects. The liver endpoint in the Lau et al. 

(2006) and York et al. (2002) studies were accompanied by developmental effects and effects on 

kidney weights, respectively. The modeled serum value from Thumford (2001) based on liver 

effects in the monkey, also strongly supports the chosen RfD.

 The results from the modeled serum values are supported by the benchmark dose results 

wherein the Palazzolo et al. (1993) POD supports a potential RfD of 0.00002 mg/kg/day for 

change in liver weight. Other than Lau et al. (2006) all of the studies supporting an RfD of 

0.00002 mg/kg/day are studies with exposures to PFOA for ≥ 84 days meeting the duration 

requirements for determination of a lifetime exposure value. Confidence in the RfD is medium 

because of the uncertainties surrounding mode of action for some of the observed effects and the 

role of saturable renal resorption in determining the body burden from chronic exposures. 

5.1.2 RfC Determination

Derivation of the RfC requires evaluation of toxicity data for concentration-response and 

identification of target organs and noncancer effects. 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 meet the minimal requirements for derivation of the 

RfC as defined by US EPA (1994). 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 
Perfluorooctanoic Acid – February 2014 5-24 

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animal toxicological database. While inhalation is an important route of exposure to workers, 

drinking water was identified as the primary contributor to exposure in the general population. 

In all of the epidemiology studies, wide ranges of serum levels were reported which 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 dose-response via inhalation

Animal Data. Inhalation toxicity data in laboratory animals were limited to acute exposure, 

single and repeated exposures for pharmacokinetic 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 three 

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 six hours in males compared to just one 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 limit of quantitation (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 for use in quantitative risk assessment.

In a developmental toxicity study, pregnant Sprague-Dawley rats were exposed wholebody 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. 

5.2 Dose-Response for Cancer Effects

As discussed in section 4.4.4, there is some evidence that PFOA exposure may be 

associated with an increased risk for cancer from the human epidemiology database and animal 

studies. The Dupont (2003b) report on the West Virginia Washington Works Plant found 

statistically significant elevations in cancer of the bladder, kidney, and urinary tract organs and a 
Perfluorooctanoic Acid – February 2014 5-25 

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nonstatistically significant elevation in leukemia incidence. The findings for an increased 

incidence of leukemia from two earlier worker studies were statistically significant, but both had 

methodological limitations (Walrath and Burke, 1989; Karns and Fayerweather, 1991). The 

mortality data for liver cancer, pancreatic cancer, prostate cancer, and testicular cancer from 

human epidemiological studies all showed no association to exposure. 

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 

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 males and females; the Biegel et al. study only evaluated males. The tumor types 

observed were: 

• Liver (Butenhoff et al., 2012) 

• Leydig Cell (Butenhoff et al., 2012; Biegel et al., 2001) 

• 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) study are summarized in Table 5-15, below. The data are limited 

in that only Butenhoff et al. tested more than one dose and only one tumor-type (Leydig Cell 

adenoma) demonstrated a dose-response relationship.

TABLE 5-15. Summary of Tumor Data from Animal Studies

Tissue Concentration in Diet 

(ppm)

Tumor Type Reference

0a 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 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 

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, mitigate concern for their relevance to humans.

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. 
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PPARα activation may 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 be a consequence of 

the following sequence of events: 

• Decreased testosterone synthesis

• Increased GnRH and increased levels of leutinizing hormone 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). 

 

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 elevated testosterone levels or intestinal tissue hormones (cholecystokinin 

and/or gastrin) in promoting proliferation of acinar cell preneoplastic foci (Osboourn et al., 1997; 

Klaunig et al., 2003). 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. Minimal data for an effect on bile composition are 

available for the PPARα activating agent Wyeth-14,643 (Osborn et al., 1997). 

 Under the EPA 2005 cancer guidelines the evidence for the carcinogenicity of PFOA is 

considered suggestive because only one species has been evaluated and the tumor responses 

occurred primarily in males. Dose-response data are only available for the Leydig Tumors in 

one study. In cases where the evidence is suggestive, the Agency generally does not attempt 

dose-response quantification. However, if the evidence includes a well-conducted study with 

adequate data to support modeling of dose-response, quantitative analyses may be useful in

providing a sense of the magnitude and uncertainty of potential risks, ranking potential hazards, 

or setting research priorities. 

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 Leydig cell tumors in 

male rats in the Butenhoff et al. (2012) study which was confirmed by the high dose in the single 

dose mechanistic study by Biegel et al. (2011). 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. (2012) study and 4% at the low dose (1.3 mg/kg/day). In the Biegel et al (2001) study 11% 

were affected at a dose of 300 ppm in the diet (13.6 mg/kg/day). In each case there were no 

Leydig cell tumors in the controls. The PACT tumors, only detected in the single dose Biegel et 
Perfluorooctanoic Acid – February 2014 5-27 

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al (2001) study, do not support quantification. 

EPA modeled the dose-response for the Leydig cell tumors from Butenhoff et al. (2012) 

using the Agency benchmark dose software (BMDS) v 2.3.1. The multistage cancer model was 

used for a dichotomous dataset to predict 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 

study results. Both the first and second degree polynomials gave identical goodness-of-fit 

criteria (p value and AIC). Results are shown in Table 5-16. 

TABLE 5-16. Model Results for Leydig Cell Tumors (Buttenhoff et al. (2012)

BMD (mg/kg/day) BMDL (mg/kg/day)

4% increase 3.51 1.99

4% increase 3.51 1.99

AIC = 62.6936 P = 0.2245

FIGURE 5-2. BMD Model Results for Leydig Cell Tumors (Buttenhoff et al., 2012)

The Cancer Slope Factor (CSF) for PFOS is is derived from the BMDL04 of 1.99 

mg/kg/day after converting the animal BMDL to a Human Equivalent Dose (HED) using body 

weights the 3⁄4 power. The HED is calculated as follows: 

HED = Animal BMDL x (animal body weight)1/4 ÷ (human body weight)1/4 

HED = 1.99 mg/kg/day x [(0.523 kg1

)

1/4 ÷ (70 kg)1/4] = 1.99 mg/kg/day x 0.29 = 0.58 

mg/kg/day 

 

1 Body weight for male Sprague Dawley rats (chronic Exposures) U.S. EPA, 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
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Where:

1.99 mg/kg/day = BMDL04 for Leydig Cell tumors

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

The CSF should not be used at doses greater than 0.58 mg/kg/day which is the human 

equivalent dose corresponding to the point of departure for the 4% incidence of Leydig Cell 

tumors follow lifetime exposure to PFOA. The observed dose-response relationships do not 

continue linearly above this level and the fitted dose-response models better characterizes the 

dose-response for the higher exposures. The HED of 0.58 mg/kg/day is nearly 30,000-fold 

greater than the RfD (0.58 mg/kg/day ÷ 0.00002 mg/kg/day = 29,000). Thus, the proposed RfD, 

based on increased liver weight as a common denominator for loss of homeostasis and protection 

of co-occurring effects, will also be protective of Leydig cell tumors. 
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APPENDIX A

Contents: Benchmark Dose Modeling Output Files 10% increased liver weight for:

1) Palazzolo, 1993; Perkins et al., 2004: male rat; 

2) Loveless et al., 2008: male rat; 

3) York, 2002; Butenhoff et al., 2004a; York et al., 2010: male rat; 

4) Loveless et al., 2008: male mouse; 

5) Lau et al., 2006: female mouse 
Perfluorooctanoic Acid – February 2014 A-2 

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1) Palazzolo, 1993; Perkins et al., 2004: male rat 

==================================================================== 

 Exponential Model. (Version: 1.7; Date: 12/10/2009) 

 

 Thu Aug 18 07:37:14 2011 

 ==================================================================== 

 BMDS Model Run 

~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~ 

 

 The form of the response function by Model: 

 Model 2: Y[dose] = a * exp{sign * b * dose} 

 Model 3: Y[dose] = a * exp{sign * (b * dose)^d} 

 Model 4: Y[dose] = a * [c-(c-1) * exp{-b * dose}] 

 Model 5: Y[dose] = a * [c-(c-1) * exp{-(b * dose)^d}] 

 Note: Y[dose] is the median response for exposure = dose; 

 sign = +1 for increasing trend in data; 

 sign = -1 for decreasing trend. 

 Model 2 is nested within Models 3 and 4. 

 Model 3 is nested within Model 5. 

 Model 4 is nested within Model 5. 

 Dependent variable = Mean 

 Independent variable = Dose 

 Data are assumed to be distributed: normally 

 Variance Model: exp(lnalpha +rho *ln(Y[dose])) 

 The variance is to be modeled as Var(i) = exp(lalpha + log(mean(i)) * rho) 

 Total number of dose groups = 5 

 Total number of records with missing values = 0 

 Maximum number of iterations = 250 

 Relative Function Convergence has been set to: 1e-008 

 Parameter Convergence has been set to: 1e-008 

 MLE solution provided: Exact 

 Initial Parameter Values 

Variable Model 2 Model 3 Model 4 Model 5 

-------- ------- ------- ------- ------- 

Lnalpha -11.0843 -11.0843 -11.0843 -11.0843 

rho 4.39551 4.39551 4.39551 4.39551 

 a 19.3889 19.3889 17.1285 17.1285 

 b 0.0525111 0.0525111 0.329637 0.329637 

 c -- -- 1.64165 1.64165 

 d -- 1 -- 1 

 Parameter Estimates by Model 

Variable Model 2 Model 3 Model 4 Model 5 

-------- ------- ------- ------- ------- 

Lnalpha -10.3398 -10.3398 -11.3983 -11.4672 

rho 4.17399 4.17399 4.50168 4.523 

 a 19.3333 19.3333 18.8392 18.9098 

 b 0.055133 0.055133 0.281713 0.393804 

 c -- -- 1.49669 1.42426 

 d -- 1 -- 1.2698 
Perfluorooctanoic Acid – February 2014 A-3 

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 Table of Stats From Input Data 

 Dose N Obs Mean Obs Std Dev 

 ----- --- ---------- ------------- 

 0 15 19.73 2.01 

 0.06 15 18.03 2.81 

 0.64 15 20.44 2.87 

 1.94 15 22.74 4.21 

 6.5 15 26.78 5.47 

 Estimated Values of Interest 

 Model Dose Est Mean Est Std Scaled Residual 

 ------- ------ ---------- --------- ---------------- 

 2 0 19.33 2.75 0.5588 

 0.06 19.4 2.769 -1.913 

 0.64 20.03 2.96 0.5396 

 1.94 21.52 3.437 1.379 

 6.5 27.67 5.809 -0.5906 

 3 0 19.33 2.75 0.5588 

 0.06 19.4 2.769 -1.913 

 0.64 20.03 2.96 0.5396 

 1.94 21.52 3.437 1.379 

 6.5 27.67 5.809 -0.5906 

 4 0 18.84 2.482 1.39 

 0.06 19 2.529 -1.479 

 0.64 20.38 2.964 0.07453 

 1.94 22.78 3.806 -0.03971 

 6.5 26.7 5.441 0.05896 

 5 0 18.91 2.496 1.273 

 0.06 18.98 2.516 -1.46 

 0.64 20.19 2.894 0.3352 

 1.94 22.99 3.882 -0.2494 

 6.5 26.64 5.416 0.1028 

 Other models for which likelihoods are calculated: 

 Model A1: Yij = Mu(i) + e(ij) 

 Var{e(ij)} = Sigma^2 

 Model A2: Yij = Mu(i) + e(ij) 

 Var{e(ij)} = Sigma(i)^2 

 Model A3: Yij = Mu(i) + e(ij) 

 Var{e(ij)} = exp(lalpha + log(mean(i)) * rho) 

 Model R: Yij = Mu + e(i) 

 Var{e(ij)} = Sigma^2 

 Likelihoods of Interest 

 Model Log(likelihood) DF AIC 

 ------- ----------------- ---- ------------ 

 A1 -132.6899 6 277.3799 

 A2 -123.7484 10 267.4967 

 A3 -125.6728 7 265.3456 

 R -153.0718 2 310.1437 

 2 -129.1329 4 266.2658 

 3 -129.1329 4 266.2658 
Perfluorooctanoic Acid – February 2014 A-4 

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 4 -126.8109 5 263.6218 

 5 -126.6814 6 265.3627 

 Additive constant for all log-likelihoods = -68.92. This constant added to 

the above values gives the log-likelihood including the term that does not 

depend on the model parameters. 

 Explanation of Tests 

 Test 1: Does response and/or variances differ among Dose levels? (A2 vs. R) 

 Test 2: Are Variances Homogeneous? (A2 vs. A1) 

 Test 3: Are variances adequately modeled? (A2 vs. A3) 

 Test 4: Does Model 2 fit the data? (A3 vs. 2) 

 Test 5a: Does Model 3 fit the data? (A3 vs 3) 

 Test 5b: Is Model 3 better than Model 2? (3 vs. 2) 

 Test 6a: Does Model 4 fit the data? (A3 vs 4) 

 Test 6b: Is Model 4 better than Model 2? (4 vs. 2) 

 Test 7a: Does Model 5 fit the data? (A3 vs 5) 

 Test 7b: Is Model 5 better than Model 3? (5 vs. 3) 

 Test 7c: Is Model 5 better than Model 4? (5 vs. 4) 

 Tests of Interest 

 Test -2*log(Likelihood Ratio) D. F. p-value 

 -------- ------------------------ ------ -------------- 

 Test 1 58.65 8 < 0.0001 

 Test 2 17.88 4 0.001301 

 Test 3 3.849 3 0.2783 

 Test 4 6.92 3 0.07449 

 Test 5a 6.92 3 0.07449 

 Test 5b 0 0 N/A 

 Test 6a 2.276 2 0.3204 

 Test 6b 4.644 1 0.03116 

 Test 7a 2.017 1 0.1555 

 Test 7b 4.903 2 0.08616 

 Test 7c 0.2591 1 0.6108 

 The p-value for Test 1 is less than .05. There appears to be a 

difference between response and/or variances among the dose levels, it seems 

appropriate to model the data. 

 The p-value for Test 2 is less than .1. A non-homogeneous variance model 

appears to be appropriate. 

 The p-value for Test 3 is greater than .1. The modeled variance appears 

to be appropriate here. 

 The p-value for Test 4 is less than .1. Model 2 may not adequately 

describe the data; you may want to consider another model. 

 The p-value for Test 5a is less than .1. Model 3 may not adequately 

describe the data; you may want to consider another model. 

 Degrees of freedom for Test 5b are less than or equal to 0. The Chi-Square 

test for fit is not valid. 
Perfluorooctanoic Acid – February 2014 A-5 

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 The p-value for Test 6a is greater than .1. Model 4 seems to adequately 

describe the data. 

 The p-value for Test 6b is less than .05. Model 4 appears to fit the data 

better than Model 2. 

 The p-value for Test 7a is greater than .1. Model 5 seems to adequately 

describe the data. 

 The p-value for Test 7b is greater than .05. Model 5 does not seem to fit 

the data better than Model 3. 

 The p-value for Test 7c is greater than .05. Model 5 does not seem to fit 

the data better than Model 4. 

 Benchmark Dose Computations: 

 Specified Effect = 0.100000 

 Risk Type = Relative deviation 

 Confidence Level = 0.950000 

 BMD and BMDL by Model 

 Model BMD BMDL 

 ------- ------------ ---------- 

 2 1.72873 1.34495 

3 1.72873 1.34495 

 4 0.798012 0.456224 

 5 0.902365 0.468724 

 16

 18

 20

 22

 24

 26

 28

 30

 0 1 2 3 4 5 6

Mean Response

dose

Exponential Model 4 with 0.95 Confidence Level

07:37 08/18 2011

BMDL BMD

 

Exponential
Perfluorooctanoic Acid – February 2014 A-6 

Draft – Do Not Cite or Quote 

2) Loveless et al., 2008: male rat 

==================================================================== 

 Exponential Model. (Version: 1.7; Date: 12/10/2009) 

 

 Thu Dec 22 12:43:43 2011 

 ==================================================================== 

 BMDS Model Run 

~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~ 

 

 The form of the response function by Model: 

 Model 2: Y[dose] = a * exp{sign * b * dose} 

 Model 3: Y[dose] = a * exp{sign * (b * dose)^d} 

 Model 4: Y[dose] = a * [c-(c-1) * exp{-b * dose}] 

 Model 5: Y[dose] = a * [c-(c-1) * exp{-(b * dose)^d}] 

 Note: Y[dose] is the median response for exposure = dose; 

 sign = +1 for increasing trend in data; 

 sign = -1 for decreasing trend. 

 Model 2 is nested within Models 3 and 4. 

 Model 3 is nested within Model 5. 

 Model 4 is nested within Model 5. 

 Dependent variable = Mean 

 Independent variable = Dose 

 Data are assumed to be distributed: normally 

 Variance Model: exp(lnalpha +rho *ln(Y[dose])) 

 The variance is to be modeled as Var(i) = exp(lalpha + log(mean(i)) * rho) 

 Total number of dose groups = 5 

 Total number of records with missing values = 0 

 Maximum number of iterations = 250 

 Relative Function Convergence has been set to: 1e-008 

 Parameter Convergence has been set to: 1e-008 

 MLE solution provided: Exact 

 Initial Parameter Values 

Variable Model 2 Model 3 Model 4 Model 5 

-------- ------- ------- ------- ------- 

 Lnalpha -7.96162 -7.96162 -7.96162 -7.96162 

 rho 3.3934 3.3934 3.3934 3.3934 

 a 15.5663 15.5663 12.54 12.54 

 b 0.00879874 0.00879874 0.0515133 0.0515133 

 c -- -- 1.80024 1.80024 

 d -- 1 -- 1 

 Parameter Estimates by Model 

Variable Model 2 Model 3 Model 4 Model 5 

-------- ------- ------- ------- ------- 

 Lnalpha -4.07115 -4.07115 -10.0648 -10.3628 

 rho 2.30379 2.30379 4.14318 4.24407 

 a 15.7112 15.7112 13.0813 13.1947 

 b 0.00896759 0.00896759 0.861545 0.977136 

 c -- -- 1.5326 1.51565 

d -- 1 -- 1.36232 
Perfluorooctanoic Acid – February 2014 A-7 

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 Table of Stats From Input Data 

 Dose N Obs Mean Obs Std Dev 

 ----- --- ---------- ------------- 

 0 10 13.2 1.4 

 0.3 10 14.4 1.6 

 1 10 17.2 2.9 

 10 10 21.5 2.9 

 30 10 18.7 2.9 

 Estimated Values of Interest 

 Model Dose Est Mean Est Std Scaled Residual 

 ------- ------ ---------- --------- ---------------- 

 2 0 15.71 3.118 -2.547 

 0.3 15.75 3.128 -1.368 

 1 15.85 3.15 1.352 

 10 17.19 3.457 3.947 

 30 20.56 4.251 -1.385 

 3 0 15.71 3.118 -2.547 

 0.3 15.75 3.128 -1.368 

 1 15.85 3.15 1.352 

 10 17.19 3.457 3.947 

 30 20.56 4.251 -1.385 

 4 0 13.08 1.342 0.2798 

 0.3 14.67 1.701 -0.4985 

 1 17.1 2.339 0.1288 

 10 20.05 3.249 1.414 

 30 20.05 3.25 -1.312 

 5 0 13.19 1.34 0.01248 

 0.3 14.36 1.604 0.07811 

 1 17.42 2.416 -0.2836 

 10 20 3.24 1.466 

 30 20 3.24 -1.268 

 Other models for which likelihoods are calculated: 

 Model A1: Yij = Mu(i) + e(ij) 

 Var{e(ij)} = Sigma^2 

 Model A2: Yij = Mu(i) + e(ij) 

 Var{e(ij)} = Sigma(i)^2 

 Model A3: Yij = Mu(i) + e(ij) 

 Var{e(ij)} = exp(lalpha + log(mean(i)) * rho) 

 Model R: Yij = Mu + e(i) 

 Var{e(ij)} = Sigma^2 

 Likelihoods of Interest 

 Model Log(likelihood) DF AIC 

 ------- ----------------- ---- ------------ 

 A1 -66.95077 6 145.9015 

 A2 -62.37207 10 144.7441 

 A3 -63.18657 7 140.3731 

 R -91.38557 2 186.7711 

 2 -86.12482 4 180.2496 

 3 -86.12482 4 180.2496 

 4 -65.31811 5 140.6362 
Perfluorooctanoic Acid – February 2014 A-8 

Draft – Do Not Cite or Quote 

 5 -64.98792 6 141.9758 

 Additive constant for all log-likelihoods = -45.95. This constant added to 

the above values gives the log-likelihood including the term that does not 

depend on the model parameters. 

 Explanation of Tests 

 Test 1: Does response and/or variances differ among Dose levels? (A2 vs. R) 

 Test 2: Are Variances Homogeneous? (A2 vs. A1) 

 Test 3: Are variances adequately modeled? (A2 vs. A3) 

 Test 4: Does Model 2 fit the data? (A3 vs. 2) 

 Test 5a: Does Model 3 fit the data? (A3 vs 3) 

 Test 5b: Is Model 3 better than Model 2? (3 vs. 2) 

 Test 6a: Does Model 4 fit the data? (A3 vs 4) 

 Test 6b: Is Model 4 better than Model 2? (4 vs. 2) 

 Test 7a: Does Model 5 fit the data? (A3 vs 5) 

 Test 7b: Is Model 5 better than Model 3? (5 vs. 3) 

 Test 7c: Is Model 5 better than Model 4? (5 vs. 4) 

 Tests of Interest 

 Test -2*log(Likelihood Ratio) D. F. p-value 

 -------- ------------------------ ------ -------------- 

 Test 1 58.03 8 < 0.0001 

 Test 2 9.157 4 0.05728 

 Test 3 1.629 3 0.6528 

 Test 4 45.88 3 < 0.0001 

 Test 5a 45.88 3 < 0.0001 

 Test 5b -1.563e-012 0 N/A 

 Test 6a 4.263 2 0.1187 

 Test 6b 41.61 1 < 0.0001 

 Test 7a 3.603 1 0.05769 

 Test 7b 42.27 2 < 0.0001 

 Test 7c 0.6604 1 0.4164 

 The p-value for Test 1 is less than .05. There appears to be a difference 

between response and/or variances among the dose levels, it seems appropriate 

to model the data. 

 The p-value for Test 2 is less than .1. A non-homogeneous variance model 

appears to be appropriate. 

 The p-value for Test 3 is greater than .1. The modeled variance appears 

to be appropriate here. 

 The p-value for Test 4 is less than .1. Model 2 may not adequately 

describe the data; you may want to consider another model. 

 The p-value for Test 5a is less than .1. Model 3 may not adequately 

describe the data; you may want to consider another model. 

 Degrees of freedom for Test 5b are less than or equal to 0. The Chi-Square 

test for fit is not valid. 

 The p-value for Test 6a is greater than .1. Model 4 seems to adequately 

describe the data. 
Perfluorooctanoic Acid – February 2014 A-9 

Draft – Do Not Cite or Quote 

 The p-value for Test 6b is less than .05. Model 4 appears to fit the data 

better than Model 2. 

 The p-value for Test 7a is less than .1. Model 5 may not adequately 

describe the data; you may want to consider another model. 

 The p-value for Test 7b is less than .05. Model 5 appears to fit the data 

better than Model 3. 

 The p-value for Test 7c is greater than .05. Model 5 does not seem to fit 

the data better than Model 4. 

 Benchmark Dose Computations: 

 Specified Effect = 0.100000 

 Risk Type = Relative deviation 

 Confidence Level = 0.950000 

 BMD and BMDL by Model 

 Model BMD BMDL 

 ------- ------------ ---------- 

 2 10.6283 6.02231 

 3 10.6283 6.02231 

 4 0.241376 0.151833 

 5 0.331813 0.161379 
Perfluorooctanoic Acid – February 2014 A-10 

Draft – Do Not Cite or Quote 

 12

 14

 16

 18

 20

 22

 24

 0 5 10 15 20 25 30

Mean Response

dose

Exponential Model 4 with 0.95 Confidence Level

13:12 12/22 2011

BMDLBMD

 

Exponential
Perfluorooctanoic Acid – February 2014 A-11 

Draft – Do Not Cite or Quote 

3) York, 2002; Butenhoff et al., 2004a; York et al., 2010 

==================================================================== 

 Exponential Model. (Version: 1.7; Date: 12/10/2009) 

 

 Wed Oct 26 10:57:26 2011 

 ==================================================================== 

 BMDS Model Run 

~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~ 

 

 The form of the response function by Model: 

 Model 2: Y[dose] = a * exp{sign * b * dose} 

 Model 3: Y[dose] = a * exp{sign * (b * dose)^d} 

 Model 4: Y[dose] = a * [c-(c-1) * exp{-b * dose}] 

 Model 5: Y[dose] = a * [c-(c-1) * exp{-(b * dose)^d}] 

 Note: Y[dose] is the median response for exposure = dose; 

 sign = +1 for increasing trend in data; 

 sign = -1 for decreasing trend. 

 Model 2 is nested within Models 3 and 4. 

 Model 3 is nested within Model 5. 

 Model 4 is nested within Model 5. 

 Dependent variable = Mean 

 Independent variable = Dose 

 Data are assumed to be distributed: normally 

 Variance Model: exp(lnalpha +rho *ln(Y[dose])) 

 The variance is to be modeled as Var(i) = exp(lalpha + log(mean(i)) * rho) 

 Total number of dose groups = 5 

 Total number of records with missing values = 0 

 Maximum number of iterations = 250 

 Relative Function Convergence has been set to: 1e-008 

 Parameter Convergence has been set to: 1e-008 

 MLE solution provided: Exact 

 Initial Parameter Values 

Variable Model 2 Model 3 Model 4 Model 5 

-------- ------- ------- ------- ------- 

 lnalpha -3.83873 -3.83873 -3.83873 -3.83873 

 rho 1.89443 1.89443 1.89443 1.89443 

 a 24.2168 24.2168 19.285 19.285 

 b 0.00591038 0.00591038 0.0671202 0.0671202 

 c -- -- 1.56261 1.56261 

 d -- 1 -- 1 

 Parameter Estimates by Model 

Variable Model 2 Model 3 Model 4 Model 5 

-------- ------- ------- ------- ------- 

 lnalpha 5.96811 5.96811 -4.75762 -4.61574 

 rho -0.961701 -0.961702 2.17955 2.13551 

 a 24.5758 24.5758 20.2791 20.2956 

 b 0.0049661 0.0049661 0.737783 0.745578 

 c -- -- 1.39208 1.38868 

 d -- 1 -- 1.12767 
Perfluorooctanoic Acid – February 2014 A-12 

Draft – Do Not Cite or Quote 

 Table of Stats From Input Data 

 Dose N Obs Mean Obs Std Dev 

 ----- --- ---------- ------------- 

 0 30 20.3 2.5 

 1 30 24.3 3.2 

 3 30 27.7 2.7 

 10 30 28.7 3.9 

 30 29 27.5 3.7 

 Estimated Values of Interest 

 Model Dose Est Mean Est Std Scaled Residual 

 ------- ------ ---------- --------- ---------------- 

 2 0 24.58 4.24 -5.524 

 1 24.7 4.23 -0.5157 

 3 24.94 4.209 3.585 

 10 25.83 4.14 3.801 

 30 28.52 3.947 -1.397 

 3 0 24.58 4.24 -5.524 

 1 24.7 4.23 -0.5157 

 3 24.94 4.209 3.585 

 10 25.83 4.14 3.801 

 30 28.52 3.947 -1.397 

 4 0 20.28 2.462 0.04647 

 1 24.43 3.016 -0.2328 

 3 27.36 3.412 0.5445 

 10 28.23 3.53 0.7368 

 30 28.23 3.531 -1.114 

 5 0 20.3 2.476 0.009804 

 1 24.34 3.005 -0.06777 

 3 27.52 3.427 0.2834 

 10 28.18 3.515 0.8048 

 30 28.18 3.515 -1.048 

 Other models for which likelihoods are calculated: 

 Model A1: Yij = Mu(i) + e(ij) 

 Var{e(ij)} = Sigma^2 

 Model A2: Yij = Mu(i) + e(ij) 

 Var{e(ij)} = Sigma(i)^2 

 Model A3: Yij = Mu(i) + e(ij) 

 Var{e(ij)} = exp(lalpha + log(mean(i)) * rho) 

 Model R: Yij = Mu + e(i) 

 Var{e(ij)} = Sigma^2 

 Likelihoods of Interest 

 Model Log(likelihood) DF AIC 

 ------- ----------------- ---- ------------ 

 A1 -247.2345 6 506.469 

 A2 -242.9088 10 505.8177 

 A3 -244.9146 7 503.8291 

 R -296.4613 2 596.9226 

 2 -286.6477 4 581.2954 

 3 -286.6477 4 581.2954 

 4 -245.8857 5 501.7713 

 5 -245.8276 6 503.6552 
Perfluorooctanoic Acid – February 2014 A-13 

Draft – Do Not Cite or Quote 

 Additive constant for all log-likelihoods = -136.9. This constant added to 

the above values gives the log-likelihood including the term that does not 

depend on the model parameters. 

 Explanation of Tests 

 Test 1: Does response and/or variances differ among Dose levels? (A2 vs. R) 

 Test 2: Are Variances Homogeneous? (A2 vs. A1) 

 Test 3: Are variances adequately modeled? (A2 vs. A3) 

 Test 4: Does Model 2 fit the data? (A3 vs. 2) 

 Test 5a: Does Model 3 fit the data? (A3 vs 3) 

 Test 5b: Is Model 3 better than Model 2? (3 vs. 2) 

 Test 6a: Does Model 4 fit the data? (A3 vs 4) 

 Test 6b: Is Model 4 better than Model 2? (4 vs. 2) 

 Test 7a: Does Model 5 fit the data? (A3 vs 5) 

 Test 7b: Is Model 5 better than Model 3? (5 vs. 3) 

 Test 7c: Is Model 5 better than Model 4? (5 vs. 4) 

 Tests of Interest 

 Test -2*log(Likelihood Ratio) D. F. p-value 

 -------- ------------------------ ------ -------------- 

 Test 1 107.1 8 < 0.0001 

 Test 2 8.651 4 0.07043 

 Test 3 4.011 3 0.2602 

 Test 4 83.47 3 < 0.0001 

 Test 5a 83.47 3 < 0.0001 

 Test 5b -2.274e-013 0 N/A 

 Test 6a 1.942 2 0.3787 

 Test 6b 81.52 1 < 0.0001 

 Test 7a 1.826 1 0.1766 

 Test 7b 81.64 2 < 0.0001 

 Test 7c 0.1162 1 0.7332 

 The p-value for Test 1 is less than .05. There appears to be a difference 

between response and/or variances among the dose levels, it seems appropriate 

to model the data. 

 The p-value for Test 2 is less than .1. A non-homogeneous variance model 

appears to be appropriate. 

 The p-value for Test 3 is greater than .1. The modeled variance appears 

to be appropriate here. 

 The p-value for Test 4 is less than .1. Model 2 may not adequately 

describe the data; you may want to consider another model. 

 The p-value for Test 5a is less than .1. Model 3 may not adequately 

describe the data; you may want to consider another model. 

 Degrees of freedom for Test 5b are less than or equal to 0. The ChiSquare test for fit is not valid. 

 The p-value for Test 6a is greater than .1. Model 4 seems to adequately 

describe the data. 
Perfluorooctanoic Acid – February 2014 A-14 

Draft – Do Not Cite or Quote 

 The p-value for Test 6b is less than .05. Model 4 appears to fit the data 

better than Model 2. 

 The p-value for Test 7a is greater than .1. Model 5 seems to adequately 

describe the data. 

 The p-value for Test 7b is less than .05. Model 5 appears to fit the data 

better than Model 3. 

 The p-value for Test 7c is greater than .05. Model 5 does not seem to fit 

the data better than Model 4. 

 Benchmark Dose Computations: 

 Specified Effect = 0.100000 

 Risk Type = Relative deviation 

 Confidence Level = 0.950000 

 BMD and BMDL by Model 

 Model BMD BMDL 

 ------- ------------ ---------- 

 2 19.1922 13.6199 

 3 19.1922 13.6199 

 4 0.399087 0.274007 

 5 0.45764 0.276304 

 20

 22

 24

 26

 28

 30

 0 5 10 15 20 25 30

Mean Response

dose

Exponential Model 4 with 0.95 Confidence Level

10:57 10/26 2011

BMDLBMD

 

Exponential
Perfluorooctanoic Acid – February 2014 A-15 

Draft – Do Not Cite or Quote 

4) Loveless et al., 2008: male mouse 

==================================================================== 

 Exponential Model. (Version: 1.7; Date: 12/10/2009) 

 

 Wed Aug 24 10:48:22 2011 

 ==================================================================== 

 BMDS Model Run 

~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~ 

 

 The form of the response function by Model: 

 Model 2: Y[dose] = a * exp{sign * b * dose} 

 Model 3: Y[dose] = a * exp{sign * (b * dose)^d} 

 Model 4: Y[dose] = a * [c-(c-1) * exp{-b * dose}] 

 Model 5: Y[dose] = a * [c-(c-1) * exp{-(b * dose)^d}] 

 Note: Y[dose] is the median response for exposure = dose; 

 sign = +1 for increasing trend in data; 

 sign = -1 for decreasing trend. 

 Model 2 is nested within Models 3 and 4. 

 Model 3 is nested within Model 5. 

 Model 4 is nested within Model 5. 

 Dependent variable = Mean 

 Independent variable = Dose 

 Data are assumed to be distributed: normally 

 Variance Model: exp(lnalpha +rho *ln(Y[dose])) 

 rho is set to 0. 

 A constant variance model is fit. 

 Total number of dose groups = 5 

 Total number of records with missing values = 0 

 Maximum number of iterations = 250 

 Relative Function Convergence has been set to: 1e-008 

 Parameter Convergence has been set to: 1e-008 

 MLE solution provided: Exact 

 Initial Parameter Values 

Variable Model 2 Model 3 Model 4 Model 5 

-------- ------- ------- ------- ------- 

 lnalpha -0.698397 -0.698397 -0.698397 -0.698397 

 rho(S) 0 0 0 0 

 a 2.65446 2.65446 1.691 1.691 

 b 0.0323602 0.0323602 0.0970613 0.0970613 

 c -- -- 3.76286 3.76286 

 d -- 1 -- 1 

(S) = Specified 
Perfluorooctanoic Acid – February 2014 A-16 

Draft – Do Not Cite or Quote 

 Parameter Estimates by Model 

Variable Model 2 Model 3 Model 4 Model 5 

-------- ------- ------- ------- ------- 

 lnalpha 0.636609 0.636609 -0.686179 -0.686255 

 rho 0 0 0 0 

 a 3.06456 3.06456 1.8131 1.8194 

 b 0.0242388 0.0242388 0.446049 0.450771 

 c -- -- 3.30881 3.29563 

 d -- 1 -- 1.01732 

 Table of Stats From Input Data 

 Dose N Obs Mean Obs Std Dev 

 ----- --- ---------- ------------- 

 0 20 1.78 0.18 

 1 20 3.27 0.23 

 0.3 20 2.41 0.26 

 30 20 5.9 0.85 

 10 20 6.06 1.32 

 Estimated Values of Interest 

 Model Dose Est Mean Est Std Scaled Residual 

 ------- ------ ---------- --------- ---------------- 

 2 0 3.065 1.375 -4.179 

 1 3.14 1.375 0.4237 

 0.3 3.087 1.375 -2.202 

 30 6.341 1.375 -1.435 

 10 3.905 1.375 7.01 

 3 0 3.065 1.375 -4.179 

 1 3.14 1.375 0.4237 

 0.3 3.087 1.375 -2.202 

 30 6.341 1.375 -1.435 

 10 3.905 1.375 7.01 

 4 0 1.813 0.7096 -0.2086 

 1 3.319 0.7096 -0.3117 

 0.3 2.337 0.7096 0.4575 

 30 5.999 0.7096 -0.6252 

 10 5.951 0.7096 0.688 

 5 0 1.819 0.7095 -0.2483 

 1 3.318 0.7095 -0.3055 

 0.3 2.331 0.7095 0.4988 

 30 5.996 0.7095 -0.6055 

 10 5.955 0.7095 0.6605 

 Other models for which likelihoods are calculated: 

 Model A1: Yij = Mu(i) + e(ij) 

 Var{e(ij)} = Sigma^2 

 Model A2: Yij = Mu(i) + e(ij) 

 Var{e(ij)} = Sigma(i)^2 

 Model A3: Yij = Mu(i) + e(ij) 

 Var{e(ij)} = exp(lalpha + log(mean(i)) * rho) 

 Model R: Yij = Mu + e(i) 

 Var{e(ij)} = Sigma^2 

 Likelihoods of Interest 
Perfluorooctanoic Acid – February 2014 A-17 

Draft – Do Not Cite or Quote 

 Model Log(likelihood) DF AIC 

 ------- ----------------- ---- ------------ 

 A1 -15.08015 6 42.16031 

 A2 40.89337 10 -61.78674 

 A3 -15.08015 6 42.16031 

 R -114.7709 2 233.5419 

 2 -81.83045 3 169.6609 

 3 -81.83045 3 169.6609 

 4 -15.69106 4 39.38211 

 5 -15.68723 5 41.37446 

 Additive constant for all log-likelihoods = -91.89. This constant added 

to the above values gives the log-likelihood including the term that does not 

depend on the model parameters. 

 Explanation of Tests 

 Test 1: Does response and/or variances differ among Dose levels? (A2 vs. R) 

 Test 2: Are Variances Homogeneous? (A2 vs. A1) 

 Test 3: Are variances adequately modeled? (A2 vs. A3) 

 Test 4: Does Model 2 fit the data? (A3 vs. 2) 

 Test 5a: Does Model 3 fit the data? (A3 vs 3) 

 Test 5b: Is Model 3 better than Model 2? (3 vs. 2) 

 Test 6a: Does Model 4 fit the data? (A3 vs 4) 

 Test 6b: Is Model 4 better than Model 2? (4 vs. 2) 

 Test 7a: Does Model 5 fit the data? (A3 vs 5) 

 Test 7b: Is Model 5 better than Model 3? (5 vs. 3) 

 Test 7c: Is Model 5 better than Model 4? (5 vs. 4) 

 Tests of Interest 

 Test -2*log(Likelihood Ratio) D. F. p-value 

 -------- ------------------------ ------ -------------- 

 Test 1 311.3 8 < 0.0001 

 Test 2 111.9 4 < 0.0001 

 Test 3 111.9 4 < 0.0001 

 Test 4 133.5 3 < 0.0001 

 Test 5a 133.5 3 < 0.0001 

 Test 5b -7.788e-012 0 N/A 

 Test 6a 1.222 2 0.5429 

 Test 6b 132.3 1 < 0.0001 

 Test 7a 1.214 1 0.2705 

 Test 7b 132.3 2 < 0.0001 

 Test 7c 0.007649 1 0.9303 

 The p-value for Test 1 is less than .05. There appears to be a difference 

between response and/or variances among the dose levels, it seems appropriate 

to model the data. 

 The p-value for Test 2 is less than .1. Consider running a nonhomogeneous variance model. 

 The p-value for Test 3 is less than .1. You may want to consider a 

different variance model. 

 The p-value for Test 4 is less than .1. Model 2 may not adequately 

describe the data; you may want to consider another model. 
Perfluorooctanoic Acid – February 2014 A-18 

Draft – Do Not Cite or Quote 

 The p-value for Test 5a is less than .1. Model 3 may not adequately 

describe the data; you may want to consider another model. 

 Degrees of freedom for Test 5b are less than or equal to 0. The Chi-Square 

test for fit is not valid. 

 The p-value for Test 6a is greater than .1. Model 4 seems to adequately 

describe the data. 

 The p-value for Test 6b is less than .05. Model 4 appears to fit the data 

better than Model 2. 

 The p-value for Test 7a is greater than .1. Model 5 seems to adequately 

describe the data. 

 The p-value for Test 7b is less than .05. Model 5 appears to fit the data 

better than Model 3. 

 The p-value for Test 7c is greater than .05. Model 5 does not seem to fit 

the data better than Model 4. 

 Benchmark Dose Computations: 

 Specified Effect = 0.100000 

 Risk Type = Relative deviation 

 Confidence Level = 0.950000 

 BMD and BMDL by Model 

 Model BMD BMDL 

 ------- ------------ ---------- 

 2 3.93213 3.38551 

 3 3.93213 3.38551 

 4 0.099268 0.0680743 

 5 0.104178 0.0681103 
Perfluorooctanoic Acid – February 2014 A-19 

Draft – Do Not Cite or Quote 

 2

 3

 4

 5

 6

 7

 0 5 10 15 20 25 30

Mean Response

dose

Exponential Model 4 with 0.95 Confidence Level

10:48 08/24 2011

BMDLBMD

 

Exponential
Perfluorooctanoic Acid – February 2014 A-20 

Draft – Do Not Cite or Quote 

5) Lau et al., 2006: female mouse

==================================================================== 

 Exponential Model. (Version: 1.7; Date: 12/10/2009) 

 Tue Oct 25 15:12:47 2011 

 ==================================================================== 

 BMDS Model Run 

~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~~ 

 

 The form of the response function by Model: 

 Model 2: Y[dose] = a * exp{sign * b * dose} 

 Model 3: Y[dose] = a * exp{sign * (b * dose)^d} 

 Model 4: Y[dose] = a * [c-(c-1) * exp{-b * dose}] 

 Model 5: Y[dose] = a * [c-(c-1) * exp{-(b * dose)^d}] 

 Note: Y[dose] is the median response for exposure = dose; 

 sign = +1 for increasing trend in data; 

 sign = -1 for decreasing trend. 

 Model 2 is nested within Models 3 and 4. 

 Model 3 is nested within Model 5. 

 Model 4 is nested within Model 5. 

 Dependent variable = Mean 

 Independent variable = Dose 

 Data are assumed to be distributed: normally 

 Variance Model: exp(lnalpha +rho *ln(Y[dose])) 

 The variance is to be modeled as Var(i) = exp(lalpha + log(mean(i)) * rho) 

 Total number of dose groups = 6 

 Total number of records with missing values = 0 

 Maximum number of iterations = 250 

 Relative Function Convergence has been set to: 1e-008 

 Parameter Convergence has been set to: 1e-008 

 MLE solution provided: Exact 

 Initial Parameter Values 

Variable Model 2 Model 3 Model 4 Model 5 

-------- ------- ------- ------- ------- 

 lnalpha -5.21753 -5.21753 -5.21753 -5.21753 

 rho 2.68479 2.68479 2.68479 2.68479 

 a 2.91529 2.91529 2.3085 2.3085 

 b 0.0531403 0.0531403 0.149585 0.149585 

 c -- -- 2.46979 2.46979 

 d -- 1 -- 1 

 Parameter Estimates by Model 

Variable Model 2 Model 3 Model 4 Model 5 

-------- ------- ------- ------- ------- 

 lnalpha -5.42859 -5.42858 -4.62127 -4.62765 

 rho 3.52242 3.52242 2.25579 2.26038 

 a 2.63415 2.63415 2.4311 2.43306 

 b 0.0990961 0.0990961 0.279706 0.286892 

 c -- -- 2.27404 2.25849 

d -- 1 -- 1.03267 

 Table of Stats From Input Data 

 Dose N Obs Mean Obs Std Dev 

 ----- --- ---------- ------------- 
Perfluorooctanoic Acid – February 2014 A-21 

Draft – Do Not Cite or Quote 

 0 40 2.43 0.29 

 1 15 3.17 0.27 

 3 16 4.3 0.51 

 5 20 4.66 0.56 

 10 15 5.43 0.74 

 20 5 5.36 0.77 

 Estimated Values of Interest 

 Model Dose Est Mean Est Std Scaled Residual 

 ------- ------ ---------- --------- ---------------- 

 2 0 2.634 0.3648 -3.54 

 1 2.909 0.4343 2.331 

 3 3.546 0.6158 4.897 

 5 4.323 0.873 1.724 

 10 7.096 2.089 -3.088 

 20 19.12 11.97 -2.57 

 3 0 2.634 0.3648 -3.54 

 1 2.909 0.4343 2.331 

 3 3.546 0.6158 4.897 

 5 4.323 0.873 1.724 

 10 7.096 2.089 -3.088 

 20 19.12 11.97 -2.57 

 4 0 2.431 0.2702 -0.02564 

 1 3.187 0.3666 -0.1777 

 3 4.19 0.4992 0.8807 

 5 4.763 0.5769 -0.8022 

 10 5.34 0.6562 0.5341 

 20 5.517 0.6809 -0.5152 

 5 0 2.433 0.2701 -0.07155 

 1 3.17 0.3643 -0.002448 

 3 4.195 0.4999 0.8422 

 5 4.778 0.5791 -0.91 

 10 5.338 0.6564 0.5436 

 20 5.488 0.6773 -0.4226 

 Other models for which likelihoods are calculated: 

 Model A1: Yij = Mu(i) + e(ij) 

 Var{e(ij)} = Sigma^2 

 Model A2: Yij = Mu(i) + e(ij) 

 Var{e(ij)} = Sigma(i)^2 

 Model A3: Yij = Mu(i) + e(ij) 

 Var{e(ij)} = exp(lalpha + log(mean(i)) * rho) 

 Model R: Yij = Mu + e(i) 

 Var{e(ij)} = Sigma^2 

 Likelihoods of Interest 

 Model Log(likelihood) DF AIC 

 ------- ----------------- ---- ------------ 

 A1 29.04002 7 -44.08005 

 A2 44.97661 12 -65.95321 

 A3 43.40484 8 -70.80967 

 R -80.9645 2 165.929 

 2 -15.64228 4 39.28456 

 3 -15.64228 4 39.28456 
Perfluorooctanoic Acid – February 2014 A-22 

Draft – Do Not Cite or Quote 

 4 42.25345 5 -74.50691 

 5 42.2837 6 -72.56741 

 Additive constant for all log-likelihoods = -102. This constant added 

to the above values gives the log-likelihood including the term that does not 

depend on the model parameters. 

 Explanation of Tests 

 Test 1: Does response and/or variances differ among Dose levels? (A2 vs. R) 

 Test 2: Are Variances Homogeneous? (A2 vs. A1) 

 Test 3: Are variances adequately modeled? (A2 vs. A3) 

 Test 4: Does Model 2 fit the data? (A3 vs. 2) 

 Test 5a: Does Model 3 fit the data? (A3 vs 3) 

 Test 5b: Is Model 3 better than Model 2? (3 vs. 2) 

 Test 6a: Does Model 4 fit the data? (A3 vs 4) 

 Test 6b: Is Model 4 better than Model 2? (4 vs. 2) 

 Test 7a: Does Model 5 fit the data? (A3 vs 5) 

 Test 7b: Is Model 5 better than Model 3? (5 vs. 3) 

 Test 7c: Is Model 5 better than Model 4? (5 vs. 4) 

 Tests of Interest 

 Test -2*log(Likelihood Ratio) D. F. p-value 

 -------- ------------------------ ------ -------------- 

 Test 1 251.9 10 < 0.0001 

 Test 2 31.87 5 < 0.0001 

 Test 3 3.144 4 0.5341 

 Test 4 118.1 4 < 0.0001 

 Test 5a 118.1 4 < 0.0001 

 Test 5b -9.564e-012 0 N/A 

 Test 6a 2.303 3 0.512 

 Test 6b 115.8 1 < 0.0001 

 Test 7a 2.242 2 0.3259 

 Test 7b 115.9 2 < 0.0001 

 Test 7c 0.0605 1 0.8057 

 The p-value for Test 1 is less than .05. There appears to be a difference 

between response and/or variances among the dose levels, it seems appropriate 

to model the data. 

 The p-value for Test 2 is less than .1. A non-homogeneous variance model 

appears to be appropriate. 

 The p-value for Test 3 is greater than .1. The modeled variance appears 

to be appropriate here. 

 The p-value for Test 4 is less than .1. Model 2 may not adequately 

describe the data; you may want to consider another model. 

 The p-value for Test 5a is less than .1. Model 3 may not adequately 

describe the data; you may want to consider another model. 

 Degrees of freedom for Test 5b are less than or equal to 0. The Chi-Square 

test for fit is not valid. 
Perfluorooctanoic Acid – February 2014 A-23 

Draft – Do Not Cite or Quote 

 The p-value for Test 6a is greater than .1. Model 4 seems to adequately 

describe the data. 

 The p-value for Test 6b is less than .05. Model 4 appears to fit the data 

better than Model 2. 

 The p-value for Test 7a is greater than .1. Model 5 seems to adequately 

describe the data. 

 The p-value for Test 7b is less than .05. Model 5 appears to fit the data 

better than Model 3. 

 The p-value for Test 7c is greater than .05. Model 5 does not seem to fit 

the data better than Model 4. 

 Benchmark Dose Computations: 

 Specified Effect = 0.100000 

 Risk Type = Relative deviation 

 Confidence Level = 0.950000 

 BMD and BMDL by Model 

 Model BMD BMDL 

 ------- ------------ ---------- 

 2 0.961795 0.832065 

 3 0.961796 0.832065 

 4 0.292244 0.246387 

 5 0.312258 0.246874 
Perfluorooctanoic Acid – February 2014 A-24 

Draft – Do Not Cite or Quote 

 2

 2.5

 3

 3.5

 4

 4.5

 5

 5.5

 6

 6.5

 0 5 10 15 20

Mean Response

dose

Exponential Model 4 with 0.95 Confidence Level

15:12 10/25 2011

BMDLBMD

 

Exponential
Perfluorooctanoic Acid – February 2014 B-1 

Draft – Do Not Cite or Quote 

APPENDIX B 

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 
Perfluorooctanoic Acid – February 2014 B-2 

Draft – Do Not Cite or Quote 

 95.0% Wald Confidence 

Interval 

 Variable Estimate 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. 

 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 

 Scaled 

 Dose Est._Prob. Expected Observed Size 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 – February 2014 B-3 

Draft – Do Not Cite or Quote 

 ==================================================================== 

 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 

 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 – February 2014 B-4 

Draft – Do Not Cite or Quote 

 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 

 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 

 95.0% Wald Confidence 

Interval 

 Variable Estimate 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 

 Scaled 

 Dose Est._Prob. Expected Observed Size 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 
Perfluorooctanoic Acid – February 2014 B-5 

Draft – Do Not Cite or Quote 

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