ve/EPA
United States Office of Water EPA 822-R-16-003
Environmental Protection Mail Code 4304T May 2016
Agency
Health Effects Support
Document for
Perfluorooctanoic Acid
(PFOA)
Perfluorooctanoic Acid - May 2016
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Health Effects Support Document
for
Perfluorooctanoic Acid (PFOA)
U.S. Environmental Protection Agency
Office of Water (43 04T)
Health and Ecological Criteria Division
Washington, DC 20460
EPA Document Number: 822-R-16-003
May 2016
Perfluorooctanoic Acid - May 2016
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BACKGROUND
The Safe Drinking Water Act (SDWA), as amended in 1996, requires the Administrator of
the U.S. Environmental Protection Agency (EPA) to periodically publish a list of unregulated
chemical contaminants known or anticipated to occur in public water systems and that may
require regulation under SDWA. The SDWA also requires the Agency to make regulatory
determinations on at least five contaminants on the Contaminant Candidate List (CCL) every
5 years. For each contaminant on the CCL, before EPA makes a regulatory determination, the
Agency needs to obtain sufficient data to conduct analyses on the extent to which the
contaminant occurs and the risk it poses to populations via drinking water. Ultimately, this
information will assist the Agency in determining the most appropriate course of action in
relation to the contaminant (e.g., developing a regulation to control it in drinking water,
developing guidance, or deciding not to regulate it).
The PFOA health assessment was initiated by the Office of Water, Office of Science and
Technology in 2009. The draft Health Effects Support Document for Perfluoroctanoic Acid
(PFOA) was completed in 2013 and released for public comment in February 2014. An external
peer-review panel meeting was held on August 21 and 22, 2014. The final document reflects
input from the panel as well as public comments received on the draft document. Both the peer-
reviewed draft and this document include only the sections of a health effects support document
(HESD) that cover the toxicokinetics and health effects of PFOA. If a decision is made to
regulate the contaminant, this document will be expanded.
One of the challenges inherent in conducting this assessment was the wealth of experimental
data published before and during its development. This section provides a synopsis of the
approach used in identifying and selecting the publications reflected in the final assessment.
Data were identified through the following:
Monthly/bimonthly literature searches conducted by EPA library staff (2009-2015) and
New Jersey Department of Environmental Protection library staff (2012-2015).
Papers identified by EPA internal and external peer reviewers.
Papers identified through public comments on the draft assessments.
Papers submitted to EPA by the public.
In mid-2013, the EPA library searches were expanded to cover all members of the
perfluoroalkane carboxylate family (C4 through C12). Appendix A describes the literature search
strategy used by the libraries. Through the literature search, documents were identified for
retrieval, review, and inclusion in the HESD using the following criteria:
The study examines a toxicity endpoint or population not examined by studies already
included in the draft document.
Aspects of the study design such as the size of the population exposed or quantification
approach make it superior to key studies already included in the draft document.
The data contribute substantially to the weight of evidence for any of the toxicity
endpoints covered by the draft document.
Elements of the study design merit its inclusion in the draft document based on its
contribution to the mode of action (MoA) or the quantification approach.
Perfluorooctanoic Acid - May 2016 iii
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The study elucidates the MoA for any toxicity endpoint or toxicokinetic property
associated with PFOA exposure.
The effects observed differ from those in other studies with comparable protocols.
In addition to each publication being evaluated against the criteria above, the relevance of the
study to drinking water exposures and to the U.S. population also were considered.
The studies included in the final draft were determined to provide the most current and
comprehensive description of the toxicological properties of PFOA and the risk it poses to
humans exposed to it in their drinking water. Appendix B summarizes the studies evaluated for
inclusion in the FIESD following the August 2014 peer review and identifies those selected for
inclusion in the final assessment. Appendix B includes epidemiology data that provide a high-
level summary of the outcomes across the studies evaluated.
Development of the hazard identification and dose-response assessment for PFOA has
followed the general guidelines for risk assessment forth by the National Research Council
(1983) and EPA's Framework for Human Health Risk Assessment to Inform Decision Making
(USEPA 2014a). Other EPA guidelines used in the development of this assessment include the
following:
Guidelines for the Health Risk Assessment of Chemical Mixtures (USEPA 1986a)
Guidelines for Mutagenicity Risk Assessment (USEPA 1986b)
Recommendations for and Documentation of Biological Values for Use in Risk
Assessment (USEPA 1988)
Guidelines for Developmental Toxicity Risk Assessment (USEPA 1991)
Interim Policy for Particle Size and Limit Concentration Issues in Inhalation Toxicity
Studies (USEPA 1994a)
Methods for Derivation of Inhalation Reference Concentrations and Application of
Inhalation Dosimetry (USEPA 1994b)
Use of the Benchmark Dose Approach in Health Risk Assessment (USEPA, 1995)
Guidelines for Reproductive Toxicity Risk Assessment (USEPA 1996)
Guidelines for Neurotoxicity Risk Assessment (USEPA 1998)
Science Policy Council Handbook: Peer Review (2nd edition) (USEPA 2000a)
Supplemental Guidance for Conducting Health Risk Assessment of Chemical Mixtures
(USEPA 2000b)
A Review of the Reference Dose and Reference Concentration Processes (USEPA 2002a)
Guidelines for Carcinogen Risk Assessment (USEPA 2005a)
Supplemental Guidance for Assessing Susceptibility from Early-Life Exposure to
Carcinogens (USEPA 2005b)
Science Policy Council Handbook: Peer Review (3rd edition) (USEPA 2006a)
A Framework for Assessing Health Risks of Environmental Exposures to Children
(USEPA 2006b)
Exposure Factors Handbook (USEPA 2011)
Benchmark Dose Technical Guidance Document (USEPA 2012)
Child-Specific Exposure Scenarios Examples (USEPA 2014b)
Perfluorooctanoic Acid - May 2016 iv
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AUTHORS
Joyce Morrissey Donohue, Ph.D. (Chemical Manager)
Office of Water
U.S. Environmental Protection Agency, Washington, DC
Tina Moore Duke, M.S. (previously with Office of Water, U.S. Environmental Protection
Agency)
John Wambaugh, Ph.D.
Office of Research and Development
U.S. Environmental Protection Agency, Research Triangle Park, NC
The following contractors supported the development of this document:
Jennifer Rayner, Ph.D., D.A.B.T.
Environmental Sciences Division
Oak Ridge National Laboratory, Oak Ridge, TN
Carol S. Wood, Ph.D., D.A.B.T.
Environmental Sciences Division
Oak Ridge National Laboratory, Oak Ridge, TN
This document was prepared under the U.S. EPA Contract No. DW-8992342701, Work
Assignment No. 2011-001 with Oak Ridge National Laboratory. The Lead U.S. EPA Scientist is
Joyce Morrissey Donohue, Ph.D., Health and Ecological Criteria Division, Office of Science and
Technology, Office of Water.
The Oak Ridge National Laboratory is managed and operated by UT-Battelle, LLC., for the U.S.
Department of Energy under Contract No. DE-AC05-OOOR22725.
Contributors and Reviewers
Internal Contributors and Reviewers
Office of Water, U.S. Environmental Protection Agency
Elizabeth Doyle, Ph.D. (retired)
Edward Hackett
Office of Research and Development, U.S. Environmental Protection Agency
Glinda Cooper, Ph.D.
Barbara Glenn, Ph.D.
Erin Hines, Ph.D.
Christopher Lau, Ph.D.
Matthew Lorber, Ph.D.
Jaqueline Moya
Linda Phillips, Ph.D.
Paul White, Ph.D.
Michael Wright, Sc.D.
Perfluorooctanoic Acid - May 2016
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Office of Chemical Safety and Pollution Prevention, U.S. Environmental Protection Agency
E. Laurence Lib el o
Andrea Pfehales-Hutchens, Ph.D.
Tracy Williamson
David Lai, Ph.D. (retired)
Jennifer Seed, Ph.D. (retired)
Office of Childrens Health Protection, U.S. Environmental Protection Agency
Gregory Miller
Office of Land and Emergency Management, U.S. Environmental Protection Agency
External Reviewers
James Bruckner, Ph.D.
Department of Pharmacology and Toxicology
University of Georgia, Athens, GA
Deborah Cory-Slechta, Ph.D.
Department of Environmental Medicine
University of Rochester Medical Center, Rochester, NY
Jamie DeWitt, Ph.D.
Pharmacology and Toxicology
East Carolina University, Greenville, NC
Jeffrey Fisher, Ph.D.
Biochemical Toxicology, NCTR
U.S. Food & Drug Administration, Jefferson, AK
William Hayton, Ph.D.
College of Pharmacy (Emeritus)
The Ohio State University, Columbus, OH
Matthew Longnecker, M.D., Sc.D.
Biomarker-based Epidemiology Group
National Institute of Environmental Health Sciences, Research Triangle Park, NC
Angela Slitt, Ph.D.
Biomedical and Pharmaceutical Sciences
University of Rhode Island, Kingston, RI
Perfluorooctanoic Acid - May 2016 vi
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CONTENTS
BACKGROUND iii
ABBREVIATIONS AND ACRONYMS xiv
EXECUTIVE SUMMARY ES-1
1 IDENTITY: CHEMICAL AND PHYSICAL PROPERTIES 1-1
2 TOXICOKINETICS 2-1
2.1 Absorption 2-1
2.1.1 Oral Exposure 2-2
2.1.2 Inhalation Exposure 2-2
2.1.3 Dermal Exposure 2-3
2.2 Distribution 2-3
2.2.1 Oral Exposure 2-7
2.3 Metabolism 2-23
2.4 Excretion 2-23
2.5 Animal Studies 2-24
2.5.1 Mechanistic Studies of Renal Excretion 2-28
2.6 Toxicokinetic Considerations 2-34
2.6.1 PK Models 2-34
2.6.2 Half-Life Data 2-46
2.6.3 Volume of Distribution Data 2-51
2.6.4 Toxicokinetic Summary 2-51
3 HAZARD IDENTIFICATION 3-1
3.1 Human Studies 3-1
3.1.1 Noncancer 3-3
3.1.1.1 Serum Lipids and Cardiovascular Diseases 3-3
3.1.1.2 Cardiovascular Diseases 3-13
3.1.1.3 Liver Enzymes and Liver Disease 3-13
3.1.1.4 Biomarkers of Kidney Function and Kidney Disease 3-18
3.1.1.5 Immunotoxicity 3-22
3.1.1.6 Thyroid Effects 3-26
3.1.1.7 Diabetes and Related Endpoints 3-35
3.1.1.8 Reproductive and Developmental Endpoints 3-37
3.1.1.9 Steroid Hormones 3-47
3.1.1.10 Neurodevelopment 3-48
3.1.1.11 Postnatal Development 3-51
3.1.1.12 Summary and Conclusions from the Human Epidemiology Studies 3-52
3.1.2 Cancer 3-55
3.1.2.1 Summary and Conclusions from the Human Cancer Epidemiology Studies .. 3-60
3.2 Animal Studies 3-60
3.2.1 Acute Toxicity 3-61
3.2.2 Short-Term Studies 3-62
Perfluorooctanoic Acid - May 2016 vii
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3.2.3 Subchronic Studies 3-74
3.2.4 Neurotoxicity 3-77
3.2.5 Developmental/Reproductive Toxicity 3-79
3.2.6 Prenatal Development 3-84
3.2.7 Mammary Gland Development and Other Specialized Developmental Studies ... 3-89
3.2.8 Chronic Toxicity 3-102
3.2.9 Carcinogenicity 3-106
3.3 Other Key Data 3-110
3.3.1 Mutagenicity and Genotoxicity 3-110
3.3.2 Immunotoxicity 3-112
3.3.3 Hormone Disruption 3-124
3.3.4 Physiological or Mechanistic Studies 3-129
3.3.5 Structure-Activity Relationship 3-139
3.4 Hazard Characterization 3-140
3.4.1 Synthesis and Evaluation of Major Noncancer Effects 3-140
3.4.2 Synthesis and Evaluation of Carcinogenic Effects 3-150
3.4.3 Mode of Action and Implications in Cancer Assessment 3-152
3.4.4 Weight of Evidence Evaluation for Carcinogenicity 3-158
3.4.5 Potentially Sensitive Populations 3-159
4 DOSE-RESPONSE ASSESSMENT 4-1
4.1 Dose-Response for Noncancer Effects 4-1
4.1.1 RfD Determination 4-1
4.1.1.1 PK Model approach 4-8
4.1.1.2 RfD Quantification 4-14
4.1.2 RfD Selection 4-16
4.1.3 RfC Determination 4-17
4.2 Dose-Response for Cancer Effects 4-18
5 REFERENCES 5-1
Appendix A: Literature Search Strategy Developing the Search A-l
Appendix B: Studies Evaluated Since August 2014 B-l
Appendix C: Multistage Model for Leydig Cell Tumors C-l
Perfluorooctanoic Acid - May 2016 viii
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TABLES
Table 1-1. Chemical and Physical Properties of PFOA 1-3
Table 2-1. Protein Binding in Rat, Human, and Monkey Plasma 2-4
Table 2-2. Dissociation Constants (Kd) of Binding Between PFOA and Albumin 2-4
Table 2-3. Percent (%) Binding of PFOA to Human Plasma Protein Fractions 2-6
Table 2-4. Tissue Distribution of PFOA in Wistar Rats After 28 Days of Treatment 2-9
Table 2-5. Distribution of PFOA in Male Sprague-Dawley Rats After Oral Exposure
Dose 2-10
Table 2-6. Distribution of PFOA in Female Sprague-Dawley Rats after Oral Exposure
Dose 2-11
Table 2-7. PFOA Concentrations in Wild-type and PPARa-null Mice (ug/mL) 2-12
Table 2-8. Plasma PFOA Concentrations (|ig/ml) in Postweaning Sprague-Dawley Rats 2-14
Table 2-9. Plasma PFOA Concentrations in Male Rats 2-14
Table 2-10. Plasma PFOA Concentrations in Female Rats 2-15
Table 2-11. Maternal Plasma PFOA Levels (jig/ml) in Rats During Gestation and
Lactation 2-17
Table 2-12. Placenta, Amniotic Fluid, and Embryo/Fetus PFOA Concentrations in Rats
(jig/ml) 2-17
Table 2-13. Fetus/Pup PFOA Concentration (jig/ml) in Rats During Gestation and
Lactation 2-18
Table 2-14. PFOA Levels (jig/ml) in Rats Maternal Milk During Lactation 2-18
Table 2-15. PFOA Levels (ng/ml) in Mice During Gestation and Lactation in Selected
Fluids and Tissues 2-19
Table 2-16. Female Offspring PFOA Levels (ng/ml) in Mice After GD 1-17 Exposure 2-20
Table 2-17. Female Offspring Serum PFOA Levels (ng/ml) in Mice After GD 10-17
Exposure 2-21
Table 2-18. Serum PFOA Levels (ng/ml) in Mice Over Three Generations 2-21
Table 2-19. Urine PFOA Concentrations in Male and Female Rats 2-25
Table 2-20. Cumulative Percent 14C-PFOA Excreted in Urine and Feces by Rats 2-26
Table 2-21. Cumulative Percent 14C-PFOA Excreted in Urine and Feces 2-27
Table 2-22. Kinetic Parameters of Perfluorinated Carboxylate Transport by OAT1,
OAT3, and OATPlal 2-33
Table 2-23. Plasma and Urine PFOA Concentration 24-hr After Treatment with 30 mg/kg
PFOA 2-34
Table 2-24. Model Parameters for 1 and 10 mg/kg Single Doses of PFOA to CD1 Mice 2-37
Table 2-25. Pharmacokinetic Parameters from Wambaugh et al. (2013) Meta-Analysis of
Literature Data 2-46
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Table 2-26. PK Parameters in Male Rats Following Administration of PFOA 2-50
Table 2-27. PK Parameters in Female Rats Following Administration of PFOA 2-50
Table 3-1. Summary of PFOA Occupational Exposure Studies of PFOA and Serum
Lipids 3-7
Table 3-2. Summary of High-Exposure Community Studies of PFOA and Serum Lipids 3-9
Table 3-3. Summary of General Population Epidemiology Studies of PFOA with Serum
Lipids 3-12
Table 3-4. Summary of Epidemiology Studies of PFOA and Liver Enzymes 3-14
Table 3-5. Summary of Epidemiology Studies of PFOA and Measures of Kidney
Function 3-18
Table 3-6. Summary of Epidemiology Studies of PFOA and Immune Suppression
(Vaccine Response) 3-23
Table 3-7. Summary of Epidemiology Studies of PFOA and Thyroid Effects in Adults 3-27
Table 3-8. Summary of Epidemiology Studies of PFOA and Thyroid Effects in Special
Populations 3-29
Table 3-9. Summary of Epidemiology Studies of PFOA and Pregnancy-Induced
Hypertension or Preeclampsia 3-39
Table 3-10. Summary of Epidemiology Studies of PFOA and Birth Weight 3-41
Table 3-11. Summary of Epidemiology Studies of PFOA and Pubertal Development 3-45
Table 3-12. Summary of PFOA Epidemiology Studies of Kidney and Testicular Cancer 3-56
Table 3-13. Comparison of PPAR-a Related Effects in Rats for PFOA, DEHP, and
Fenofibrate after a 3-day Exposure 3-62
Table 3-14. Hepatic Effects of Rats Exposed to PFOA 3-65
Table 3-15. Hepatic Effects in PFOA-Treated Mice 3-68
Table 3-16. Mouse Hepatocyte Ultrastructure After PFOA or Wythe 14,643 Treatment 3-69
Table 3-17. Relative Response of hPPARa, mPPARa, and PPARa-null Mice to PFOA 3-70
Table 3-18. Liver Effects in Male Rats 3-76
Table 3-19. Organ Weight Data from FO Male Rats 3-80
Table 3-20. Organ Weight Data from Fl Male Rats 3-82
Table 3-21. Studies of Pregnant CD-I Mice Following Administration of PFOA 3-90
Table 3-22. Mammary Gland Measurements at PND 21 from Female Offspring of Dams
Treated GD 10-17 3-95
Table 3-23. Dosing Regimens Used in the Multigeneration Study of CD-I Mice 3-96
Table 3-24. Mammary Gland Scores from Three Generations of CD-I Female Mice 3-97
Table 3-25. Liver Weight Data in Monkeys Administered PFOA for 6 Months 3-103
Table 3-26. Subcellular Liver Enzyme Activities and Liver PFOA Concentrations 3-104
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Table 3-27. Clinical Chemistry Values from Male Rats Given PFOA for 2 Years 3-105
Table 3-28. Incidence of Nonneoplastic Lesions in Rats Given PFOA for 2 Years 3-106
Table 3-29. Incidence of Ovarian Stromal Hyperplasia and Adenoma in Rats 3-107
Table 3-30. Mammary Gland Tumor Incidence Comparison 3-108
Table 3-31. Liver Tumors in Three Strains of Mice at 18 Months with Exposure to PFOA
Only during Gestation and Lactation 3-109
Table 3-32. Genotoxicity of PFOA In Vitro 3-111
Table 3-33. Selected Clinical Chemistry Parameters in Mice Treated with PFOA for 5
Days 3-118
Table 3-34. Selected Clinical Chemistry Parameters in Mice Treated with PFOA for
10 Days 3-119
Table 3-35. Impact of PFOA on Splenic and Thymic Lymphocyte Populations 3-120
Table 3-36. Estimated ECso Values 3-126
Table 3-37. Data Collection for Female Mice Gestationally Exposed to PFOA 3-127
Table 3-38. mRNA Expression of Hepatic PPARa and Related Genes 3-132
Table 3-39. Activation of Mouse and Human PPAR by PFOA 3-135
Table 4-1. NOAEL/LOAEL Data for Oral Subchronic and Chronic Studies of PFOA 4-3
Table 4-2. Shorter-term and Developmental Oral Exposure Studies 4-5
Table 4-3. Predicted Final Serum Concentration and Time-Integrated Serum
Concentration (AUC) for Studies in Rats 4-9
Table 4-4. Predicted Final Serum Concentration and Time-Integrated Serum
Concentration (AUC) for Studies in Mice 4-10
Table 4-5. Predicted Final Serum Concentration and Time-Integrated Serum (AUC) in
Studies of Monkeys 4-10
Table 4-6. Average Serum Concentrations Derived from the AUC and the Duration of
Dosing 4-11
Table 4-7. Comparison of Average Serum Concentration and Steady-State Concentration 4-12
Table 4-8. HEDs Derived from the Modeled Animal Average Serum Values 4-13
Table 4-9. The Impact of Quantification Approach on the RfD Outcomes for the HEDs
from thePK Model Average Serum Values 4-15
Table 4-10. Summary of Tumor Data from Animal Studies 4-19
Table 4-11. Multistage Cancer Model Dose Prediction Results for a 4% Increase in LCT
Incidence 4-21
Table B-1. PFOS Epi PapersPost Peer Review (Retrieved and Reviewed) B-1
Table B-2. PFOA Post Peer Review Animal Toxicity Studies B-3
Table B-3. Toxicokinetics: Post Peer Review B-5
Table B-4. Association between Serum PFOA and Serum Lipids and Uric Acid B-6
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Table B-5. Association of Serum PFOA and Biochemical and Hematological Measures B-8
Table B-6. Association between PFOA level and prevalence of thyroid disease and
thyroid hormone levels B-9
Table B-7. Association between Serum PFOA and Markers of Immunotoxicity B-ll
Table B-8. Association between Serum PFOA and Reproductive and Developmental
Outcomes B-12
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FIGURES
Figure 1-1. Chemical Structures of PFOA and APFO 1-1
Figure 1-2. PFOA Anti-Conformer 1-1
Figure 1-3. PFOA Lowest Energy Conformer 1-2
Figure 2-1. PFOA Binding Sites on HSA 2-5
Figure 2-2. Localization of Transport Proteins 2-29
Figure 2-3. Schematic for a Physiologically Motivated Renal Resorptions PK Model 2-35
Figure 2-4. Physiologically Motivated Pharmacokinetic Model Schematic for PFOA-
Exposed Rats 2-36
Figure 2-5. Schematic for One-Compartment Model 2-37
Figure 2-6. PK Model of Gestation and Lactation in Mice 2-39
Figure 2-7. Structure of the PFOA PBPK Model in Monkeys and Humans 2-40
Figure 2-8. Structure of the PBPK Model for PFOA in the Adult Sprague-Dawley Rat 2-42
Figure 2-9. PBPK Model Structure for Simulating PFOA and PFOS Exposure During
Pregnancy in Humans (Maternal, Left; Fetal, Right) 2-43
Figure 3-1. PPARa Agonist Mo A for Liver Tumors 3-154
Figure 4-l.BMD Model Results for LCTs (Butenhoff et al. 2012) 4-21
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ABBREVIATIONS AND ACRONYMS
8-OH-dG
Acotl
Acox
ADHD
ADX
AIC
ALP
ALT
ANOVA
APFO
Areg
AST
ATP
AUC
AUClNF
AUCiNF/D
BAX
BMD
BMDL
BMDS
BMI
BrdU
BSA
BSEP
BSP
BUN
bw
C
Cmax
CaMKII
CAR
CAT
CCK
CCL
CCL3
CFSE
ChAT
CHO
CI
CL
C1P
CLR
CoA
ConA
COPD
CORT
8-hydroxydeoxyguanosine
acyl-CoA thioesterase (human)
acyl-CoA oxidase
attention deficit hyperactivity disorder
adrenalectomized
Akaike's Information Criterion
alkaline phosphatase
alanine aminotransferase
analysis of variance
ammonium perfluorooctanoate
amphiregulin
aspartate aminotransferase
adenosine triphosphate
area under the plasma concentration time curve
area under the plasma concentration time curve, extrapolated to infinity
area under the plasma concentration time curve, extrapolated to infinity,
normalized to dose
BCL2-associated X protein
benchmark dose
lower 95th percentile confidence bound on benchmark dose
Benchmark Dose Software
body mass index
Bromodeoxyuridine (5-bromo-2-deoxyuridine)
bovine serum albumin
bile salt export pump
sulfobromophthalein
blood urea nitrogen
body weight
Celsius
peak plasma concentration at the first intestinal absorption loci
calcium/calmodulin-dependent protein kinase II
constitutive androstane receptor
carnitine acyltransferase
cholecystokinin
Contaminant Candidate List
Contaminant Candidate List 3
6-carboxyfluorescein succinimidyl ester
choline acetyltransferase
Chinese hamster ovary
confidence interval
clearance
plasma clearance
renal clearance
coenzyme A
concanavalinA
chronic obstructive pulmonary disease
corticosterone
Perfluorooctanoic Acid - May 2016
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Cox II
Cox IV
CPT
Crl
CSF
Cte
CYP4A10
d
DCDQ
Ddit3
DEHP
DHT
dL
DMSO
DNA
DR
DTK
DWI
E2
ESS
ECso
ECF
eGFR
EGFR
EPA
ER
ERa
Erra
FID
FSH
FT4
FXR
g
GAP-43
GD
GEE
GFR
GGT
GJIC
GlyT
GnRH
GSD
GST
hCG
HDL
HED
HEK
HESD
HET
cytochrome c oxidase subunit II
cytochrome c oxidase subunit IV
carnitine palmitoyltransferase
Charles River Laboratory
cancer slope factor
acyl-CoA thioesterase (rat)
cytochrome P450 4alO
day
Developmental Coordination Disorder Questionnaire
DNA damage inducible transcript
di(2-ethylhexyl) phthalate
5a-dihydroxy-testosterone
deciliter
dimethyl sulfoxide
deoxyribonucleic acid
dose rate
delayed-type hypersensitivity
drinking water intake
17-P estradiol
estrone-3-sulfate
half maximal effective concentration
electrochemical fluorination
estimated glomerular filtration rate
epidermal growth factor receptor
U.S. Environmental Protection Agency
endoplasmic reticulum
estrogen receptor a
estrogen-related receptor a
flame ionization detector
follicle-stimulating hormone
free thyroxine
farnesoid receptor
gram
growth-associated protein-43
gestation day
generalized estimating equation
glomerular filtration rate
gamma-glutamyl transpeptidase
gap junction intercellular communication
glycogen trophoblast cell
gonadotropin releasing hormone
geometric standard deviation
glutathione-S-transferase
human chorionic gonadotropin
high-density lipoprotein
human equivalent dose
human embryonic kidney
health effects support document
heterozygous
Perfluorooctanoic Acid - May 2016
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HFD
HGFa
HL-60
HMG-CoA
HPLC/MS
HPLC/MS/MS
HR
HRBC
HSA
HSD17P1
HSD3P1
ICso
ICR
IDL
IgE
IGF-I
IgM
IHD
IL-6
INUENDO
IQR
IRR
IU
IV
Ka
Kd
Ke
kg
Km
-K.OC
Kt
L
L-FABP
LCso
LCT
LD
LDso
LDH
LDL
LH
LHWA
LLOQ
LOAEL
LOD
LOQ
LPS
m
MCAD
MDA
high-fat diet
hepatocyte growth factor
human promyelocytic leukemia cell line
3-hydroxy-3-methylglutaryl coenzyme A
High-performance liquid chromatography mass spectrometry
High-performance liquid chromatography tandem mass spectrometry
hazard ratio
horse red blood cells
human serum albumin
hydroxysteroid 1?P dehydrogenase 1
hydroxysteroid 3p dehydrogenase 1
half-maximal inhibiting concentration
imprinting control region
intermediate density lipoprotein
immunoglobulin E
insulin like growth factor I
immunoglobulin M
ischemic heart disease
interleukin 6
Biopersistent Organochlorines in Diet and Human Fertility study
interquartile range
incidence rate ratio
international unit
intravenous
adsorption rate constant
dissociation constant
elimination rate constant
kilogram
substrate concentration at which the initial reaction rate is half maximal
organic carbon water partitioning coefficient
affinity constant
liter
liver fatty acid binding protein
lethal concentration for 50% of animals
Leydig cell tumor
lactation day
lethal dose for 50% of animals
lactic dehydrogenase
low-density lipoprotein
luteinizing hormone
Little Hocking Water Association
lower limit of quantification
lowest observed adverse effect level
limit of detection
limit of quantitation
lipopolysaccharide
meter
medium chain acyl-CoA dehydrogenase
malondialdehyde
Perfluorooctanoic Acid - May 2016
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Mdr2
US
mg
min
mL
jimol
MMAD
MOA
MoA
mol
mPL
mPLP
mRNA
MRP
MTT
Nd2
NdufsS
ng
NHANES
NIEHS
NJDEP
NK
NM
nmol
NMR
NMRI
NOAEL
Nrfl
Nrf2
NTCP
OAT
OATP
OR
OVA
OVX
OW
P
PACT
PAH
PB
PBMC
PBPK
PCNA
PenH
PFAA
PFAS
PFC
PFDA
multidrug resistance protein 2
microgram
milligram
minute
milliliter
micrometer
micromole
mass median aerodynamic diameter
mechanism of action
mode of action
mole
mouse placental lactogen
mouse prolactin-like protein
messenger ribonucleic acid
multidrug resistance-associated protein
3-(4,5-dimethylthiazol-2-yl)-2,5-diphenyltetrazolium bromide
NADH dehydrogenase 2
NADH dehydrogenase iron-sulfur protein 8
nanogram
National Health and Nutrition Examination Survey
National Institute for Environmental Health Sciences
New Jersey Department of Environmental Protection
natural killer
not monitored
nanomolar
nuclear magnetic resonance
Naval Medical Research Institute
no observed adverse effect level
nuclear respiratory factor 1
nuclear respiratory factor 2
sodium-taurocholate contransporting polypeptide
organic anion transporter
organic anion transporting polypeptide
odds ratio
ovalbumin
ovariectomized
Office of Water
progesterone
pancreatic acinar cell tumor
polycyclic aromatic hydrocarbon
phenobarbital
peripheral blood mononuclear cells
physiologically based pharmacokinetic
proliferating cell nuclear antigen
enhanced pause airway respiration
perfluoroalkyl acid
perfluoroalkyl substance
plaque-forming cell
perfluorodecanoic acid
Perfluorooctanoic Acid - May 2016
XVII
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PFHxA
PFHxS
PFNA
PFOA
PFOS
Pgc-la
PH
PHA
PK
pKa
PND
POD
PPAR
ppb
ppm
PT
PWG
PXR
Q
Qfiic
RfC
RfD
RFD
ROS
RR
RSA
RSC
RT-PCR
RXRa
SD
SDH
SDQ
SDWA
SHBG
STAR
SIR
SMR
SOD
SPI
SRBC
S-TGC
T3
T4
Tl/2
Tm
Imax
TC
TCPOBOP
Tfam
perfluorohexanoic acid
perfluorohexanesulfonic acid
perfluorononanoic acid
perfluorooctanoic acid
perfluorooctane sulfonate
peroxisome proliferator-activated receptor gamma coactivator la
peroxisomal bifunctional protein
phytohemagglutinin
pharmacokinetic
acid dissociation constant
postnatal day
point of departure
peroxisome proliferator-activated receptor
parts per billion
parts per million
peroxisomal thiolase
pathology working group
pregnane X receptor
flow in and out of tissues
median fraction of blood flow to the filtrate
reference concentration
reference dose
regular fat diet
reactive oxygen species
relative risk
rodent serum albumin
relative source contribution
reverse transcription polymerase chain reaction
retinoid X receptor alpha
standard deviation
sorbitol dehydrogenase
Strengths and Difficulties Questionnaire
Safe Drinking Water Act
sex hormone-binding globulin
SIDS Initial Assessment Report
standardized incidence ratio
standardized mortality ratio
superoxide dismutase
Society of the Plastics Industry
sheep red blood cells
sinusoidal trophoblast giant cells
triiodothyronine
thyroxine
elimination half-time
transporter maximum
time of maximum plasma concentration
total cholesterol
l,4-bis[2-(3,5-dichloropyridyloxy)] benzene
transcription factor A
Perfluorooctanoic Acid - May 2016
xvm
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TG
TH
TNFa
TPO
TRR
TSH
TIP
TTR
UA
UCMR
UCMR1
UCMR 2
UF
URAT
USGS
Vd
Vmax
VLCAD
VLDL
VOC
WHO
WRF
triglyceride
tyrosine hydroxylase
tumor necrosis factor a
thyroid peroxidise
total reactive residues
thyroid stimulating hormone
time to pregnancy
thyroid hormone transport protein, transthyretin
uric acid
Unregulated Contaminant Monitoring Rule
Unregulated Contaminant Monitoring Rule 1
Unregulated Contaminant Monitoring Rule 2
uncertainty factor
urate transporter
U.S. Geological Survey
volume of distribution
maximum initial rate of an enzyme catalyzed reaction
very long chain acyl-CoA dehydrogenase
very low-density lipoprotein
volatile organic compound
World Health Organization
Water Research Foundation
Perfluorooctanoic Acid - May 2016
xix
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EXECUTIVE SUMMARY
Perfluorooctanoic acid (PFOA) is a synthetic, fully fluorinated, organic acid used in a variety
of consumer products and in the production of fluoropolymers and generated as a degradation
product of other perfluorinated compounds. Because of strong carbon-fluorine bonds, PFOA is
stable to metabolic and environmental degradation. PFOA is one of a large group of
perfluoroalkyl substances (PFASs) that are used to make products more resistant to stains,
grease, and water. These compounds have been widely found in consumer and industrial
products as well as in food items. Major U.S. manufacturers voluntarily agreed to phase out
production of PFOA by the end of 2015. Exposure to PFOA in the United States remains
possible due to its legacy uses, existing and legacy uses on imported goods, degradation of
precursors, and extremely high persistence in the environment and the human body.
Extensive data on humans and animals indicate ready absorption of PFOA and distribution of
the chemical throughout the body by noncovalent binding to plasma proteins. Studies of
postmortem human tissues identify its presence in liver, lung, kidney, and bone. PFOA is not
readily eliminated from the human body as evidenced by the half-life of 2.3 years among
members of the general population. In contrast, half-life values for the monkey, rat, and mouse
are 20.8 days, 11.5 days, and 15.6 days, respectively.
Human epidemiology data report associations between PFOA exposure and high cholesterol,
increased liver enzymes, decreased vaccination response, thyroid disorders, pregnancy-induced
hypertension and preeclampsia, and cancer (testicular and kidney). Epidemiology studies
examined workers at PFOA production plants, a high-exposure community population near a
production plant in the United States (i.e., the C8 cohort), and members of the general population
in the United States, Europe, and Asia. These studies examined the relationship between serum
PFOA concentration (or other measures of PFOA exposure) and various health outcomes.
Exposures in the highly exposed C8 community are based on the concentrations in contaminated
drinking water and serum measures. Exposures among the general population typically included
multiple PFASs as indicated by serum measurements. The correlation among eight carbon
PFASs is often moderately strong (e.g., Spearman r > 0.6 for PFOA and perfluorooctane
sulfonate (PFOS) in the general population). Mean serum levels among the occupational cohorts
ranged from approximately 1 to 4 micrograms per milliliter (|ig/mL) and in the C8 cohort ranged
from 0.01 to 0.10 |ig/mL. Geometric mean serum values for the National Health and Nutrition
Examination Survey (NHANES) general population (> age 12; 2003-2008) were 0.0045 ug/mL
for males and 0.0036 ug/mL for females.
These epidemiology studies have generally found positive associations between serum PFOA
concentration and total cholesterol (TC) in the PFOA-exposed workers and the high-exposure
community (i.e., increasing lipid level with increasing PFOA); similar patterns are seen with
low-density lipoproteins (LDLs) but not with high-density lipoproteins (HDLs). These
associations were seen in most of the general population studies, but similar results also were
seen with PFOS, and the studies did not always adjust for these correlations. Associations
between serum PFOA concentrations and elevations in serum levels of alanine aminotransferase
(ALT) and gamma-glutamyl transpeptidase (GOT) were consistently observed in occupational
cohorts, the high-exposure community, and the U.S. general population. The associations are not
large in magnitude, but indicate the potential for PFOA to affect liver function.
Perfluorooctanoic acid (PFOA) - May 2016 ES-1
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Diagnosed thyroid disease in females and female children was increased both in the high-
exposure C8 study population and in females with background exposure; thyroid hormones are
not consistently associated with PFOA concentration. Associations between PFOA exposure
and risk of infectious diseases (as a marker of immune suppression) were not identified, but a
decreased response to vaccines in relation to PFOA exposure was reported in studies in adults
in the high-exposure community population and in studies of children in the general
population; in the latter studies, it is difficult to distinguish associations with PFOA from those
of other correlated PFASs. Studies in the high-exposure community reported an association
between serum PFOA and risk of pregnancy-related hypertension or preeclampsia, conditions
related to renal function during pregnancy; this outcome has not been examined in other
populations. An inverse association between maternal PFOA (measured during the second or
third trimester) or cord blood PFOA concentrations and birth weight was seen in several
studies. It has been suggested that low glomerular filtration rate (GFR) can impact fetal birth
weight (Morken et al. 2014). Pharmacokinetic (PK) analyses have shown, however, that in
individuals with low GFR, there are increased levels of serum PFOA and lower birth weights.
Thus, the impact on body weight is likely due to a combination of the low GFR and the serum
PFOA.
The epidemiology studies did not find associations between PFOA and diabetes,
neurodevelopmental effects, or preterm birth and other complications of pregnancy.
Developmental outcomes including delayed puberty onset in girls also have been reported;
however, in the two studies examining PFOA exposure in relation to menarche, conflicting
results were observed: either no association or a possible indication of an earlier menarche seen
with higher maternal PFOA levels in one study and a later menarche seen with higher maternal
PFOA levels in the other study. Increased risk of ulcerative colitis was reported in the high-
exposure community study as well as in a study limited to workers in that population.
For PFOA, oral animal studies of short-term subchronic and chronic duration are available in
multiple species including monkeys, rats, and mice. These studies report developmental effects,
liver and kidney toxicity, immune effects, and cancer (liver, testicular, and pancreatic).
Developmental effects observed in animals include decreased survival, delayed eye opening and
reduced ossification, skeletal defects, altered puberty (delayed vaginal opening in females and
accelerated puberty in males), and altered mammary gland development.
In most animal studies, changes in relative and/or absolute liver weight appear to be the most
common effect observed with or without other hepatic indicators of adversity, identifying
increased liver weight as a common indicator of PFOA exposure. The liver also contains the
highest levels of PFOA when analyzed after test animal sacrifice. The increases in liver weight
and hypertrophy, however, also can be associated with activation of cellular peroxisome
proliferator-activated receptor a (PPARa) receptors, making it difficult to determine if this
change is a reflection of PPARa activation or an indication of PFOA toxicity. The PPARa
response is greater in rodents than it is in humans. The U.S. Environmental Protection Agency
(EPA) evaluated liver disease and liver function resulting from PFOA exposure in studies where
liver weight changes and other indicators of adversity such as necrosis, inflammation, fibrosis,
and/or steatosis (fat accumulation in the liver) or increases in liver or serum enzymes indicative
of liver damage were observed.
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In repeat PFOA dosing studies, rats given 0.64 milligrams per kilogram per day (mg/kg/day)
for 13 weeks and monkeys given 3 mg/kg/day for 26 weeks had increased liver weight
accompanied by hepatocellular hypertrophy. As part of a two-generation study, male rats had
increased liver and kidney weights as well as decreased body weight at 1 mg/kg/day. In shorter
term studies, slightly higher or lower doses to rats resulted in increased liver weight, liver
necrosis, and developmental delays. In mice, developmental toxicity and increased spleen weight
was observed at a dose of 1 mg/kg/day accompanied by increased liver weight. Other doses of
similar magnitudes in mice were associated with developmental delays and liver necrosis.
Slightly higher doses resulted in decreased immunoglobulin levels. As supported by the
epidemiology data, suppression of the immune system in response to PFOA exposure is an area
of concern for humans as well as animals.
PFOA is known to activate PPAR pathways by increasing transcription of mitochondrial and
peroxisomal lipid metabolism, sterol, and bile acid biosynthesis and retinol metabolism genes.
Based on PFOA-induced transcriptional activation of many other genes in PPARa-null mice,
however, other receptors such as the constitutive androstane receptor (CAR), farnesoid receptor
(FXR), and pregnane X receptor (PXR) could be involved in PFOA-induced toxicity.
EPA used a peer-reviewed PK model to calculate the average serum concentrations
associated with candidate no observed adverse effect levels (NOAELs) and lowest observed
adverse effect levels (LOAELs) from six studies for multiple effects to calculate corresponding
human equivalent doses (HEDs) for the derivation of candidate reference doses (RfDs). Overall,
the toxicity studies available for PFOA demonstrate that the developing fetus is particularly
sensitive to PFOA-induced toxicity. In addition to the critical developmental effects described
above, other adverse effects include decreased survival, delays in eye opening and ossification,
skeletal defects, delayed vaginal opening in females, and altered mammary gland development.
The EPA Office of Water (OW) selected an RfD of 0.00002 mg/kg/day based on effects
observed in a developmental toxicity study in mice for PFOA (Lau et al. 2006). The RfD is
based on reduced ossification and accelerated puberty (in males). The total uncertainty factor
(UF) applied to the HED LOAEL from Lau et al. (2006) is 300 and includes a UF of 10 for
intrahuman variability, a UF of 3 to account for toxicodynamic differences between animals and
humans, and a UF of 10 to account for use of a LOAEL as the point of departure (POD).
Decreased pup body weights also were observed in studies conducted in mice receiving
external doses within the same order of magnitude (1,3, and 5 mg/kg/day, respectively) as those
chosen for the RfD. These studies, however, lacked serum levels and were not amenable to
modeling. Overall, the developmental and reproductive toxicity studies available for PFOA
demonstrate that the developing fetus is particularly sensitive to PFOA-induced toxicity. The
selected RfD is supported by the other candidate RfDs (also 0.00002 mg/kg/day) based on
effects on the immune system in a 15-day short-term study by DeWitt et al. (2008) and on the
kidneys of FO and Fl males in a two-generation study of developmental and reproductive
toxicity.
Under EPA's Guidelines for Carcinogen Risk Assessment (USEPA 2005a), there is
"suggestive evidence of carcinogenic potential" for PFOA. Epidemiology studies demonstrate an
association of serum PFOA with kidney and testicular tumors among highly exposed members of
the general population. Two chronic bioassays of PFOA support a positive finding for its ability
to be tumorigenic in one or more organs of rats, including the liver, testes, and pancreas. EPA
estimated a cancer slope factor (CSF) of 0.07 (mg/kg/day)"1 based on testicular tumors. As a
Perfluorooctanoic acid (PFOA) - May 2016 ES-3
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comparative analysis, the concentration of PFOA in drinking water that would have a one-in-a-
million increased cancer risk was calculated using the oral slope factor for testicular tumors,
assuming a default adult body weight of 80 kg and a default drinking water intake (DWI) rate of
2.5 liter per day (L/day) (USEPA 2011). This concentration is lower than the concentration for
cancer (also derived with adult exposure values), indicating that a guideline derived from the
developmental endpoint will be protective for the cancer endpoint.
Perfluorooctanoic acid (PFOA) - May 2016 ES-4
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1 IDENTITY: CHEMICAL AND PHYSICAL PROPERTIES
Perfluorooctanoic acid (PFOA) is a completely fluorinated organic synthetic acid used to
produce fluoropolymers. It is manufactured by the Simons electrochemical fluorination (ECF)
process or by telomerization. In the ECF process, the carbon-hydrogen bonds on molecules of
organic feedstock are replaced with carbon-fluorine bonds when an electric current is passed
through a solution of hydrogen fluoride and the organic feedstock. In the telomerization process,
fluorine-bearing chemicals and tetrafluoroethylene react to produce fluorinated intermediates
that are converted into PFOA (HSDB 2006). The telomerization process produces linear chains
(Beesoon et al. 201 1). Ammonium perfluorooctanoate (APFO) is the salt of PFOA and is a
processing aid in the manufacture of certain fluoropolymers, especially as an emulsifier in
aqueous solution during the emulsion polymerization of tetrafluoroethylene (see Figure 1-1).
APFO is not consumed during the polymerization process (SPI 2005). Some sources of PFOA in
the atmosphere result from the atmospheric degradation or transformation or surface deposition
of precursors, including related fluorinated chemicals (e.g., fluorotelomer alcohols, olefms, and
perfluoroalkyl sulfonamide substances) (Wallington et al. 2006).
PFOA APFO
FFFFFFO FFFFFPO
F FF FF FF F F F F F F F F F
Source: SIAR 2006
Figure 1-1. Chemical Structures of PFOA and APFO
Although PFOA is not a polar molecule, each of the carbon fluoride bonds is a dipole as a
result of the electronegativity difference between fluoride (4.1) and carbon (2.5), placing a partial
negative charge on each of the covalently bound fluorines and a partial positive charge on each
of the fluorinated carbons. Charge repulsion of the partially negative fluorines and steric factors
favor a PFOA conformation in which carbons 2 through 7 adopt an anti arrangement of
substituents resulting in a fairly linear molecular shape as the lowest energy conformer (see
Figure 1-2 and Figure 1-3).
(CF2)Z
Figure 1-2. PFOA Anti-Conformer
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Figure 1-3. PFOA Lowest Energy Conformer
The favored PFOA conformer is very similar to the preferred conformation of the eight-
carbon fatty acid, octanoic acid (also known as caprylic acid) except for the sphere of partial
negative charge on the fluorines of the exterior surface. The ionized carboxylate grouping and
the fluorine's partial negative charges favor electrostatic interactions between PFOA and
positively charged surfaces on proteins and other macromolecules.
The ECF process produces branched chain isomers, about 80% linear and 20% branched
(Loveless et al. 2008). The samples studied by Loveless et al. (2008) had the following mole
(mol) percents of branched chain isomers: 12.6% internal monomethyl (nonalpha), 9% isopropyl,
0.2%, tert-butyl, 0.1 gem-dimethyl, and 0.1 alpha monomethyl. A study by Yoshikane et al.
(2010) reported finding perfluoro-6-methylheptanoic acid (the isopropyl isomer) using mass
spectroscopy analysis of environmental fluorosurfactants in Japan. Branched chain samples
evaluated by Beesoon and Martin (2015) had a 7 carbon linear chain with methyl groups on
carbons 3, 4, 5, or 6, designated as 3m, 4m, 5m, or 6m (iso), respectively. The composition of a
PFOA product is important because the toxicokinetic and physiological properties of the linear
and branched chain isomers are different. The nomenclature for the branched chain isomers
varied between authors and indicates that differences exist in the composition of the commercial
products that were evaluated.
The physical and chemical properties and other reference information for PFOA and its salt
APFO are provided in Table 1-1. These properties help to define the behavior of PFOA in living
systems and the environment. PFOA and its salt are highly stable compounds. They are solids at
room temperature with low vapor pressures. The melting point for PFOA is identified as
54.3 degrees Celsius, and vapor pressures increase at temperatures near the melting point.
PFOA is moderately soluble in water and APFO is even more soluble. Both compounds are
considered insoluble in nonpolar solvents, which results in their being described as olephobic.
Water solubility is increased by the presence of other ions and is an important factor governing
solubility in body fluids. As the concentration of PFOA in aqueous solution increases, it forms
colloidal micelles with the carboxyl functional groups on the exterior and the fluorocarbon chain
on the interior. The critical micelle concentration has been identified as 3.6-3.7 g/L. Once the
critical concentration has been reached, micelles will form and the PFOA molecules will
colloidally distribute in the aqueous environment. At levels below the critical micelle
concentration, the individual molecules are individually distributed in the solvent.
The acid dissociation constant (pKa) for PFOA has been reported as 2.8. As a result, it will
be present in most biological fluids (gastric secretions excluded) primarily as the
perfluorooctanoate anion. This is an important feature in governing absorption and membrane
transport.
Perfluorooctanoic acid (PFOA) - May 2016 1-2
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Table 1-1. Chemical and Physical Properties of PFOA
Property
Chemical Abstracts
Service Registry No.
(CASRN) a
CA Index Name
Synonyms
Chemical Formula
Molecular Weight (g/mol)
Color/Physical State
Boiling Point
Melting Point
Vapor Pressure
Henry's Law constant
pKa
Koc
Kow
Solubility in water
Half-life in water (25°C)
Half -life in air
Perfluorooctanoic Acid
335-67-1
2,2,3,3,4,4,5,5,6,6,7,7,8,8,8-
pentadecafluorooctanoic acid
PFOA; Pentadecafluoro-1-octanoic acid;
Pentadecafluoro-n-octanoic acid; Octanoic
acid, pentadecafluoro-; Perfluorocaprylic
acid; Pentadecafluorooctanoic acid;
Perfluoroheptanecarboxylic acid
C8HFi5O2
414.09
White powder (ammonia salt)
192.4 °C; Stable when bound
54.3 °C
0.525 mm Hg at 25 °C (measured)
0.962 mm Hg at 59.25 °C (measured)
Not measureable
2.80
2.06
Not measurable
9.50 x 103 mg/L at 25 °C (estimated)
Stable
Stable when bound
Source
(HSDB 2012); (Lide 2007); (SRC 2016)
(HSDB 2012); (Lewis 2004)
(HSDB 2012); (Lide 2007); (SRC 2016)
(HSDB 2012); (Lide 2007); (SRC 2016)
(Heksteretal. 2003); (HSDB 2012);
(SRC 2016)
(ATSDR 2015); (Kaiser et al. 2005)
(ATSDR2015)
(SRC 2016)
(Higgins and Luthy 2006)
(ATSDR 2015); (EFSA 2008)
(ATSDR 2015); (Hekster et al. 2003);
(HSDB 2012); (Kauck and Diesslin
1951); (SRC 2016)
(UNEP 2015)
(UNEP 2015)
Note: aThis CASRN is for linear PFOA, but the toxicity studies are based on a mixture of linear and branched and the RfD
applies to both.
Perfluorooctanoic acid (PFOA) - May 2016
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2 TOXICOKINETICS
PFOA is stable to metabolic and environmental degradation because of strong carbon-
fluorine bonds. It also is resistant to metabolic biotransformation. Thus, the toxicity of the parent
compound is the concern. Because of its impact on cellular receptors and proteins, it possesses
the ability to impact the biotransformation of dietary constituents, intermediate metabolites, and
other xenobiotic chemicals by altering enzyme activities and transport kinetics. PFOA is known
to activate PPAR pathways by increasing transcription of mitochondrial and peroxisomal lipid
metabolism, sterol, and bile acid biosynthesis and retinol metabolism genes. Based on
transcriptional activation of many genes in PPARa-null mice, however, the effects of PFOA
involve far more than activation of PPAR and consequent peroxisome proliferation. The data
indicate that it also can activate the CAR, FXR, and PXR and metabolic activities linked to these
nuclear receptors.
PFOA is not readily eliminated from humans and other primates. Toxicokinetic profiles and
the underlying mechanism for half-life differences are not completely understood, although
many of the differences appear to be related to elimination kinetics and factors that control
membrane transport. Thus far, three transport families appear to play a role in PFOA absorption,
distribution, and excretion: organic anion transporters (OATs), organic anion transporting
polypeptides (OATPs), and multidrug resistance-associated proteins (MRPs) (Klaassen and
Aleksunes 2010; Launay-Vacher et al. 2006). The transporters are critical for gastrointestinal
absorption, uptake by the tissues, and excretion via bile and the kidney. These transport systems
are located at the membrane surfaces of the intestines, liver, lungs, heart, blood brain barrier,
blood placental barrier, blood testes barrier, and mammary glands where they function to protect
the organs, tissues, and fetus from foreign compounds (Ito and Alcorn 2003; Klaassen and
Aleksunes 2010, Zai'r et al. 2008).
There are differences in transporters across species, genders, and individuals. For example,
more PFOA-specific information is available about the OAT and OATP families than about the
MRPs. These limitations have hindered the development of PK models for use in predicting
effects in humans based on the data from animal studies. Abbreviations for the various
transporters are not totally standardized, and there are inconsistencies across individual
publications. The current convention for distinguishing between the transporters in humans and
those in animals is to use uppercase letters for humans and lowercase letters for animals. In this
document, uppercase letters are used uniformly, thus, the abbreviations indicate the transporter
family and not the species studied.
2.1 Absorption
Absorption data are available for oral, inhalation, and dermal exposure in laboratory animals,
and extensive data are available from humans demonstrating the presence of PFOA in serum.
These data demonstrate absorption by one or more routes but do not quantify the amounts
absorbed relative to dose.
The absorption process requires transport across the interface of the gut, lung, and skin with
the external environment. Since PFOA is moderately soluble in aqueous solutions and
oleophobic (i.e., minimally soluble in body lipids), movement across the apical and basal
membrane surfaces of the lung, gastrointestinal tract, and skin involves transporters or
mechanisms other than simple diffusion across the lipid bilayer. As discussed above, there are
Perfluorooctanoic acid (PFOA) - May 2016 2-1
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data that identify involvement of OATs, OATPs, and MRPs in enterocytes in uptake of PFOA
(Klaassen and Aleksunes 2010; Zai'r et al. 2008). OAT2, OATS, OATP2bl, 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 (Zai'r et al.
2008). Few studies have been conducted of the intestinal transporters for PFOA in humans or
laboratory animals. Most of the research has focused on the kidney and has been carried out
using cultured carrier cells transfected with the transporter proteins.
2.1.1 Oral Exposure
Based on animal data, PFOA is well absorbed following oral exposure. Gibson and Johnson
(1979) administered a single dose of 14C-PFOA averaging 11.4 mg/kg by gavage to groups of
three male 10-week-old CD rats. Twenty-four hours after administration, at least 93% of total
carbon-14 was absorbed. In another study, Cui et al. (2010) exposed male Sprague-Dawley rats
(10 per group) to PFOA (96% active ingredient) at 0, 5, and 20 mg/kg/day once daily by gavage
for 28 days. The percent of the dose absorbed was 92.8% and 92.3% for the low and high dose,
respectively, under the assumption that fecal excretion over the first 24 hours after dosing was
estimated to be unabsorbed material and did not include biliary loss.
The data from studies of adverse effects on monkeys, rats, and mice receiving PFOA in
capsules, food, or drinking water demonstrate gastrointestinal absorption. In cynomolgus
monkeys, steady-state serum PFOA levels were reached within 4-6 weeks after dosing with
capsules containing 3, 10, and 20 mg/kg PFOA for 6 months (Butenhoff et al. 2004b). Urine
steady-state levels were reached after 4 weeks. Serum PFOA concentration in male rats fed diets
containing 0.06, 0.64, 1.94, and 6.5 mg PFOA/kg for 90 days was 7.1, 41, 70, and 138 ug/mL,
respectively (Perkins et al. 2004). Peak blood levels of PFOA were attained 1-2 hours following
a 25-mg/kg dose to male and female rats (Kennedy et al. 2004). Blood levels of PFOA over time
were similar in female rats given a single dose of 25 mg PFOA/kg to a female rat given 10 daily
doses of 25 mg PFOA/kg (Kennedy et al. 2004). Plasma PFOA concentrations in male Sprague-
Dawley rats fed a diet containing 300 parts per million (ppm) PFOA for 1, 7, and 28 days were
259, 234, and 252 ug/mL, respectively (Elcombe et al. 2010).
In rats, a marked gender difference in serum and tissue levels exists following PFOA
administration. Males consistently have much higher levels than females with the difference
maintained and becoming more pronounced over time. Female rats show much greater urinary
excretion of PFOA than do male rats with serum half-life values in hours for females compared
with days for males. These differences account for variability in postexposure serum
concentrations between males and females.
2.1.2 Inhalation Exposure
Hinderliter (2003) measured the serum concentrations of PFOA following single and
repeated inhalation exposures in Sprague-Dawley rats. For the single-exposure study, male and
female rats (3/gender/group) were exposed to a single nose-only exposure of an aerosol of 0, 1,
10, and 25 mg/m3 PFOA. Preliminary range-finding studies demonstrated that aerosol particle
sizes were 1.8-2.0 jim mass median aerodynamic diameter (MMAD) with geometric standard
deviations (GSDs) ranging from 1.9 to 2.1 jam. Blood samples were collected before exposure; at
0.5, 1, 3, and 6 hours during exposure; and at 1, 3, 6, 12, 18, and 24 hours after exposure. Plasma
Perfluorooctanoic acid (PFOA) - May 2016 2-2
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was analyzed by liquid chromatography-mass spectrometry (LC-MS). PFOA plasma
concentrations increased proportional to aerosol exposure concentrations.
The male plasma Cmax values were approximately 2-3 times higher than the female Cmax. The
female Cmax occurred approximately 1 hour after the exposure period with plasma concentrations
then declining. In males, Cniax was observed immediately after the exposure period ended and
persisted for up to 6 hours. The data are illustrative of absorption of PFOA via inhalation and are
consistent with the gender differences in rate of excretion.
2.1.3 Dermal Exposure
There is evidence that PFOA is absorbed following dermal exposure. Kennedy (1985) treated
rabbits and rats dermally with a total of 10 applications of PFOA at doses of 0, 20, 200, and
2,000 mg/kg. Treatment resulted in elevated blood organofluorine levels that increased in a dose-
related manner. Organofluorine was measured because, at the time of the study, reliable
analytical techniques for measuring serum or plasma PFOA were still under development.
O'Malley and Ebbens (1981) treated groups of two male and two female New Zealand White
rabbits dermally with doses of 100, 1,000, and 2,000 mg/kg PFOA for 14 days. Mortality among
the exposed animals demonstrated dermal uptake. All of the animals died at the highest dose,
three of four died in the mid-dose group, and none in the low-dose group. Although these data
demonstrate dermal absorption, they do not provide quantitative dose-response data for effects
other than mortality.
The results of in vitro percutaneous absorption studies of PFOA through rat and human skin
have been reported (Fasano et al. 2005). The permeability coefficient for PFOA was calculated
to be 3.25 ± 1.51 x 10"5 centimeters per hour (cm/h) and 9.49 ± 2.86 x 10"7 cm/h in rat and
human skin, respectively.
2.2 Distribution
Distribution of absorbed material requires vascular transport from the portal of entry to
receiving tissues. It has been suggested that PFOA circulates in the body by noncovalently
binding to plasma proteins. Several studies have investigated the binding of PFOA to plasma
proteins in rats, humans, or monkeys to gain an understanding of its absorption, distribution, and
elimination as well as information on species and gender differences.
Protein Binding. Protein binding in plasma from cynomolgus monkeys, rats, and humans was
tested with PFOA via in vitro methods (Kerstner-Wood et al. 2003). The results are summarized
in Table 2-1. Rat, human, and monkey plasma proteins were able to bind 97-100% of the PFOA
added at concentrations ranging from 1 to 500 ppm. Human serum albumin (HSA) carried the
largest portion of the PFOA among the protein components of human plasma. Serum albumin is
a common carrier of hydrophobic materials in the blood, including short- and medium-chain
fatty acids, thyroxine (T4), heme, inorganic ions, and some pharmaceuticals (Fasano et al. 2005).
Approximately 60% of the serum protein in humans and rats is albumin (Harkness and Wagner
1983; Saladin 2004). At 68%, the percentage bound to albumin in mice is slightly higher than in
humans and rats (Harkness and Wagner 1983).
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Table 2-1. Protein Binding in Rat, Human, and Monkey Plasma
PFOA Concentration
(ppm)
1
10
100
250
500
Rat (%)
-100
99.5
98.6
97.6
97.3
Monkey (%)
-100
99.8
99.8
99.8
99.5
Human (%)
-100
99.9
99.9
99.6
99.4
Source: Kerstner-Wood et al. 2003
Note: % binding values reported as "~100" reflect a nonquantifiable amount of test article in the plasma water below the
quantifiable limit <6.25 ng/mL.
Han et al. (2003) investigated the binding of PFOA to rat and human plasma proteins in vitro.
The authors concluded that there was no correlation between the PFOA persistence and binding
of the PFOA to rat serum. The primary PFOA binding protein in plasma was serum albumin.
However, the method used (ligand blotting) would not theoretically allow the identification of
low-abundance proteins with high affinity for PFOA. Further investigation of purified rodent and
HSA binding using labeled 19F nuclear magnetic resonance (NMR) allowed the calculation of
disassociation constants for PFOA binding to rodent and HSA. No significant difference in
binding to the serum albumin of rat versus human was detected (Table 2-2).
Male and female rats treated in vivo showed no gender difference in the binding of PFOA to
serum proteins though the persistence of PFOA in vivo is much greater in male than female rats.
Table 2-2. Dissociation Constants (Kd) of Binding Between PFOA and Albumin
Parameter
Kd(mM)
Kd(mM)
Number of Binding Sites
Method
NMRa
micro-SECb
micro-SECb
RSA
0.29±0.10C
0.36±0.08C
7.8 ± 1.5
HSA
0.38 ±0.04
7.2 ±1.3
Source: Han et al. 2003
Notes:
RSA = rodent serum albumin; HSA= human serum albumin
a = Average of the two Kd values (0.31 ±0.15 and 0.27 ± 0.05 mM) obtained by NMR.
b = Values were obtained from three independent experiments and their SDs are shown.
c = On the basis of the result of unpaired t-test at 95% confidence interval, the difference of Kd values determined by NMR and
micro-SEC is statistically insignificant.
Wu et al. (2009) examined the interaction of PFOA and HSA. The authors tested their
hypothesis that PFOA, after absorption, was transported bound to albumin by dialyzing PFOA
solutions in the presence and absence of HSA. In the absence of HSA, 98% of the dissolved
PFOA crossed the dialysis membrane into the dialysate within 4 hours. In the presence of HSA,
the amount of PFOA found in the dialysate after 4 hours decreased in direct proportion to the
albumin concentration, demonstrating binding to the protein. No albumin was identified in the
dialysate.
Using the dialysis data and thermodynamic considerations, the authors concluded that
albumin could bind up to 12 PFOA molecules on its surface via chemical monolayer absorption
with a 13th molecule bound noncovalently in the more hydrophobic interior of the protein. The
surface nature of the binding could well indicate potential binding to other serum proteins as
well. Circular dichroism measurements of the albumin/PFOA complex suggested a
conformational change in the protein as a result of the PFOA binding. The beta-pleated sheet
Perfluorooctanoic acid (PFOA) - May 2016
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content of the albumin decreased, and the alpha-helix content increased by 15%. These
conformational changes could interfere with the functional properties of serum albumin or other
serum proteins impacted by surface monolayers of PFOA. For example, albumin's ability to
transport its natural ligands could be decreased by the presence of PFOA on the protein surface.
The interaction of albumin with target cellular receptors also could be altered.
MacManus-Spencer et al. (2010) used a variety of approaches to quantify the binding of
PFOA to serum albumin (e.g., surface tension measurements, 19FNMR spectroscopy, and
fluorescence spectroscopy). When taken together, the results from these analyses suggest the
presence of primary and secondary binding sites on albumin. The PFOA-albumin association
constants for the primary site (Kia) was about 1.5 ± 0.2 x 105/mol bovine serum albumin (BSA)
while the association constant for the secondary site (K2a) was 0.8 ± 0.1 x 102/mol BSA at a
concentration of Ijimol. When the concentration of BSA increased to 10 jimol, the binding per
mol of BSA decreased Kia = 0.33 ± 0.004 and K2a = 0.53 ± 0.1. Qin et al. (2010) also used
fluorescence spectroscopy quenching analysis to study PFOA binding to BSA and concluded that
van der Waals forces and hydrogen bonds were the dominant intermolecular binding forces.
The results of the fluorescence spectroscopy suggested a conformational change in BSA
following binding of PFOA that moved a tryptophan residue (#214) from a slightly polar region
of the protein to a less polar region. The shift in a tryptophan position is consistent with the
observations of Wu et al. (2009) and Qin et al. (2010), who reported that BSA underwent a
conformational change following the binding of PFOA. The authors considered the results from
the fluorescence spectroscopy to be relevant to the potential physiological impact of PFOA at
levels found in the environment. Because serum albumin is a carrier for a variety of endogenous
and exogenous substrates, a change in conformation can alter the bonding constants between
albumin and other serum constituents.
A modeling study by Salvalaglio et al. (2010) was conducted to determine the binding sites
of PFOA on HSA and classify them by their interaction energy using molecular modeling; this
study builds on the binding studies of Wu et al. (2009) and MacManus-Spencer et al. (2010). It
was estimated that the maximum number of PFOA binding sites on HSA was nine. The site
locations were common to the natural binding sites for fatty acids, T4, Warfarin, indol, and
benzodiazepine (see Figure 2-1) (Salvalaglio et al. 2010). The binding site closest to tryptophan
residue #214 had the highest binding affinity (-8.0 kilocalorie/mol).
*
Indole benzodiazepine
Warfann
Figure 2-1. PFOA Binding Sites on HSA
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Weiss et al. (2009) screened several perfluorinated compounds (n = 30), differing by carbon
chain length C4-18, fluorination degree, and functional groups for potential binding to the serum
thyroid hormone transport protein, transthyretin (TTR), using a radioligand-binding assay. The
natural ligand of TTR is T4. PFOA was one of the chemicals evaluated. Human TTR was
incubated overnight with 125I-labeled T4, unlabeled T4, and 10-10,000 nanomolar (nmol) PFOA
as a competitor for the T4 binding sites. The unlabeled T4 was used as a reference compound.
The levels of T4 in the assay were close to the lower range for total T4 measured in healthy
adults. The authors concluded that binding affinity for TTR was highest for the fully fluorinated
compounds tested and those having at least an eight carbon length chain, characteristics that
apply to PFOA. PFOA demonstrated a high binding affinity for TTR with 949 nmol, causing a
50% inhibition of T4 binding to the TTR.
Beesoon and Martin (2015) examined differences in the binding of the linear and branched
chain isomers to serum albumin and human serum proteins. The linear PFOA molecule was
found to bind more strongly to calf serum albumin than the branched chain isomers. When
arranged in order of increasing binding, the order was 4m < 3m <5m < 6m (iso) 6m >4m> 3m (15-30% free). Binding was estimated based on the concentrations in the
ultrafiltrate after spiking with 5-60 mg/L of technical PFOA. The human serum was diluted
tenfold before spiking.
When incubated with separate human-derived plasma protein fractions (Kerstner-Wood et al.
2003), PFOA was highly bound (99.7%) to albumin and showed some affinity for LDLs,
formerly beta-lipoproteins (9.6%) with limited binding to alpha-globulins (11.0%) and gamma-
globulins (3.0%). Low levels of binding to alpha-2-macroglobulin and transferrin were measured
when the protein concentrations were approximately 10% of physiological concentration (see
Table 2-3).
Table 2-3. Percent (%) Binding of PFOA to Human Plasma Protein Fractions
Fraction
Albumin
Gamma-globulin
Alpha-globulin
Fibrinogen
Alpha-2-macroglobulin
Transferrin
LDLs
~10% Physiological Cone.
96.4
3.5
28.5
5.4
7.9
1.0
19.6
100% Physiological Cone.
99.7
3.0
11.0
<0.1
<0.1
2.1
39.6
Source: Kerstner-Wood et al. 2003
It also is possible that PFOA will display nonspecific binding to proteins within the cellular
matrix as well as in the serum but little work has been done to investigate that probability.
Luebker et al. (2002) conducted in vitro studies of the ability of a variety of perfluorinated
chemicals to displace a fluorescent substrate (ll-(5-dimethylamino-napthalenesulphonyl)-
undecanoic acid) from liver fatty acid binding protein (L-FABP). L-FABP is an intracellular
lipid carrier protein that reversibly binds long-chain fatty acids, phospholipids, and an assortment
of peroxisome proliferators (Erol et al. 2003). It constitutes 2-5% of the cytosolic protein in
hepatocytes. Luebker et al. (2002) reported that PFOA (ICsoMOjimol) exhibited some binding to
Perfluorooctanoic acid (PFOA) - May 2016
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L-FABP, but the binding potential was only about 50% of that for PFOS (ICso = 4.9 |imol) and
far less than that of oleic acid (ICso = 0.01 jimol).
L. Zhang et al. (2013) cloned the human L-FABP gene and used it to produce purified
protein for evaluation of the binding of PFOA and other PFASs. Nitrobenzoxadiazole-labeled
lauric acid was the fluorescent substrate used in the displacement assays. ICso values and
dissociation constants were generated for the PFASs studied. Oleic and palmitic acids served as
the normal substrates for L-FABP binding. The nitrobenzoxadiazole-labeled lauric acids
indicated that there were two distinct binding sites for fatty acids in human FABP with the
primary site having a twentyfold higher affinity than the secondary site. The ICso value for
PFOA was 9.0 ± 0.7 jimol, suggesting that it has a lower binding affinity than PFOS (ICso=3.3 ±
0.1 jimol). A similar approach was used to compare perfluorohexanoic acid (PFHxA,
perfluorohexanesulfonic acid (PFHxS), PFOA, and perfluorononanoic acid (PFNA). The affinity
of PFNA for human L-FABP was found to be greater than that for PFOA. The affinities of
PFHxA and PFHxS for the protein were much lower. Both PFOA and PFNA bound to the carrier
protein in a 1:1 ratio and the interaction was mediated by electrostatic interactions and hydrogen
binding of the PFAS with the fatty acid binding site.
2.2.1 Oral Exposure
Tissue Distribution
Human. No clinical studies are available that examined tissue distribution in humans following
administration of a controlled dose of PFOA. However, samples collected in biomonitoring and
epidemiology studies provide data showing distribution of PFOA. Olsen et al. (2001a) analyzed
human sera and postmortem liver samples and found that more than 90% of the liver samples
(n = 30) were < limit of quantitation (LOQ). Serum levels ranged from
-------
Monkey. Butenhoff et al. (2002, 2004b) studied the fate of PFOA in cynomolgus monkeys in a
6-month oral exposure study. Groups of four to six male monkeys each were administered PFOA
daily via oral capsule at dose rates (DRs) of 0, 3, 10, or 20 mg/kg. The highest dose was initially
30 mg/kg, but due to its toxicity, it was suspended after 12 days. Dosing was resumed on test day
22 using the 20 mg/kg/day dose for the remainder of the 6-month period, resulting in a
normalized dose of 20 mg/kg/day for the study. Serum, urine, and fecal samples were collected
at 2-week intervals and were analyzed for PFOA concentrations. Liver samples were collected at
time of sacrifice.
Serum concentrations reached steady-state levels within 4-6 weeks in all dose groups.
Steady-state concentrations of PFOA in serum were 77 ± 39, 86 ± 33, and 158 ± 100 |ig/ml after
6 weeks and 81 ± 40, 99 ± 50, and 156 ± 103 |ig/ml after 6 months for the 3-, 10-, and 20-mg/kg
dose groups, respectively (Butenhoff et al. 2002, 2004b). The mean serum concentration of
PFOA in control monkeys was 0.134-0.203 |ig/ml. Urine PFOA concentrations reached steady
state after 4 weeks and were 53 ± 25, 166 ± 83, and 181 ± 100 |ig/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|ig/g. Liver PFOA concentrations in two monkeys exposed to 20 mg/kg were 16.0 and
83.3 |ig/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 |ig/g (Butenhoff et al. 2004b).
Rat. Ylinen et al. (1990) administered newly weaned Wistar rats (18/gender/group) doses of 3,
10, and 30 mg/kg/day PFOA by gavage for 28 days. At necropsy, serum was collected as well as
brain, liver, kidney, lung, spleen, ovary, testis, and adipose tissue. The concentration of PFOA in
the serum and tissues was determined with capillary gas chromatography equipped with a flame
ionization detector (FID). A mass spectrometer was used in the selected ion monitoring mode
when the PFOA concentration was below the LOQ of the FID (1 jig/ml).
The concentration of PFOA in the serum and tissues following 28 days of administration is
presented in Table 2-4. PFOA was not detected in the adipose tissue. The concentrations of
PFOA in the serum and tissues were much higher in males than in females. In the males, the
levels of PFOA in the serum were generally lower in the 30 mg/kg/day dose group than in the
10 mg/kg/day dose group due to increased urinary elimination in the 30 mg/kg/day group. The
tissue levels were similar for the 10 and 30 mg/kg/day doses. In females, there was a dose-related
increase in tissue levels while the serum levels were comparable for the 10 and 30 mg/kg/day
dose groups. Among solid tissues, the liver had the highest tissue concentration followed by the
testis, spleen, lung, kidney, and brain, respectively. In females, the concentration in the kidneys
exceeded that in the liver for the 10 and 30 mg/kg/day doses but not at the lowest dose. Ovary
and spleen tissue had similar concentrations followed by lower levels in the lung and brain.
Perfluorooctanoic acid (PFOA) - May 2016 2-8
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Table 2-4. Tissue Distribution of PFOA in Wistar Rats After 28 Days of Treatment
Tissue
Serum
Liver
Kidney
Spleen
Lung
Brain
Ovary
Testis
Dose (Males3)
mg/kg/day
3
48.6 ±10.3
39.9 ±7.25
1.55 ±0.71
4.75 ±1.66
2.95 ±0.54
0.398 ±0.144
6.24 ± 2.04
10
87.27 ± 20.09
51.71 ±11. 18
40.56 ±14.94
7.59 ±3. 5
22.58 ±4.59
1.464 ±0.211
9.35 ±4.02
30
51.65 ±1.47
49.77 ±10.76
39.81 ±17.67
4.1 ±1.57
23.71 ±5.42
0.71 ±0.32
7.22 ±3. 17
Dose (Females3)
mg/kg/day
3
2.4b
1.81 ±0.49
0.06 ± 0.02
0.15 ±0.04
0.24b
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Table 2-5. Distribution of PFOA in Male Sprague-Dawley Rats After Oral Exposure Dose
Tissue
Prostate
Skin3
Blood3
Brain
Fata
Heart
Lungs
Spleen
Liver
Kidney
G.I. tract
G.I.
contents
Thyroid
Thymus
Testes
Adrenals
Muscle3
Bone3
Totalb
Img/kg
% at Tmax
0.083 ±0.039
14.77 ±2.135
22.148 ±0.692
0.071 ±0.018
2.281 ±0.467
0.451 ±0.119
0.74 ±0.147
0.086 ±0.011
21.708 ±5.627
1.949 ±0.402
2.930 ±0.929
2.083 ±0.625
0.008 ± 0.005
0.085 ±0.008
0.755 ±0.079
0.019 ±0.004
12.025 ±0.648
3.273 ±0.538
85.465 ± 6.426
% at Tmax/2
0.030 ±0.002
6.061 ±0.274
8.232 ±1.218
0.022 ± 0.002
0.593 ±0.136
0.195 ±0.024
0.341 ±0.043
0.045 ± 0.006
32.627 ±3.601
1.14 ±0.215
0.980 ±0.300
0.239 ±0.025
0.004 ± 0.003
0.051 ±0.018
0.356 ±0.037
0.010 ±0.001
4.984 ±0.745
1.120 ±0.094
57.026 ±3. 379
5mg/kg
% at Tmax
0.071 ±0.045
15. 565 ±0.899
24.919 ±1.942
0.051 ±0.021
2.815 ±0.225
0.443 ±0.037
0.593 ±0.376
0.096 ±0.017
18.750 ±2.434
2.170 ±0.354
2.508 ±0.713
2.632 ±0.934
0.011 ±0.006
0.085 ±0.012
0.693 ±0.180
0.022 ± 0.004
13. 565 ±0.576
3.047 ±0.544
88.033 ± 1.420
% at Tmax/2
0.057 ± 0.020
7.233 ±0.430
11. 140 ±0.624
0.023 ± 0.008
0.916 ±0.205
0.252 ±0.030
0.344 ±0.194
0.060 ± 0.007
25.231 ±1.289
1.212±0.115
1.052 ±0.202
0.270 ± 0.028
0.004 ± 0.002
0.053 ±0.003
0.372 ±0.062
0.009 ±0.001
6.429 ± 0.648
1.375 ±0.169
56.031 ±1.025
25 mg/kg
% at Tmax
0.067 ±0.018
13.836 ±0.969
22.905 ±1.177
0.063 ± 0.007
2.153 ±0.430
0.461 ±0.053
0.863 ±0.103
0.106 ±0.015
17.528 ±0.900
2.293 ±0.286
2.784 ± 0.608
4.186 ±1.349
0.009 ± 0.002
0.120 ±0.025
0.623 ±0.098
0.026 ± 0.004
12.855 ±0.841
3.062 ±0.438
83.937 ±3.680
% at Tmax/2
0.028 ±0.012
5.419 ±0.237
7.904 ±1.032
0.019 ±0.002
0.628 ±0.110
0.164 ±0.032
0.303 ±0.057
0.042 ± 0.005
20. 145 ±3.098
1.003 ±0.122
0.808 ±0.189
0.210 ±0.084
0.005 ±0.001
0.045 ±0.010
0.224 ±0.031
0.009 ± 0.003
4.253 ±0.358
0.906 ±0.100
42. 112 ±4.740
Source: Kemper 2003
Notes: Percent of dose recovered at Tmax and Tmax/2 in tissues.
a Percent recovery scaled to whole animal assuming the following: skin= 19%, whole blood=7.4%, fat=7%, muscle=40.4%,
bone=7.3% of body weight.
b Totals are calculated from individual animal data.
In the female tissues at Tmax/2 (4 hours), approximately 30% of the dosed PFOA retained was
present in the liver, blood, kidney, muscle, and skin tissues in decreasing amounts (Table 2-6).
About 14% of the administered dose remained in the gastrointestinal tissues and contents. Based
on the timing of the measurements and the results, females appear to absorb and excrete PFOA
more rapidly than males.
Lau et al. (2006) studied PFOA's toxicokinetic properties in rats as part of a larger study. The
authors gavage-dosed adult male and female Sprague-Dawley rats (n = 8) with 10 mg/kg for 20
days and sacrificed them 24 hours after the last treatment. After 20 days of treatment, male rats
had serum PFOA levels of 111 |ig/mL compared to 0.69 |ig/mL in female rats.
Martin et al. (2007) administered 20 mg PFOA/kg to adult male Sprague-Dawley rats (n = 4
or 5) for 1, 3, and 5 days by oral gavage and determined the liver and serum levels of PFOA.
Blood was collected via cardiac puncture and PFOA concentration was determined by high-
performance liquid chromatography-electrospray tandem mass spectrometry (LC-MS/MS). The
mean liver PFOA concentration was 92 ± 6, 250 ± 32, and 243 ± 23 ug/g after 1, 3, and 5 daily
doses of 20 mg PFOA/kg/day, respectively. The mean serum concentration was 245 ±41 ug/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.
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Table 2-6. Distribution of PFOA in Female Sprague-Dawley Rats after
Oral Exposure Dose
Tissue
Skin3
Blood3
Brain
Fata
Heart
Lungs
Spleen
Liver
Kidney
G.I. tract
G.I.
contents
Thyroid
Thymus
Ovaries
Adrenals
Muscle3
Uterus
Bone3
Totalb
Img/kg
% at Tmax
0.434 ±0.162
5.740 ±1.507
0.037 ±0.009
0.134 ±0.032
0.198 ±0.079
0.454 ±0.148
0.063 ± 0.027
7.060 ±1.266
3.288 ±0.948
10.699 ±9.066
21.956 ±13.48
0.010 ±0.003
0.052 ±0.017
0.047 ±0.019
0.014 ±0.005
0.170 ±0.051
0.243 ±0.091
0.101 ±0.017
50.698 ±16.48
% at Tmax/2
0.403 ± 0.096
4.438 ±1.625
0.047 ± 0.008
0.164 ±0.079
0.253 ±0.055
0.546 ±0.082
0.058 ±0.006
6.817 ±1.537
2.769 ± 0.784
8.462 ±6.519
3.891 ±2.395
0.016 ±0.021
0.058 ±0.024
0.048 ± 0.006
0.018 ±0.004
0.258 ±0.089
0.374 ±0.247
0.153 ±0.052
28.772 ± 10.98
5mg/kg
% at Tmax
0.624 ±0.142
8.089 ±2.080
0.066 ±0.019
0.220 ±0.111
0.388 ±0.057
0.827 ±0.102
0.101 ±0.021
11.190±2.192
4.293 ±0.771
7.142 ±2.594
2.896 ±2.305
0.008 ± 0.002
0.105 ±0.030
0.071 ±0.012
0.026 ± 0.005
0.325 ±0.010
0.354 ±0.046
0.174 ±0.057
36.897 ±3. 187
% at Tmax/2
0.307 ±0.121
5.411 ±1.466
0.045 ±0.010
0.110 ±0.069
0.236 ±0.051
0.570 ±0.179
0.060 ±0.012
7.176 ±0.982
2.685 ±0.736
8.255 ± 8.967
5.601 ±6.165
0.006 ± 0.002
0.068 ±0.021
0.041 ±0.012
0.015 ±0.004
0.229 ±0.031
0.247 ± 0.068
0.142 ±0.078
31.201 ±12.63
25 mg/kg
% at Tmax
0.380 ±0.166
7.158 ±2.232
0.058 ±0.008
0.147 ±0.053
0.317 ±0.035
0.678 ± 0.067
0.091 ±0.007
10.538 ±1.723
5.867 ±0.946
6.923 ± 1.846
2.491 ±1.548
0.009 ± 0.003
0.091 ±0.032
0.071 ±0.012
0.031 ±0.005
0.441 ±0.116
0.358 ±0.124
0.157 ±0.072
35.803 ±2.554
% at Tmax/2
0.415 ±0.175
6.407 ±1.406
0.058 ±0.018
0.148 ±0.065
0.287 ± 0.069
0.775 ± 0.204
0.070 ± 0.002
9.080 ±0.895
4.749 ±0.393
3.547 ±1.306
1.121 ±1.010
0.007 ± 0.002
0.077 ± 0.020
0.070 ±0.012
0.021 ±0.001
0.304 ±0.099
0.365 ±0.029
0.181 ±0.090
27.680 ±2.569
Source: Kemper 2003
Notes: Percent of dose recovered at Tmax and Tmax/2 in tissues.
a Percent recovery scaled to whole animal assuming the following: skin=19%, whole blood=7.4%, fat=7%, muscle=40.4%,
bone=7.3% of body weight.
b Totals are calculated from individual animal data.
Mouse. Lau et al. (2006) gavage-dosed adult male and female CD-I mice (5-7/group) with
20 mg/kg for 7 and 17 days. The animals were sacrificed 24 hours after the last treatment. After
7 days of treatment, male mice had serum PFOA levels of 181 |ig/mL and females had levels of
178 |ig/mL. After 17 days of treatment, male mice had serum PFOA levels of 199 |ig/mL and
females had levels of 171 |ig/mL. These data suggest that the gender difference observed by Lau
et al (2006) in rats was not seen in the mice under the conditions of this study.
As part of a physiologically based pharmacokinetic (PBPK) modeling exercise, Lou et al.
(2009) administered single doses of 1 and 10 mg/kg to groups of three male and three female
CD-I mice. The mice were sacrificed for analysis of plasma, liver, and kidney tissues after 4, 8,
and 12 hours and at 1, 3, 6, 9, 13, 20, 27, 34, and 48 days after dosing. This study was repeated
for a second analysis that extended the sacrifice times to 55, 62, 70, and 80 days.
Measures of PFOA in serum were presented graphically and indicate that the order of
magnitude difference between the doses led to a comparable order of magnitude difference in
serum concentrations for both males and females across the 80-day observation period. [The
study procedures indicated that serum was collected and analyzed, but the graphic presentation
described the values as plasma values. Contact with one of the authors confirmed that the values
should have been listed as serum rather than plasma.] The peak serum concentrations were
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10 and 100 mg/L for the 1 and 10 mg/kg/day doses, respectively. Declines in serum
concentrations for females over time were roughly parallel reaching concentrations of about
2 mg/L and <0.2 mg/L for the high and low doses, respectively, at the end of 80 days. Peak
serum concentrations were slightly lower in the males (~8 and 80 mg/L) than in the females, and
final serum concentrations were higher in the males (-0.5 and 8 mg/L). Liver and kidney
concentrations also were higher in males than in females for each of the two doses. These data
suggest a longer half-life in males than in females.
Lou et al. (2009) also collected serum data for up to 28 days after administration of a
60-mg/kg dose to groups of three female mice. Based on the graphic presentation of the data, the
60-mg/kg dose was cleared from the serum much more rapidly than the 1- and 10-mg/kg doses.
For example, a serum concentration of about 0.4 mg/L was reached in about 28 days for the
60-mg/kg dose, 61 days for the 10-mg/kg dose, and 70 days for the 1-mg/kg dose (values
estimated from Figure 3 in Lou et al. [2009]). No measurements were made for liver or kidney in
the high-dose animals.
In the final experimental portion of the study, Lou et al. (2009) exposed groups of five
female CD-I 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 |ig/mL reported by Lau et al. (2006).
Minata et al. (2010) orally administered 0, 12.5, 25, and 50 micromole per kilogram
(umol/kg) PFOA (~0, 5.4, 10.8, and 21.6 mg/kg PFOA) to groups of male wild-type
12984/SvlmJ mice (n = 39) and PPARa- null 129S4/SvJae-PparatmlGonz/J mice (n = 40) for
4 weeks. Blood, liver, and bile were collected for determination of PFOA concentration at the
end of 4 weeks, as shown in Table 2-7. The PFOA concentration in whole blood and the liver
were similar between wild-type and PPARa-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 umol/kg and by a factor of 2.8 from 25 to 50 umol/kg. In the bile
of PPARa-null mice, PFOA concentration increased by a factor of 3.2 from 12.5 to 25 umol/kg
and by a factor of 19.5 from 25 to 50 umol/kg. The data suggested saturation of PFOA transport
from the liver to bile ducts in wild-type mice, but not PPARa-null mice. This may indicate that
PPARa plays a role in the clearance of PFOA.
Table 2-7. PFOA Concentrations in Wild-type and PPARa-null Mice (ug/mL)
Dose
fimol/kg
0
12.5
25
50
Whole Blood
Wild-type
ND
20.6 ±2.4
46.9 ±3.2
64.2 ±6.5
PPARa-null
ND
19.3 ±2.2
36.4 ±2.7*
71.2 ±8.0
Bile
Wild-type
ND
56.8 ±26.9
784 ±137.6
2174 ±322.4
PPARa-null
ND
19.6 ±2.2
62.9 ±16.7"
383 ± 109.9"
Liver
Wild-type
ND
181.2 ±6.3
198.8 ±15.4
211.6±13.3
PPARa-null
ND
172.3 ±8.9
218.3 ±14.5
239.7 ±25.0
Source: Minata et al. 2010
Notes: Mean ± SD; ND= not detected (< 0.001 |ig/mL); > < 0.05; "p < 0.01
Tissue Transporters. As identified earlier, protein transporters from a number of families play a
role in the tissue uptake of orally ingested PFOA. The transporters are located at the interface
between the serum and the liver, kidneys, lungs, heart, brain, testes, ovaries, placenta, and uterus
(Klaassen and Aleksunes 2010). The liver is an important uptake site for PFOA. OATPs and
MRPs, at least one OAT, and the sodium-taurocholate cotransporting polypeptide (NTCP), a
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hepatic bile uptake transporter, have been identified at the interface of the liver with the portal
blood and/or the canalicular membranes within the liver (Kim 2003; Kusuhara and Sugiyama
2009; Zai'r et al. 2008).
The impact of PFOA on several membrane transporter systems linked to biliary transport was
studied by Maher et al. (2008) as part of a more detailed study of perfluorodecanoic acid
(PFDA). A dose of 80 mg/kg by intraperitoneal (i.p.) injection (propylene glycol: water vehicle)
was found to significantly increase (p<0.05) the expression of MRP3 and MRP4 in the livers of
C57BL/6 mice 2 days after treatment as reflected in quantification of their deoxyribonucleic acid
(DNA) transcripts. MRP3 and MRP4 are believed to protect the liver from accumulation of bile
acids, bilirubin, and potentially toxic exogenous substances by promoting their excretion in bile.
There were significant increases in serum bilirubin and bile acids after PFDA exposure, signifying
increased export. Conversely, there were significant decreases (p<0.05) in the protein levels for
OATPlal, OATPla4, and OATPlb2 as determined by Western Blot analysis and messenger
ribonucleic acid (mRNA) measurements following exposure to 40 mg PFOA/kg (Cheng and
Klaassen 2008). There was no significant impact on NTCP protein or the serum levels of bile
acids. The OATPs are transporters responsible for the uptake of bile acids and other hydrophobic
substances such as steroid conjugates, ecosinoids, and thyroid hormones into the liver.
These studies, all by the same laboratory, were carried out at high, single-dose exposures,
which limit their value in extrapolating to low- and repeat-dose scenarios. The results suggest a
decrease in the uptake of favored substrates into the liver and an increase in removal of favored
substrates from the liver via bile. Upregulation of MRP3 and MRP4, coupled with decreased
OATp levels, could be beneficial due to increased biliary excretion of bile acids, bilirubin, and
conjugated metabolites of toxic chemicals, including PFOA. Based on the results with the more
extensive evaluation of PFDA including mouse strains null for several receptors (PPARa, CAR,
PXR, and FXR), the authors concluded that the changes in receptor proteins were primarily
linked to activation of PPARa.
Impact of Developmental Age. Han (2003) administered groups of 4-8-week-old Sprague-
Dawley rats (10 per gender per age) a single dose of 10 mg/kg/day PFOA by oral gavage. Blood
samples were collected 24 hours after dosing and the plasma concentration of PFOA was
measured by high-performance liquid chromatography mass spectrometry (HPLC/MS). In the
4-week-old rats, the concentration of plasma PFOA was approximately 2.7 times higher in males
than in the females (Table 2-8). In the 5- and 6-week-old female rats, the plasma PFOA
concentrations were about twofold lower than in the 4-week-old rats. However, in the 5-week-
old males, the concentration of plasma PFOA was about fivefold higher than in the 4-week-old
group, suggesting a developmental change in excretion rate. Plasma concentrations did not differ
appreciably among 5-, 6-, 7-, and 8-week-old rats within each gender but did differ between
genders. In fact, PFOA plasma concentrations were 35-65-fold higher in males than in females
at every age except at 4 weeks. Thus, it appears that maturation of the transport features
responsible for the gender difference in elimination occurs between the ages of 4 and 5 weeks in
the rat.
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Table 2-8. Plasma PFOA Concentrations (ug/ml) in Postweaning Sprague-Dawley Rats
Age (weeks)
4
5
6
7
8
Males
7.32 ± l.OP
39.24 ±3. 89
43. 19 ±3.79
37.12 ±4.07
38.55 ±5.44
Females
2.68 ±0.64
1.13 ±0.46
1.18 ±0.52
0.57 ±0.29
0.81 ±0.27
Source: Han 2003
Notes:
a Mean ± SD; samples from 10 animals/gender/group
Hinderliter (2004) and Hinderliter et al. (2006a) continued the investigation of the
relationship between age and plasma PFOA in male and female Sprague-Dawley rats. Immature
rats at 3, 4, and 5 weeks of age were administered PFOA via oral gavage at a single dose of 10 or
30 mg/kg. Rats were not fasted prior to dosing. Two hours after dosing, five rats per gender per
age group and dose group were sacrificed and blood samples were collected. The remaining five
rats per gender per age and dose group were placed in metabolism cages for 24-hour urine
collection. These rats were sacrificed at 24 hours and blood samples were collected.
In the male rats, plasma PFOA concentrations for either the 10- or 30-mg/kg dosage groups
did not differ significantly by sample time (at 2 and 24 hours) or by animal age (3, 4, and
5 weeks), except at 2 hours for the 5-week-old group (p<0.01), which showed the lowest PFOA
level (Table 2-9). PFOA plasma concentrations following a 30-mg/kg dose were 2-3 times
higher than those following a 10-mg/kg dose. These data do not demonstrate a difference
between the 5-week-old rats and the younger 3- and 4-week-old groups at 24 hours after dosing,
and thus do not support the observations from the Han study (2003).
Table 2-9. Plasma PFOA Concentrations in Male Rats
Age
(weeks)
3
4
5
3
4
5
Dose
(mg/kg)
10
10
10
30
30
30
Plasma PFOA (|J.g/ml)
2 Hours Post-Dose
Mean
41.87
39.92
26.32*
120.65
117.40
65.66*
SD
4.01
4.45
6.89
12.78
18.10
15.53
24 Hours Post-Dose
Mean
34.22
42.94
40.60
74.16
100.81
113.86
SD
7.89
5.33
3.69
18.23
13.18
23.36
Source: Hinderliter 2004
Note: 'Statistically significantly different by sample time and animal age (p<0.01).
In the female rats, plasma PFOA concentrations were significantly lower in the 5-week-old
group than in the 3- or 4-week-old groups at the 24-hour time period for both doses and for the
30-mg/kg dose group at 2 hours (Table 2-10). Plasma PFOA concentrations following a
30-mg/kg dose were approximately one and one half to four times higher than those observed
following a 10-mg/kg dose.
At 24 hours post-dose, plasma PFOA levels in the female rats were significantly lower than the
plasma PFOA levels in male rats, especially at 5 weeks of age. The data for the 5-week-old female
rats compared to the 3- and 4-week-old groups at 24 hours are consistent with the Han (2003) data
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in that they demonstrate a decline in plasma levels compared to their earlier measurements. Thus,
the developmental change is one that appears to be unique to the female rat.
Table 2-10. Plasma PFOA Concentrations in Female Rats
Age
(weeks)
3
4
5
o
J
4
5
Dose
(mg/kg)
10
10
10
30
30
30
Plasma PFOA (|ig/ml)
2 Hours Post-Dose
Mean
37.87
29.88
33.23
84.86
80.67
56.90 a
SD
5.77
12.15
7.41
10.51
14.10
29.66
24 Hours Post-Dose
Mean
13.55 b
18.98b
1.36 a-b
51.43b
28.01 b
3A2*-b
SD
3.83
7.01
0.87
13.61
9.90
1.95
Source: Hinderliter 2004
Notes:
a Statistically significantly different from the 3- and 4-week values (p < 0.01).
b Statistically significantly different from 2-hour values (p < 0.01).
The data demonstrate that both dose and gender influence plasma levels. Post-dosing
clearance (CL) is slow for both doses at 2 and 24 hours in males and females at postnatal weeks
3 and 4. At 5 weeks, however, the plasma levels after 24 hours are greater than those at 2 hours
in males. In females, for the high dose at 2 hours, plasma levels are similar to those in males,
while at 24 hours they are only 3% of the value for males. This suggests that uptake from the
intestines is similar while the rate of excretion at 5 weeks and beyond is considerably greater for
female rats than males. They are comparable for postnatal weeks 3 and 4.
In a supplemental study to determine the effect of fasting (Hinderliter [2004] and Hinderliter
et al. [2006a]), 4-week-old rats, 4 rats per gender, were administered 10 mg/kg PFOA via oral
gavage. Animals (two per gender) were fasted overnight for 12 hours before dosing with PFOA.
All the rats were sacrificed at 24 hours post dosing and blood was collected for analysis of PFOA
in plasma. Plasma PFOA concentrations in male rats were 64.95 and 30.00 |ig/ml for the fasted
and nonfasted animals, respectively. Plasma PFOA concentrations in the female rats were 68.16
and 26.54 |ig/ml for the fasted and nonfasted animals, respectively. Given the consistency in the
4-week-old rat plasma PFOA concentrations, the authors concluded that age-dependent changes
in female PFOA elimination are observable between 3 and 5 weeks of age. PFOA uptake was
greater in the fasted animals than the fed animals, suggesting competition for uptake in the
presence of food components that share common transporters and/or decreased contact of PFOA
with the intestinal epithelium in the presence of dietary materials.
Distribution during Pregnancy and Lactation
Humans. T. Zhang et al. (2013) recruited pregnant females for a study to examine the
distribution of PFOA between maternal blood, cord blood, the placenta, and amniotic fluid.
Thirty-two females from Tianjin, China, volunteered to take part in the study. Samples were
collected at time of delivery. Maternal ages ranged from 21 to 39 years, gestation periods ranged
from 35 to 37 weeks. It was the first child for 26 of the females and a second child for 6. The
study yielded 31 maternal whole blood samples, 30 cord blood samples, 29 amniotic fluid
samples, and 29 placentas. The maternal blood contained variable levels of 10 PFASs, eight
acids, and two sulfonates. The mean maternal blood concentration was highest for PFOS
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(14.6 ng/mL) followed by PFOA (3.35 ng/mL). In both cases, the mean was greater than the
median, indicating a distribution skewed toward the higher concentrations.
PFOA was found in all fluids/tissues sampled. PFOA was transferred to the amniotic fluid to
a greater extent than PFOS, based on their relative proportions in the maternal blood and cord
blood. Compared to the mean PFOA blood levels in the pregnant females, the mean levels in the
cord blood, placenta, and amniotic fluid were 47%, 59%, and 1.3%, respectively, of those in the
mother's blood. The correlation coefficients between the maternal PFOA blood levels and
placenta, cord blood, and amniotic fluid levels were good (0.7-0.9) and the relationships
statistically significant (p<0.001).
Rat. An oral two-generation reproductive toxicity study of PFOA in rats was conducted
(Butenhoff et al. 2004a). Five groups of rats (30 gender/group) were administered PFOA by
gavage at doses of 0, 1, 3, 10, and 30 mg/kg/day. At scheduled sacrifice, after completion of the
cohabitation period in FO male rats and on lactation day (LD) 22 in FO female rats, blood
samples (3/gender/group-control; 10/gender/group-treated) were collected from animals dosed
with 0, 10, and 30 mg/kg for analysis of PFOA. Serum analysis for the FO 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 |ig/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 |ig/ml, respectively, for the 10- and 30-mg/kg/day
dose groups. In the FO female controls, serum PFOA was below LOQ (0.00528 jig/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 |ig/ml, respectively, for the 10- and
30-mg/kg/day dose groups.
PFOA levels during gestation and lactation were studied by Hinderliter et al. (2005) and
Mylchreest (2003). Groups of 20 pregnant Sprague-Dawley rats were dosed with 0, 3, 10, and
30 mg/kg/day of PFOA during days 4-10, 4-15, and 4-21 of gestation, or from gestation day
(GD) 4 to LD 21. Maternal blood samples were collected at 2 hours ± 30 minutes (mins) 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 GD 10, only embryos were recovered, and on GDs 15 and 21, the
placentas, amniotic fluid, and embryos/fetuses were collected.
The remaining five rats per group were allowed to deliver their pups. On LDs 0, 3, 7, 14, and
21, the pups were counted, weighed (genders separate), and examined for abnormal appearance
and behavior. Randomly selected pups were sacrificed and blood samples were collected. On
LDs 3, 7, 14, and 21, the dams were anesthetized and milk and blood samples were collected;
dams were removed from their litters 1-2 hours prior to collection.
Plasma, milk, amniotic fluid extract, and tissue homogenate (placenta, embryo, and fetus)
supernatants were analyzed for PFOA concentrations by HPLC/MS. Maternal PFOA plasma
levels during gestation and lactation are presented in Table 2-11. Maternal plasma levels at
2 hours post-dosing (approximately the time of peak blood levels following a gavage dose) were
fairly similar during the course of the study with a mean level of 11.2, 26.8, and 66.6 |ig/ml in
the 3-, 10-, and 30-mg/kg/day groups, respectively; PFOA levels in the control group were below
the LOQ (0.05 |ig/ml).
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Table 2-11. Maternal Plasma PFOA Levels (ug/ml) in Rats During Gestation and Lactation
Exposure Period
GD 4 - GD 10
GD 4 - GD 15
GD4-GD21
GD 4 - LD 3
GD 4 - LD 7
GD 4 - LD 14
GD4-LD21
NA
Sample Time
GD 10 plasma
GD 15 plasma
GD 21 plasma
LD 3 plasma
LD 7 plasma
LD 14 plasma
LD 21 plasma
Average plasma
Dose
3 mg/kg/day
8.53 ± 1.06
15. 92 ±12.96
14.04 ±2.27
11.01±2.11
10.09 ±2.90
9.69 ±0.92
9.04 ±1.01
11. 19 ±2.76
10 mg/kg/day
23. 32 ±2.15
29.40 ±14.19
34.20 ±6.68
22.47 ± 2.74
25. 83 ±2.07
23.79 ±2,81
28.84 ±5. 15
26.84 ±4.21
30 mg/kg/day
70.49 ±8.94
79.55 ±3. 11
76.36 ±14.76
54.39 ±17.86
66.91 ±11.82
54.65 ±11.63
64.13 ± 1.45
66.64 ± 9.80
Source: Hinderliter et al. 2005; Mylchreest 2003
Notes: Mean ± SD; samples were from five dams/group/time point and were collected 2 hours post-dosing.
PFOA levels in the placenta, amniotic fluid, and embryo/fetus are presented in Table 2-12.
The levels of PFOA in the placenta on GD 21 were approximately twice the levels observed on
GD 15, and the levels of PFOA in the amniotic fluid were approximately four times higher on
GD 21 than on GD 15. The concentration of PFOA in the embryo/fetus was highest in the GD 10
embryo and lowest in the GD 15 embryo; PFOA levels in the GD 21 fetus were intermediate.
Table 2-12. Placenta, Amniotic Fluid, and Embryo/Fetus PFOA Concentrations
in Rats (ug/ml)
Exposure Period
GD 4-GD 10
GD 4-GD 15
GD 4-GD 21
Tissue
GD 10 embryo
GD 15 placenta
amniotic fluid
embryo
GD 21 placenta
amniotic fluid
fetus
Dose
3 mg/kg/day
1.40 ±0.30
2.22 ±1.79
0.60 ±0.69
0.24 ±0.19
3.55 ±0.57
1.50 ±0.32
1.27 ±0.26
10 mg/kg/day
3.33 ±0.81
5. 10 ±1.70
0.70 ±0.15
0.53 ±0.18
9.37 ±1.76
3.76 ±0.81
2.61 ±0.37
30 mg/kg/day
12.49 ±3.50
13.22 ±1.03
1.70 ±0.91
1.24 ±0.22
24.37 ±4.13
8.13 ±0.86
8.77 ±2.36
Source: Hinderliter et al. 2005; Mylchreest 2003
Note: Mean ± SD; samples were pooled by litter and were collected 2 hours post-dosing.
The concentrations of PFOA in the plasma of the GD 21 fetus were approximately half the
levels observed in the maternal plasma (Table 2-11). The values were about twice as high in the
dams as in the pups with mean values of 14.04, 34.20, and 76.36 |ig/ml, respectively, in the 3-,
10-, and 30-mg/kg/day groups for the dams and 5.88, 14.48, and 33.11 |ig/ml, respectively, for
the pups. Pup plasma levels decreased between birth and LD 7 (Table 2-13) and were, thereafter,
similar to the levels observed in the milk (Table 2-14). The pups were not separated by gender.
The concentration of PFOA in the milk also was fairly similar throughout lactation and was
approximately one-tenth of the PFOA levels in the maternal plasma (see Table 2-11); the mean
values for maternal milk were 1.1, 2.8, and 6.2 |ig/ml in the 3-, 10-, and 30-mg/kg/day groups,
respectively (Table 2-14).
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Table 2-13. Fetus/Pup PFOA Concentration (ug/ml) in Rats During
Gestation and Lactation
Exposure Period
GD4-GD21
GD 4-LD 3
GD 4-LD 7
GD 4-LD 14
GD 4-LD 21
Tissue
GD21 fetal plasma
LD 3 pup plasma
LD 7 pup plasma
LD 14 pup plasma
LD 21 pup plasma
Dose
3 mg/kg/day
5.88 ±0.69
2.89 ±0.70
0.65 ±0.20
0.77 ±0.10
1.28 ±0.72
10 mg/kg/day
14.48 ±1.51
5.94 ±1.44
2.77 ±0.58
2.22 ±0.38
3.25 ±0.52
30 mg/kg/day
33. 11 ±4.64
11.96 ±1.66
4.92 ±1.28
4.91 ±1.12
7.36 ±2.17
Source: Hinderliter et al. 2005; Mylchreest 2003
Note: Mean ± SD; samples were pooled by litter and were collected 2 hours post-dosing.
Table 2-14. PFOA Levels (ug/ml) in Rats Maternal Milk During Lactation
Exposure Period
GD 4-LD 3
GD 4-LD 7
GD 4-LD 14
GD 4-LD 21
NA
Sample Time
LD 3-milk
LD 7-milk
LD 14-milk
LD 21-milk
Average milk
Dose
3 mg/kg/day
1.07 ±0.26
0.94 ±0.22
1.15 ±0.06
1.13 ±0.08
1.07 ±0.09
10 mg/kg/day
2.03 ±0.33
2.74 ±0.91
3.45 ±1.18
3.07 ±0.51
2.82 ±0.60
30 mg/kg/day
4.97 ±1.20
5.76 ±1.26
6.45 ±1.38
7.48 ±1.63
6.16 ±1.06
Source: Hinderliter et al. 2005; Mylchreest 2003
Notes: Mean ± SD; samples were from 5 dams/group/time point and were collected 2 hours post-dosing.
Mouse. Fenton et al. (2009) orally dosed pregnant CD-I mice (n = 25/group) with 0, 0.1, 1, and
5 mg PFOA/kg on GD 17. On GD 18, five dams/group were sacrificed and trunk blood, urine,
amniotic fluid, and the fourth and fifth mammary glands were collected. One fetus/dam was
euthanized and retained for whole-pup analysis. The remaining dams were allowed to litter.
Biological samples as described above excluding amniotic fluid also were collected on postnatal
days (PNDs) 1, 4, 8, and 18. As before, at each time-point, a single pup was euthanized and
retained for whole-pup analysis. Blood from the remaining pups was collected and pooled. Milk
was collected from dams on PNDs 2, 8, 11, and 18 following a 2-hour separation of the pups
from the dam.
The concentration of PFOA in dam serum was approximately twice that detected in amniotic
fluid (Table 2-15). Compared to the amniotic fluid, the concentration of PFOA in the fetuses was
increased by 2.3-, 3.1-, and 2.7-fold at 0.1, 1, and 5 mg/kg, respectively. The highest
concentration of PFOA was detected in the serum of nursing dams. In the dams, the
concentration of PFOA in the serum exhibited a U-shaped response curve; the lowest serum
concentration was observed at the time of peak lactation. Dam mammary tissue and milk PFOA
concentrations showed a U-shaped response that mirrored that found in the dam's serum. The
concentration of PFOA in pup's serum was significantly higher than PFOA concentration in
dam's serum and appeared to decrease as the time for weaning approached. When pup PFOA
concentration was calculated with consideration for pup body weight gain, PFOA body burden
increased through the peak of lactation and began to decrease by PND 18, showing an inverse
U-shaped response curve.
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Table 2-15. PFOA Levels (ng/ml) in Mice During Gestation and Lactation in Selected
Fluids and Tissues
Tissue
Dam Serum
Amniotic Fluid
Dam Urine
Mammary Gland
Milk
Whole Pup
Pup Serum
Day
GD 18
PND 1
PND4
PND 8
PND 18
GD 18
GD 18
PND 1
PND 4
PND 8
PND 18
GD 18
PND 1
PND 4
PND 8
PND 18
PND 2
PND 8
PND 11
PND 18
GD 18
PND 1
PND 4
PND 8
PND 18
PND 1
PND 4
PND 8
PND 18
Dose
0.1 mg/kg
143 ±19
217.5 ±35
110.0 ±12
46.7 ±21
123.3 ±41
99.0 ±28
21. 9 ±8.6
7.7 ±1.7
8.4 ±6.4
0.8 ±0.22
1.8 ± 1.1
18.9 ±1.9
27.4 ±6.8
9.6± 8.4
2.4 ±3.8
17.1 ±10
32.5 ±12
11.6±8.1
5.4 ± 1.0
43. 5 ±19
136.3 ±15
150.9 ±21
91. 8 ±8.9
60.9 ±16
17.5 ±11
324.7 ±36
267.6 ± 47
260.2 ± 56
111.8±30
1 mg/kg
1697 ± 203
1957.0 ±84
1269.4 ±235
360.8 ±98
1035.2 ±305
865.3 ±191
104.9 ±69.7
116.8 ±64
53.5 ±15
11.6 ±6.2
18.7 ±8.6
307.2 ±30.4
343.8 ±53
239.2 ±53
71.7 ±22
239.9 ±76
716.7 ±145
77.4 ±19
42.3 ±9.1
251.8 ±147
1665. 8 ±213
1606.9 ±288
1183.2 ±187
729.0 ± 92
251.9 ±112
3926.8 ±480
3020.8 ± 223
2548.2 ± 245
1124.8 ±236
5 mg/kg
7897± 663
9845.6 ±1478
6776.6 ±561
1961.8 ±414
5156.5 ±1201
3203.8 ±492
666.7 ± 169
492.3 ±119
401. 5 ±117
40.1 ±17
91.7 ±49
1429± 186
1933.5 ±194
1461.8 ±267
411.8±78
1372.8 ±240
1236.6 ±1370
245.1 ±26
282.5 ± 162
909.8 ±308
6256.5 ±751
7134.5 ±1097
507 1.4 ±267
3 118.5 ±424
1391.5±118
16,286.4 ±1372
11,925.2 ±1077
9215.8 ±594
5894.3 ±743
Source: Fenton et al. 2009
Pregnant C57BL/6/Bkl mice were fed diets containing 0.3 mg PFOA/kg/day from GD 1
through the end of pregnancy. At birth, the PFOA concentrations in the offspring were
0.7 ± 0.1 [j,g/g in the brain and 16.3 ± 4.1 [j,g/g in the liver (Onishchenko et al. 2011).
Perfluorooctanoic acid (PFOA) - May 2016
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Macon et al. (2011) gavage-dosed CD-I mice with 0, 0.3, 1.0, or 3.0 mg PFOA/kg from GD
1 to GD 17 or with 0, 0.01,0.1, and 1.0 mg PFOA/kg from GD 10 to GD 17. In the full gestation
experiment (GD 1-17), offspring were sacrificed on PNDs 7, 14, 21, 28, 42, 63, and 84, and in
the half gestation experiment (GDs 10-17), female offspring were sacrificed on PNDs 1, 4, 7, 14,
and 21. Serum, liver, and brain from the offspring were analyzed for PFOA by UPLC/MS/MS.
At the lowest dose, PFOA concentration in the serum peaked at or before PND 7, but the two
higher doses peaked around PND 14 (Table 2-16). Calculated blood burdens which take into
account the increasing blood volumes and body weights for females showed an inverted
U-shaped curve peaking at PND 14 for all doses. In the liver, PFOA concentration decreased
over time with the highest concentration observed at PND 7. Lower concentrations of PFOA
were detected in the brain of the offspring on PND 7 and 14.
Table 2-16. Female Offspring PFOA Levels (ng/ml) in Mice After GD 1-17 Exposure
Tissue
Serum
Liver
Brain
Day
PND 7
PND 14
PND 21
PND 28
PND 42
PND 63
PND 84
PND 7
PND 14
PND 21
PND 28
PND 42
PND 63
PND 84
PND 7
PND 14
PND 21
PND 28
PND 42
PND 63
PND 84
Dose
0.3 mg/kg
4980 ±218
4535 ± 920
1194 ±394
630 ±162
377± 81
55 ± 17
16 ±5
2078 ± 90
972 ± 124
1188 ±182
678 ±130
342 ± 87
118 ±22
43 ± 12
150 ± 26
65 ±12
-------
Table 2-17. Female Offspring Serum PFOA Levels (ng/ml) in Mice
After GD 10-17 Exposure
Tissue
Serum
Blood Burden
(calculated)
Day
PND1
PND4
PND7
PND14
PND21
PND1
PND4
PND7
PND14
PND21
Dose
0.01 mg/kg
284.5 ±21.0
184.1 ±12.1
150.7 ±20.9
80.2 ±13. 9
16.5 ±2.1
15.2 ±1.7
20.6 ±0.1
27.3 ±3.8
27.0 ±4.6
7.9 ±1.0
0.1 mg/kg
2303.5 ±114.4
-
1277.8 ±122.6
645.4 ±114.2
131.7 ±24.5
114.3 ±5.4
-
221.7 ±24.9
218.5 ±39.8
66.4 ±12.8
1.0 mg/kg
16305.5 ±873. 5
-
11880.3 ±1447.6
6083.7 ±662.6
2025.1 ±281.9
926.0 ± 47.6
-
1965. 9 ±256.7
2033.6 ±293.5
984.7 ± 142.8
Source: Macon et al. 2011
Note: - = not measured, blood burden determined by (body weight x (58.5/1000) x serum x 0.55)
White et al. (2011) measured serum PFOA concentrations in three generation of CD-I mice
(Table 2-18). Pregnant mice (FO, n = 10-12 dams/group) were gavage-dosed with 0, 1, and 5 mg
PFOA/kg from GD 1-17. A separate group of pregnant mice (n = 7-10 dams/group) were
gavage-dosed with either 0 or 1 mg PFOA/kg from GD 1-17 and received drinking water
containing 5 parts per billion (ppb) PFOA beginning on GD 7 and continuing until the end of the
study for their offspring, except during breeding and early gestation, to simulate a chronic low-
dose exposure. An increase in serum PFOA concentration was observed in the control + 5 ppb
PFOA groups in the Fl and F2 generations and in the 1-mg/kg + 5-ppb PFOA group of the F2
generation. A decrease was observed for the remaining groups.
Table 2-18. Serum PFOA Levels (ng/ml) in Mice Over Three Generations
Dams at weaning
Offspring
Generation/
Day
FO/ PND 22
F1/-PND 91
Fl/PND 22
Fl/PND 42
Fl/PND 63
F2/PND 22
F2/PND 42
F2/PND 63
Dose
0 mg/kg + 5 ppb
74.8 ±11.3
86.9 ±14.5
21.3 ±2.1
48.9 ±4.7
66.2 ±4.1
26.6 ±2.4
57.4 ±2.9
68.5 ±9.4
1 mg/kg
6658.0 ±650.5
9.3 ±2.6
2443.8 ±256.4
609.5 ± 72.2
210.7 ±21. 9
4.6 ±1.2
0.4 ±0.0
1.1±0.5
1 mg/kg + 5 ppb
4772.0 ± 282.4
173.3 ±36.4
2743.8 ±129.7
558.0 ±55. 8
187.0 ±24.1
28.5 ±3.7
72.8 ±5.8
69.2 ±4.3
5 mg/kg
26980.0 ±1288.2
18.7 ±5.2
10045 ± 1125.6
1581.0 ±245.1
760.3 ± 188.3
7.8 ± 1.9
0.4 ±0.0
1.2 ±0.5
Source: White etal. 2011
Subcellular Distribution. Han et al. (2005) examined the subcellular distribution of PFOA in
the liver and kidney of male and female rats. Male and female Sprague-Dawley Crl:CD (SD)IGS
BR rats were gavage-dosed with 25 mg/kg [14C] PFOA and sacrificed 2 hours after dosing.
Blood was collected and the liver and kidneys were removed. Five subcellular fractions (nuclei
Perfluorooctanoic acid (PFOA) - May 2016
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and cell debris, lysosome and mitochondria, microsome, light microsome and ribosome, and
membrane-free cytosol) were obtained by differential centrifugation. The radioactivity per gram
(g) of each fraction and the total radioactivity were measured.
In the male liver, the highest proportion of total reactive residues (TRR) of PFOA was
located in the nuclei and cell debris (40%). The TRR for the other subcellular fractions were as
follows: membrane-free cytosol 26 percent% TRR, lysosome and mitochondria -14% TRR, and
microsome -16% TRR. The level of PFOA in the light microsome and ribosome was -1% TRR.
In the female liver, the highest proportion of PFOA was found in the membrane-free cytosol,
48% TRR. The TRR were nuclei and cell debris -31% TRR, lysosome and mitochondria
-12% TRR, and microsome -8% TRR. As observed in the males, the level of PFOA in the light
microsome and ribosome was -1% TRR (Han et al. 2005).
In the male kidney, the level of PFOA was 79% TRR in the membrane-free cytosol,
15% TRR in the nuclei and cell debris, and 4% TRR in the lysosome and
mitochondria/microsome/ light microsome and ribosome (combined). In the female kidney, the
level of PFOA was 71% TRR in the cytosol, 21% TRR in the nuclei and cell debris, and
8% TRR in the lysosome and mitochondria/ microsome/light microsome and ribosome
(combined). Further examination showed that in both genders, 98% of PFOA in the plasma was
protein bound. The protein-bound fraction of PFOA in the liver cytosol was 56% TRR. In the
kidney, the protein-bound fraction of PFOA in males was 42% TRR and 17% TRR in females
(Han et al. 2005).
Based on the results, the authors concluded that subcellular distribution of PFOA in the rat
liver was gender-dependent because the proportion of PFOA in the liver cytosol of female rats
was almost twice that of the male rats. They hypothesized that the female might have a greater
amount than the male of an unknown liver cytosolic binding protein with an affinity for
perfluorinated acids. They also hypothesized that the unknown protein or protein complex might
normally aid in transport of fatty acids from the liver. In the kidney, the subcellular distribution
did not show the gender difference seen with the liver; however, the protein-bound fraction for
the males (42%) was about twice that for the females (17%) (Han et al. 2005).
Inhalation Exposure
In a repeated exposure study, Hinderliter (2003) and Hinderliter et al. (2006b) exposed 6-8-
week-old male and female rats (5 per gender per group) to 0-, 1-, 10-, and 25-mg/m3 aerosol
concentrations of PFOA for 6 hours/day, 5 days/week for 3 weeks. Blood was collected
immediately before and after the daily exposure period 3 days/week. The aerosols had MMADs
of 1.3-1.9 jim with GSDs of 1.5-2.1. PFOA plasma concentrations were proportional to the
inhalation exposure concentrations, and repeated exposures produced little plasma carryover in
females, but significant day-to-day carryover in males. Male rats reached steady-state plasma
levels of 8, 21, and 36 |ig/ml for the 1-, 10-, and 25-mg/m3 groups, respectively, by 3 weeks. In
females, the post-exposure plasma levels were 1, 2, and 4 |ig/ml for the 1-, 10-, and 25-mg/m3
groups, respectively. When measured immediately before the next daily exposure, plasma levels
had returned to baseline in females, demonstrating CL within 24 hours of each daily dose.
Dermal Exposure
No data were identified on tissue distribution following dermal exposures.
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2.3 Metabolism
Several studies have examined metabolism of PFOA. However, no studies show clear
evidence of metabolism. Ophaug and Singer (1980) found no change in fluoride ion level in the
serum or urine following oral administration of PFOA to female Holtzman rats. Ylinen et al.
(1989) found no evidence of phase II metabolism of PFOA following a single intraperitoneal
PFOA dose (50 mg/kg) in male and female Wistar rats. The free anionic and possible conjugated
forms of PFOA in the urine were separated using BondElut tubes. The tubes contain NFh, which
is a weaker anion exchange sorbent and a good choice for retaining strong anions. The samples
were aspirated through the tube, washed with water, and eluted with sodium
bicarbonate/carbonate-buffer. The aspirate and eluate from the separation method were analyzed
by gas chromatography. PFOA was not detected in the aspirate, but was retained with the
cationic amino phase found in the eluate. This also occurred in control blanks spiked with PFOA.
The authors concluded that because the PFOA anion was completely bound to the weak cationic
amino phase in both the spiked controls and urine samples, PFOA in urine is not altered by phase
II metabolism (Ylinen et al. 1989).
2.4 Excretion
Excretion data are available for oral exposure in humans and laboratory animals. Several
studies have investigated the elimination of PFOA in humans, cynomolgus monkeys, and rats. In
human females, elimination pathways include pregnancy (cord blood) and lactation (breast milk)
(Apelberg et al. 2007; Tao et al. 2008; Thomsen et al. 2010; Volkel et al. 2008; von Ehrenstein et
al. 2009).
Elimination half-lives differ among species. There are also significant gender differences in
humans and some laboratory animal species. Information from humans does not, at this time,
provide sufficient data to determine the magnitude of interindividual and gender differences in
excretory half-lives. The transporters appear to play an important role in renal excretion of
PFOA and possibly its biliary elimination as well.
Humans. The urinary excretion of PFOA in humans is impacted by the isomeric composition of
the mixture present in blood and the gender/age of the individuals. The half-lives of the
branched-chain PFOA isomers are shorter than those for the linear molecule, an indication that
renal resorption is less likely with the branched chains.
Y. Zhang et al. (2013) determined half-lives for PFOA isomers based on paired serum
samples and early morning urine samples collected from healthy volunteers in two large Chinese
cities. Half-lives were determined using a one compartment model and an assumption of first
order CL. The Vd applied in the analysis as determined by Thompson et al. (2010) was
170 mL/kg. CL was estimated from the concentration in urine normalized for creatinine and
assuming excretion of 1.2 and 1.4 L/day of urine and 0.9 and 1.1 mg creatinine/day for males
and females, respectively. The mean half-life for the sum of all PFOA isomers in younger
females (n = 12) was 2.1 years (range 0.19-5.2 years) while that for all males and older females
(n = 31) was 2.6 (range 0.0059-14 years); the medians were 1.8 and 1.7 years, respectively. The
mean values for the four branched-chain isomers of PFOA were lower than the value for the
linear chain, suggesting that resorption transporters might favor uptake of the linear chain over
the branched-chain isomers. Older females and males have longer half-lives than young females,
suggesting the importance of monthly menstruation as a pathway for excretion (Y. Zhang et al.
2013).
Perfluorooctanoic acid (PFOA) - May 2016 2-23
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T. Zhang et al. (2014) derived estimates for PFOA's urinary excretion rate using paired urine
and blood samples from 54 adults (29 males and 25 females) in the general population and 27
pregnant females in Tainjin, China. The age range for the general population was 22-62 years
and for the pregnant females was 21-39 years. Urinary excretion was calculated based on the
concentration in the urine times volume of urine wherein a urinary volume of 1,200 mL/day was
applied to all females and 1,600 mL/day for all males. Urine samples were first-draw morning
samples. Total daily intakes for PFOA were calculated from the concentration in blood using
first order assumptions, a half-life of 2.3 years (Bartell et al. 2010) and a Vd of 170 mL/kg
(Lorber and Egeghy 2011; Thompson et al. 2010). PFOA was detected in the blood samples for
all participants but for only 76% of the urine samples from the general population and 30% for
the pregnant females. There was a direct correlation between the PFOA concentrations in blood
and creatinine adjusted urine (r = 0.348 p = 0.013) for the general population but not for the
pregnant females. When limited to the eight females who had detectable levels in both blood and
urine, there was a significant correlation (r= 0.724, p = 0.042).
Among the general population, the daily urinary excretion rate accounted for 25% of the
estimated intake with the excretion higher in males (31%) than in females (19%). The urine:
blood ratio was lower for pregnant females than for nonpregnant females (0.0011 versus 0.0029),
suggesting other removal pathways such as placenta and cord blood. There was little difference
between the younger menstruating females (21-50 years versus 51-61 years), but there is no
indication that data were collected from the participants relative to menstruation status on the day
of blood and urine collection.
Wong et al. (2014) looked at the role of menstrual blood as an excretory pathway to explain
the shorter half-life of PFOS in females than in males. They fit a population-based PK model to
six cross-sectional NHANES data sets (1999-2012) for males and females. They concluded that
menstruation could account for about 30% of the PFOS elimination half-life difference between
females and males. Although Wong et al. (2014) studied PFOS and not PFOA, their findings are
relevant to both chemicals.
Elimination of PFOA by way of the gastrointestinal tract was reported in a case history of a
single human male with high serum levels of perfluorinated chemicals that appeared to originate
from household dust following the installation of new carpeting (Genuis et al. 2010). Treatment
with cholestyramine, a bile acid sequestrant for 20 weeks (4g/day, three times a day), lowered his
serum PFOA concentration from 5.9 ng/g serum to 4.1 ng/g serum. More dramatic decreases
were observed with serum PFOS (23-14.4 ng/g serum) and PFHxS (58-46.8 ng/g serum), which
were present at higher levels in the serum. This observation suggests that excretion with bile and
possible enterohepatic resorption via intestinal transporters limits the loss of absorbed PFOA via
feces in the absence of a binding agent such as cholestyramine.
2.5 Animal Studies
Oral Exposure
Monkey. Butenhoff et al. (2004b) studied the fate of PFOA in cynomolgus monkeys in a
6-month oral exposure study. Groups of four to six male monkeys each were administered PFOA
daily via oral capsule at DRs of 0, 3, 10, and 30/20 mg/kg for 6 months. Two monkeys exposed
to 10 mg/kg and three monkeys exposed to 20 mg/kg were monitored for 21 weeks (recovery
period) following dosing. Urine and fecal samples were collected at 2-week intervals and were
analyzed for PFOA concentrations.
Perfluorooctanoic acid (PFOA) - May 2016 2-24
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Urine PFOA concentrations over the duration of the study were 53 ± 25, 166 ± 83, and
181 ± 100 ng/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 |ig/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. Flinderliter (2004) and Hinderliter et al. (2006a) investigated the
relationship between age and urine PFOA concentrations in male and female Sprague-Dawley
rats. Immature rats 3, 4, or 5 weeks of age were administered PFOA via oral gavage as a single
dose of 10 or 30 mg/kg. Two hours after dosing, five rats per gender per age group and dose
group were sacrificed and blood samples were collected (see section 2.2.1). The remaining five
rats per gender per age and dose group were placed in metabolism cages for 24-hour urine
collection. Urinary output (volume) was not quantified or standardized for creatinine levels.
Urine PFOA concentrations differed significantly with age, dose, and gender (p<0.01, Table
2-19). Urinary excretion of PFOA was substantially higher in females than in males, and the
female urine PFOA concentrations increased with age. In male rats, 24-hour urine PFOA
concentrations decreased with age up to five weeks. In both genders, urine PFOA was higher
(2.5 to 6.5 times) at the 30-mg/kg dose as compared to the 10-mg/kg dose.
There was a difference in urinary excretion between the 3-week-old and 4/5-week-old male
rats, with the older rats excreting -50% less PFOA in the urine than the younger rats at 10 mg/kg
and 30 mg/kg. If the data from urine are integrated with the plasma data in the same study
(Table 2-9), the male plasma levels increased from the 3-week value and were relatively stable
for weeks 4 and 5. In the females, urine excretion increased gradually with age (Table 2-19) and
plasma concentrations decreased (Table 2-10).
Table 2-19. Urine PFOA Concentrations in Male and Female Rats
Age
(weeks)
o
J
4
5
o
J
4
5
Dose
(mg/kg)
10
10
10
30
30
30
Urine PFOA (|o.g/ml at 24 hours post-dose)
Male
Mean
9.57
4.53
4.03
51.76
28.70
15.65
SD
4.86
2.45
2.36
28.86
18.84
6.24
Female
Mean
21.17
23.26
49.77
94.89
104.12
123.16
SD
8.95
15.27
24.64
26.26
28.97
51.56
Source: Hinderliter 2004
Hundley et al. (2006) examined excretion of PFOA in one male and one female CD rat
(sexually mature). Each was given a single dose of 10 mg/kg 14C-PFOA and housed in a
metabolism cage. Urine and feces were collected at 12, 24, 48, 72, 96, and 120 hours post-dose.
The female rat excreted more PFOA over the 120-hour collection period than the male rat. In the
male rat, 25.6% and 9.2% 14C-PFOA were excreted in the urine and feces, respectively. In the
female rat, 73.9% and 27.8% 14C-PFOA were excreted in the urine and feces, respectively. The
Perfluorooctanoic acid (PFOA) - May 2016
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female rat excreted almost all of the PFOA by 48 hours compared with only 19% of the dose
excreted by the male rat over the same amount of time. The cumulative percent of the dose
excreted is shown in Table 2-20.
Table 2-20. Cumulative Percent 14C-PFOA Excreted in Urine and Feces by Rats
Rat
Male
Female
Hours After Dosing
12
0.6
52.5
24
8.7
96.4
48
19.2
99.8
72
23.4
100.0
96
30.2
100.0
120
34.3
100.0
Source: Hundley et al. 2006
Adult male Sprague-Dawley rats (n = 7) were given a single gavage dose of 0.5 mg PFOA/kg
and monitored for 38 days (Benskin et al 2009). Over the course of the study, the rats were held
in metabolic cages and urine and feces were collected. The mean blood PFOA concentration was
1.1 |ig/mL 24 hours post-dose. During the first 24 hours post-dose, 65% of PFOA was excreted
in the urine; most of the PFOA that was not absorbed was excreted in the feces. After that time
period, 91-95% of the daily excreted PFOA was eliminated in the urine. On day 3, the mean
PFOA concentration in urine and feces were 265 ng/g and 28 ng/g. The half-life for elimination
from plasma in male rats was 13.4 days.
Cui et al. (2010) exposed 2-month-old male Sprague-Dawley rats (10 per group) to PFOA
(96% active ingredient) at 0, 5, and 20 mg/kg/day once daily by gavage for 28 days. Urine and
fecal samples were collected through use of metabolism cages at 24-hour intervals immediately
following dosing on days 1, 2, 5, 7, 10, 14, 18, 21, 24, and 28 of the study. Daily urine volume
and fecal weight were comparable across all groups throughout the study. As measured by
excretion 24-hours after the first dose, 17.9% of the applied dose was excreted in the urine of the
low-dose group and 22% for the high-dose group. The percent of the absorbed dose was 92.8%
and 92.3% for the low and high doses, respectively, when the fecal excretion over the 24 hours
following dosing was estimated to be unabsorbed material. During week 1, a sharp increase in
urinary and fecal excretion expressed as percent of dose/day was observed in rats of both groups.
The excretion rate leveled off at about 50% for the low-dose animals for the remainder of the
28 days. In the case of the high-dose animals, the urinary excretion remained level at about 80%
for the second and third weeks and then increased sharply to about 140% at 28 days. The fecal
excretion rates were 7.2% and 7.7% for rats in the 5- and 20-mg/kg groups, respectively, during
the first 24 hours post-dosing and continued an upward trend throughout the 28 days with the
terminal percent/day about 25% for the low-dose group and 40% for the high-dose group.
Dose is an important variable that impacts excretion. Rigden et al. (2015) exposed groups of
five male Sprague-Dawley rats to doses of 0, 10, 33, and 100 mg/kg/day for 3 days and
maintained them for 3 additional days; overnight urine was collected and body weight was
measured daily. Of greatest interest relative to the limitations on renal resorption, is the dose-
related increase in urine PFOA concentration and urine PFOA concentration per mg creatinine
for the 33- and 100-mg/kg/day groups compared to the 10-mg/kg/day group. The peak in PFOA
excretion normalized to creatinine occurred on day 3 after the cessation of dosing. The
concentration at 33 mg/kg/day was 500 times greater than that at 10 mg/kg/day. At the
100-mg/kg/day dose, the peak concentration was about 3,200 times greater than for the low dose.
The low-dose excretion was only slightly greater than the controls. The urine results support the
renal resorption hypothesis concept and suggest that there is a threshold limit on resorption that,
Perfluorooctanoic acid (PFOA) - May 2016
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once exceeded, dramatically increases PFOA loss in urine. As a consequence, half-life for
continuous low-dose exposures will be longer than for single or short-term high-dose exposures.
Other Species. Hundley et al. (2006) examined excretion of PFOA in CD mice, BIO-15.16
hamsters, and New Zealand White rabbits. One male and one female of each species was given a
single dose of 10-mg/kg 14C-PFOA and housed in metabolism cages. Urine and feces were
collected at 12, 24, 48, 72, 96, and 120 hours post-dose. Additional samples were collected from
rabbits at 144 and 168 hours post-dose.
Over 120 hours, the male mouse excreted 3.4% 14C-PFOA in urine and 8.3% 14C-PFOA in
feces, and the female mouse excreted 6.7% 14C-PFOA in urine and 5.7% 14C-PFOA in feces. The
mice were similar in the amounts excreted. The male hamster excreted 90.3% and 8.2% 14C-
PFOA in urine and feces, respectively, and the female hamster excreted 45.3% and 9.3%14C-
PFOA. The male hamster excreted a greater amount of 14C-PFOA than the female hamster. Over
84% of 14C-PFOA was excreted 24 hours after dosing by the male hamster compared to less than
25% of 14C-PFOA excreted by the female hamster at 24 hours after dosing. Over 168 hours, the
male rabbit excreted 76.8% and 4.2% 14C-PFOA in urine and feces, respectively, and the female
rabbit excreted 87.9% and 4.6% 14C-PFOA. Both rabbits excreted most of the dose by 24 hours.
The cumulative percentage of 14C-PFOA excreted is shown in Table 2-21.
Table 2-21. Cumulative Percent 14C-PFOA Excreted in Urine and Feces
Species
Mouse
Hamster
Rabbit
Gender
Male
Female
Male
Female
Male
Female
Hours After Dosing
12
0.4
0.2
67.3
11.3
77.8
86.7
24
4.1
4.1
84.5
24.6
80.2
90.5
48
6.7
6.5
96.1
36.4
80.4
92.0
72
8.6
8.4
97.4
43.9
80.4
92.2
96
9.1
9.0
98.2
50.1
80.4
92.7
120
10.8
11.0
98.4
54.0
80.4
92.9
168
-
-
-
-
80.4
93.0
Source: Hundley et al. 2006
When the data in Table 2-21 are integrated with the data from rats, the gender differences in
PFOA excretion rate appear to be species-specific. Female rats, male hamsters, and both genders
of rabbits appear to be good excreters based on their response to a radiolabeled dose of
10 mg/kg. Most of the dosed material is excreted within 24 hours after dosing. Female hamsters
apparently are moderate excreters. Males and female mice excreted only about 10% of the dose
over the 120 hours (5 days) after dosing. Mice do not show a gender difference but retain more
of the dose than do hamsters, rabbits, and female rats. The long half-lives in humans suggest that
their excretion rates are more like mice or male rats.
Inhalation Exposure
Although no data were identified on urine or fecal excretion of PFOA following inhalation
exposures, the Hinderliter study (2003) provides evidence of CL following single and repeated
inhalation exposures in Sprague-Dawley rats. Plasma PFOA concentrations following a single
exposure to 1, 10, and 25 mg/m3 PFOA declined 1 hour after exposure in females and 6 hours
after exposure in males. In females, the elimination of PFOA was rapid at all exposure levels
and, by 12 hours after exposure, their plasma levels had dropped below the analytical LOQ
(0.1 |ig/ml). In males, the plasma elimination was much slower and, at 24 hours after exposure,
the plasma concentrations were approximately 90% of the peak concentrations at all exposure
Perfluorooctanoic acid (PFOA) - May 2016
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levels. In the repeated exposure study, male and female rats were exposed to the same
concentrations for 6 hours/day, 5 days/week for 3 weeks. Steady-state plasma levels were
reached in males by 3 weeks, but plasma PFOA levels in females returned to baseline with
24 hours of each dose. The data are illustrative of distinct toxicokinetic differences between male
and female rats in their response to PFOA exposure (Hinderliter 2003).
Dermal Exposure
No data were identified on excretion following dermal exposures. Minimal fecal excretion is
anticipated for the dermal route of exposure although the biliary pathway can be a route for
excretion of material absorbed through the skin, distributed to the liver, and discharged to the
gastrointestinal tract.
2.5.1 Mechanistic Studies of Renal Excretion
Several studies have been conducted to elucidate the cause of the gender difference in the
elimination of PFOA by rats. Many of the studies have focused on the role of transporters in the
kidney tubules. Most studies have examined the OATs located in the proximal portion of the
descending tubule. OATs are found in other tissues as well and were discussed earlier for their
role in absorption and distribution. In the kidney, they are responsible for delivery of organic
anions, including a large number of medications from the serum into the kidney tubule for
excretion as well as reabsorption of anions from the glomerular filtrate. The transporters are
particularly important in excretion of PFOA because it binds to surfaces of serum proteins
(particularly albumin), which makes much of it unavailable for removal during glomerular
filtration. Other transporter families believed to be involved in renal excretion are the OATPs
and the MRPs. However, they have not been evaluated as extensively as the OATs for their role
in renal excretion.
OATs are located on both the basolateral (serum interface) and apical surfaces of the brush
boarder of the proximal tubule inner surface. At the basolateral surface, the OATs transport the
perfluorooctanoate anion from the serum to the tubular cells (Anzai et al. 2006; Cheng and
Klaassen 2008; Klaassen and Aleksunes 2010; Klaassen and Lu 2008; Nakagawa et al. 2007,
2009). OAT1, 2, and 3 are located on the basolateral membrane surface. OAT4 and OATS are
located on the apical surface of the tubular cells, where they reabsorb the PFOA anions from the
glomerular filtrate. Figure 2.2 diagrams the flow of organic anions such as the PFOA anion from
serum to the glomerular filtrate for excretion and resorption of organic acids from the glomerular
filtrate with transport back to serum. OATs can function for uptake into the cell across both the
basolateral and apical surfaces.
Several MRP transporters also appear to function in the kidney and move organic anions in
and out of cells at both the basolateral surface (e.g., MRP2/4) and the apical surface (e.g., MRP1)
as well as one or more OATPs on each surface (Cheng and Klaassen 2009; Klaassen and
Aleksunes 2010; Klaassen and Lu 2008; Kusuhara and Sugiyama 2009; Launay-Vacher et al.
2006; Yang et al. 2009). Bidirectional movement of PFOA across both the basolateral and apical
surfaces is driven by concentration gradients and/or active transport. Far more data exist on
PFOA and OATs in the kidneys than on OATPs and MRPs. Abbreviations for individual
transporters on the basolateral and apical surfaces differ across publications. The accepted
convention is to use uppercase letters to refer to human transporters and lowercase letters to refer
to animal transporters. For this report, the data are not reported by species but by transporter
family and the uppercase letters are used.
Perfluorooctanoic acid (PFOA) - May 2016 2-28
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Proximal Tubule Cells
Excretion Reabsorption
Asbt
\
Apical Side
Source: Klaassen and Aleksunes 2010
\
Basolateral
Figure 2-2. Localization of Transport Proteins
Knowledge about specific OAT, OATP, and MRP transporters in the kidneys is rapidly
evolving. A low membrane density or blockage of basolateral OATs will decrease PFOA
excretion while low membrane densities or blockage of apical OATs will increase excretion
because they decrease resorption of anions from the glomerular filtrate.
The earliest studies of the impact of gender on urinary excretion were conducted by
Hanhijarvi et al. (1982) using probenecid, an inhibitor of renal excretion of organic acids on
PFOA excretion in male and female Holtzman rats. The female rats that had not received the
probenecid excreted 76% of the administered dose of PFOA over a 7-hour period, while males
excreted only 7.8% of the administered dose over the same period of time. Probenecid
administration modified the cumulative excretion curve for males only slightly. In females,
however, probenecid markedly reduced PFOA elimination to 11.8%. The authors concluded that
the female rat possesses an active secretory mechanism that rapidly eliminates PFOA from the
body that male rats do not possess.
Kudo et al. (2002) examined the role of sex hormones and OATs on the renal clearance
(CLR) of PFOA. Renal mRNA levels of specific OATs in castrated male and ovariectomized
(OVX) female Wistar rats also were determined. Castration of male rats caused a 14-fold
increase in CLR of PFOA. The elevated PFOA CLR in castrated males was reduced by treating
them with testosterone. Treatment of male rats with estradiol increased the CLR of PFOA. In
female rats, ovariectomy caused a significant increase in CLR of PFOA (a twofold increase), but
the administration of estradiol to OVX female rats returned CLR of PFOA to normal values.
Treatments of female rats with testosterone reduced the CLR of PFOA.
Treatment with probenecid, a known inhibitor of OAT1-6 and OATS, markedly reduced the
CLR of PFOA in male rats, castrated male rats, and female rats (Kudo et al. 2002). Accordingly,
the male sex hormones appear to decrease the presence of OATs in the renal basolateral
membranes while the female sex hormones appear to increase the transporters.
Perfluorooctanoic acid (PFOA) - May 2016
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To identify the transporter molecules responsible for PFOA transport in the rat kidney, renal
mRNA levels of specific OATs were determined in male and female rats under various hormonal
states and compared with the CLR of PFOA. The level of OAT2 mRNA in male rats was only
13% of the level in female rats. Castration or estradiol treatment increased the level of OAT2
mRNA whereas treatment of castrated male rats with testosterone reduced it. Ovariectomy of
female rats significantly increased the level of OATS mRNA. Multiple regression analysis of the
data suggested that OAT2 and OATS are responsible for urinary elimination of PFOA in the rat;
however, the possibility of a resorption process mediated by OATP1 was mentioned as a
possible factor in male rat retention of PFOA. OAT2 and OATS are located on the basolateral
cell surface. OATP1 is located on the apical surface of the renal tubule cells (Kudo et al. 2002).
Cheng et al. (2006) examined whether sex hormones influenced gender-specific OATP
expression in the kidneys of adult male and female C57BL/6 mice. Gonadectomized mice were
used for the studies in conjunction with hormone replacement measures (5a-dihydroxy-
testosterone [DHT] or 17-P estradiol [E2]). OATPlal and OATPSal were evaluated. Treatment
with DHT resulted in significant increase in both OATPs in the kidneys of male and female
gonadectomized mice. In both cases, the change in males was greater than the change in females.
Treatment with E2 almost abolished the expression of OATPlal in the kidneys but caused no
significant change in OATPSal. In the intact control animals, almost no expression of OATPlal
occurred in the kidneys of females and a significantly lower expression of OATPSal (p<0.05)
occurred. In the gonadectomized control animals, little or no expression of OATPlal occurred in
either gender, and expression of OATPSal was equivalent in both genders.
Nakagawa et al. (2007) investigated the role of OATs in the renal excretion of PFOA using
in vitro methods. HEK293-transformed cells, derived from human embryonic kidney (HEK),
were transfected with human or rat OAT1, OAT2, or OATS constructs. Cells from the S2
segment of the proximal tubule were transfected with human or rat OAT2 constructs. HEK293
and S2 cells transfected with the vector served only as control cells. The transfected HEK293
cells were incubated for 1 min with or without 0, 10, and 100 (imol [14C]PFOA and/or varying
concentrations of favored OAT substrates to determine inhibitory effects of PFOA as follows:
5 jimol [14C]para-aminohippuric acid (OAT1), 20 nmol [14C]estrone sulfate (OATS), and
10 nmol [14C]prostaglandin F2« (OAT2).
PFOA significantly inhibited para-aminohippuric acid and estrone sulfate uptake mediated by
OAT1 and OATS, respectively. At 10 jimol PFOA, uptake of 5 jimol [14C] para-aminohippuric
acid was 75-85% of the control level and, at 100 jimol PFOA, uptake was reduced to 35-45% of
control. Estrone sulfate uptake by human OATS was 65% of the control level at 10 jimol PFOA
and 40% of control at 100 |imol PFOA. Estrone sulfate uptake by rat OATS was 15% of the
control level in the presence of 10 jimol PFOA and was almost completely inhibited at 100 jimol
PFOA. Prostaglandin F2a uptake by OAT2 was inhibited moderately by PFOA, 75-85% of
control at 10 |imol PFOA, and 65% of control at 100 |imol PFOA.
In the second part of their study, Nakagawa et al. (2007) incubated HEK293 and S2
transfected cells with 10 jimol [14C]PFOA for 1 min to determine uptake. Time-dependent
uptake of 5 jimol [14C]PFOA from 0 to 30 mins was conducted in the HEK293 cells transfected
with human or rat OAT1, OAT2, or OATS. Experiments were conducted in triplicate. Uptake of
PFOA was stimulated (p<0.001) in cells transfected with human or rat OAT1 or OATS, 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 OATS increased linearly up to 2 mins and reached
a plateau in about 15 mins. Kinetic evaluations resulted in substrate concentration at which the
Perfluorooctanoic acid (PFOA) - May 2016 2-30
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initial reaction rate is half maximal (Km) values of 48.0, 51.0, 49.1, and 80.2 jimol for human
OAT1, rat OAT1, human OATS, and rat OATS, respectively. The authors showed that both
human and rat OAT1 and OATS transport PFOA in the kidney while human and rat OAT2 do
not (Nakagawa et al. 2007).
Yang et al. (2009) investigated the role of OAT polypeptide lal (OATPlal) in the renal
elimination of PFOA. The polypeptide is located on the apical side of proximal tubule cells and
could be the mechanism for renal reabsorption of PFOA in rats. The level of mRNA of
OATPlal in male rat kidney is 5-20-fold higher than in female rat kidney, OATPlal protein
expression is higher in male rat kidneys, and it is regulated by sex hormones. One of its known
substrates is estrone-3-sulfate (ESS). A substantial presence of OATPlal in male rats would
favor resorption of PFOA in the glomerular filtrate and reduce excretion.
Chinese hamster ovary (CHO) cells were transfected with rat OATPlal complementary
DNA. The transfected CHO cells were incubated with 4 |imol [14C]PFOA for up to 10 mins or
with 0-1,000 jimol [14C]PFOA for 2 mins to determine uptake. The difference between the
uptake velocities of CHO OATPlal-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 jimol [14C]PFOA for 2 mins in the absence or presence of inhibitors (e.g., BSP,
taurocholate, probenecid, />-aminohippurate, and naringin [a flavonoid found in grapefruit]) for
inhibition studies. The transfected CHO cells were incubated with 2 |imol ESS and 0, 0.1, or
1 mM perfluorocarboxylates with carbon chain lengths ranging from 4 to 12, including PFOA
(C8) for 30 seconds for ESS inhibition studies.
In time-dependent uptake experiments, uptake of PFOA by OATPlal-transfected cells
increased proportionally to time during the first 2 mins of incubation. Vector-transfected cells
had a significant level of uptake of PFOA attributed to nonspecific passive diffusion. In the
concentration-dependent uptake experiments, uptake velocity of PFOA in OATPlal-transfected
cells increased with increasing concentration and saturation levels were not reached. In vector-
transfected cells, uptake velocities increased linearly with increasing concentration of PFOA,
demonstrating a passive diffusion mechanism. Active PFOA uptakethe difference between the
uptake of the OATPlal cells and the vector-transfected cellscould be described by the
Michaelis-Menton equation and exhibited saturable kinetics.
Inhibition experiments with substrates of OATs and OATPs showed that BSP, taurocholate,
and naringin inhibited PFOA uptake to 10-30% of control and^-aminohippurate inhibited PFOA
uptake to 62% of control. Probenecid, an OAT inhibitor, did not inhibit PFOA uptake at all. In
OATPlal-transfected cells, uptake of ESS was inhibited to less than 10% of control uptake
following incubation with 1 mM [14C]PFOA. Inhibition of ESS was less than 50% of control
uptake after incubation with 0.1 mM [14C]PFOA. Based on the results of the uptake and
inhibition experiments, the authors suggested that passive diffusion could be an important route
of PFOA distribution and that renal reabsorption in the male rat could be mediated by OATPlal.
Nakagawa et al. (2009) investigated the role that the human organic acid transporter (OAT4)
plays in transporting PFOA. Human OAT4 is located on the apical side of proximal tubule cells
and mediates reabsorption of organic anions. Transformed cells derived from HEK cells,
HEK293, were transfected with human OAT1, OATS, or OAT4 constructs. HEK293 cells
transfected with only the vector served as control cells. The transfected HEK293 cells were
incubated with 10 |imol [14C]PFOA for 15 mins to determine uptake. Transfected cells also were
incubated with 10 |imol [14C]PFOA for 15 mins and then washed with incubation medium
Perfluorooctanoic acid (PFOA) - May 2016 2-31
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containing 1%, 3%, and 5% BSA to investigate the contribution of nonspecific binding of PFOA
on the cell membrane. Experiments were conducted in triplicate.
Uptake of PFOA was significantly stimulated (p<0.01) in cells transfected with human
OAT1, OATS, and OAT4. Uptake of PFOA in human OAT1 transfected cells was 1.6-fold
higher than in control cells. In human OATS transfected cells, PFOA uptake was -2.4-fold
higher than in control cells. In human OAT4 transfected cells, PFOA uptake was 2.7-fold higher
than in control cells. Accumulation of PFOA in transfected human OAT4 cells also was
significantly greater than in human OAT1 cells (p<0.01). Washing the cells with BSA reduced
PFOA uptake by 30% at most, suggesting mediation by the transporters into the transfected cells.
The experiments showed that human OAT4 transports PFOA and that human OAT4 activity
might play a role in reabsorption of PFOA from the tubule, resulting in poor urinary excretion.
Yang et al. (2010) examined cellular uptake of PFOA by OATP1A2, OAT4, and urate
transporter 1 (URAT1) to determine their roles in mediating human renal reabsorption. CHO and
HEK293 cells were transfected with OATP1A2, OAT4, and URAT1 plasmid DNA or vector
DNA (control). In uptake studies, PFOA incubation times were 10 seconds (OAT4) and
30 seconds (URAT1). Cells transfected with OAT4 were incubated with 5 umol PFOA for up to
1 min in time-dependent uptake experiments. In inhibition studies, cells transfected with OAT4
were incubated with 5 umol [14C]PFOA for 10 seconds in the presence and absence of 100 umol
sulfobromophthalein (BSP), probenecid, glutarate, or polycyclic aromatic hydrocarbon (PAH).
Perfluorinated carboxylates with differing chain lengths (C4-C12) were used in chain length-
dependent inhibition experiments. Incubations with 3H-E3S (OAT4 and OATP1A2) or 6 umol
C14-uric acid (URAT1) in the presence and absence of 100 umol perfluorinated carboxylate
lasted 10 seconds (OAT4), 30 seconds (OATP1A2), and 1 min (URAT1).
PFOA uptake in OATP 1A2-transfected HEK293 cells was no different than uptake in control
cells. At 100 umol, E3S uptake was inhibited -30% by PFOA (C8), -62% by C9, -70% by CIO,
-42% by Cl 1, and -18% by C12. E3S uptake was not inhibited by C4-C7. In CHO cells
transfected with OAT4, time-dependent uptake experiments showed a saturation phase after an
incubation time of approximately 10 seconds. A pH-dependent increase in PFOA uptake was
observed with approximately 90% uptake at pH 8 and 250% at pH 5.5.
In concentration-dependent uptake experiments, uptake increased with increasing PFOA
concentration (0-1000 umol) 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 CIO (-80%). Inhibition in the presence of Cl 1 and C12 were
-52% and -30%, respectively. Inhibition of E3S in the presence of C4, C5, and C6 was less than
20% for each.
PFOA uptake in HEK293 cells transfected with URAT1 was not statistically different from
control cells in the presence and absence of Cl". Under both conditions, PFOA intake was
enhanced especially in the absence of Cl" in which PFOA uptake was greater than fourfold
compared to uptake in control cells. Time-dependent PFOA (5 umol) uptake by URAT1
increased with time during the 5-min incubation period, and a concentration-dependent increase
in PFOA uptake was observed (0-700 umol). 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 Cl 1, -50% in the presence of C5, -30% in the presence of C12, and -25% in the
presence of C4. Based on the results, Yang et al. (2010) concluded that PFOA was not a
Perfluorooctanoic acid (PFOA) - May 2016 2-32
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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, OATS, OATPlal, and URAT1. The transporters were transfected
into one of several cell lines and exposed to a series of perfluorinated carboxylates having chain
lengths ranging from 2 to 18 carbons (C). The activity of the perfluorinated carboxylate on the
transporters was quantified on the basis of its ability to inhibit the transport of a favored
radiolabeled substrate. The PFAS inhibition of the individual transporters varied with chain
length. The perfluorinated carboxylate with 6, 7, and 8 carbon chains caused a significant
decrease in OAT1 transport of tritiated p-aminohippurate, with the C7 acid having the strongest
effect. The perfluorinated carboxylates with 5 through 10 carbon chains caused a significant
decrease in transport of tritiated ESS by OATS, with C8 and C9 acids having the strongest effect.
The transport of tritiated estadiol-17p-glucuronide by OATPlal was significantly inhibited by
perfluorinated carboxylates with 6 through 11 carbon chains, with CIO acid having the strongest
effect. The perfluorinated carboxylate did not inhibit OAT2 or URAT1 transport of favored
substrates.
The kinetic response of the OAT1, OATS, and OATPlal transporters to increasing
concentrations of selected perfluorinated carboxylates also was evaluated by Weaver et al.
(2010). The change in transport velocity (ng/mg protein/min) with increasing concentrations of
the perfluorinated carboxylate exhibited a Michaelis-Menton-type response. The kinetic data
were analyzed to determine the Km and Vmax, and the results are summarized in Table 2-22
below.
Table 2-22. Kinetic Parameters of Perfluorinated Carboxylate Transport by OAT1, OATS,
and OATPlal
Transporter
OAT1
OATS
OATPlal
PFAS
C7
C8
C8
C9
C8
C9
CIO
Km (umol)
50.5 ±13. 9
43.2 ±15.5
65.7 ±12.1
174.5 ±32.4
126.4 ±23.9
20.5 ±6.8
28.5 ±5.6
Vmax (nmol/mg protein/min)
2.2 ±0.2
2.6 ±0.3
3. 8 ±0.5
8.7 ±0.7
9.3 ±1.4
3.6 ±0.5
3. 8 ±0.3
Source: Weaver et al. 2010
The Michaelis-Menton kinetic data (Km and Vmax [maximum initial rate of an enzyme
catalyzed reaction]) indicate that there are substantial differences in the affinity of the
perfluorinated carboxylate with 8 and 9 carbon chains for OATS, with the C8 acid favored over
the C9 acid. OATS 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 OATPlal, a
resorption transporter located on the apical side of the renal tubular cells, the C9 and CIO 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 CIO acid when all three fluorocarbons are present in the exposure matrix at
approximately equal concentrations. There were minimal kinetic differences between transport
Perfluorooctanoic acid (PFOA) - May 2016
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of the C7 and C8 acids by OAT1, an export transporter on the basolateral surface of the renal
tubular cells.
Based on the Hinderliter study (2004), a developmental change in renal transport occurs in
female rats between 3 and 5 weeks of age that allows for expedited excretion of PFOA. When
the transporters become active, there is a decrease in plasma PFOA levels and an increase in
urinary excretion (Table 2-23). The developmental change in male rats appears to have the
opposite effect. Sexual maturity appears to influence these events because castrated male rats
become more like females and OVX females become more like males in their PFOA excretion
capabilities. The change in female rats seems to involve the OATs (Kudo et al. 2002) while the
change in males seems to involve the OATPs (Cheng et al. 2006).
Table 2-23. Plasma and Urine PFOA Concentration 24-hr After Treatment
with 30 mg/kg PFOA
Age
(weeks)
3
4
5
Female
Plasma (jig/ml)
51.43 ±13.61
28.01 ±9.90
3.42 ± 1.95
Urine (jig/ml)
94.89 ± 26.26
104.12 ±28.97
123. 16 ±51.56
Male
Plasma (jig/ml)
74.16 ±18.23
100.81 ±13. 18
113. 86 ±23.36
Urine (jig/ml)
51.76 ±28.86
28.70 ±18.84
15.65 ±6.24
Source: Hinderliter 2004
When considered together, the studies of the transporters suggest that female rats are efficient
in transporting PFOA across the basolateral and apical membranes of the proximal kidney
tubules into the glomerular filtrate, but male rats are not. Males, on the other hand, have a higher
rate of resorption than females for the smaller amount they can transport into the glomerular
filtrate via OATPlal in the apical membrane. This scenario might explain the inverse
relationship between the levels of PFOA in female urine and plasma and the plateau of plasma
PFOA in male rats compared to their losses via urine.
Unfortunately, much work remains to be done to explain the gender differences between
male and female rats and to determine whether it is relevant to humans. Similarities are possible
because the long half-life in humans suggests that they might be more like the male rat than the
female rat. There is a broad range of half-lives in human epidemiology studies suggesting a
variability in the unbound fraction of PFOA in serum or in human transport capabilities resulting
from genetic variations in structures and consequently in function. Genetic variations in human
OATs and OATPs are described in a review by Zai'r et al. (2008).
2.6 Toxicokinetic Considerations
2.6.1 PK Models
One of the earliest PK models was done using the post-dosing plasma data from the
Butenhoff et al. study (2004b) in cynomolgus monkeys (Andersen et al. 2006). In this study,
groups of six monkeys (three per gender per group) were dosed for 26 weeks with 0, 3, 10, and
20 mg/kg PFOA (high-dose =30 mg/kg PFOA for the first 12 days), followed for >160 days after
dosing. Metabolism cages were used for overnight urine collection. Since urine specimens could
account for only overnight PFOA excretion, the total volume and total PFOA were extrapolated
to 24-hour values based on the excretion rate (volume/hour) for the volume collected and the
hours of collection.
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The Andersen et al. model (2006) was based on the hypothesis that saturable resorption
capacity in the kidney would possibly account for the unique half-life properties of PFOA across
species and genders. The model structure, shown in Figure 2-3, was derived from a published
model for glucose resorption from the glomerular filtrate via transporters on the apical surface of
renal tubule epithelial cells (Andersen et al. 2006).
Tissue
Compartment
\ "tissue!
input
(iv, oral)
Central
Compartment
(V,j; C,; fp,,^)
TM;KT
Filtrate
Compartment
-------
Input (oral)
Liver Compartment
(Liver volume; Free fraction in liver; Saturable binding)
Input (iv)
I
Fecal Elimination
Central Compartment
(Volume of distribution; Free fraction in serum; Plasma cone)
Tissue Comportment
(Amount in tissue)
Tm,
Filtrate Compartmtiit
(Volume of renal filtrate; Renal filtration rate,"
Soturable resorption)
Storage Comportment
Urinary elimination
Figure 2-4. Physiologically Motivated Pharmacokinetic Model Schematic for
PFOA-Exposed Rats
The models were parameterized and applied to the Kemper data (2003) for male and female
CD rats given doses of 1, 5, and 25 mg/kg/day. The model did not provide a satisfactory fit
between the predictions of plasma concentration or urine + fecal excretion and experimental data
for either gender.
Lou et al. (2009) used the data they collected on the serum, liver, and kidney PFOA
concentration (see section 2.2.1) in CD-I mice to examine if one- or two-compartment PK
models would fit the experimental data for 1,10, and 60 mg/kg/day single gavage doses (see
Figure 3-5 for the one-compartment model). Both models assumed first order absorption and
elimination. The two-compartment model included a central compartment that received PFOA
after absorption and transferred it to a second compartment for excretion. The excretion
compartment was coupled with bidirectional flow between the two compartments. The net loss
from the central compartment differed during and after distribution. The models were fit using a
general nonlinear least squares approach. A likelihood ratio squared approach was applied to
determine which model achieved the best fit to the data.
Perfluorooctanoic acid (PFOA) - May 2016
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Oral Dose
Ka
Central
Compartment
Ke
Urinary Loss
Figure 2-5. Schematic for One-Compartment Model.
The one compartment model performed well for serum, liver, and kidney in this analysis, and
output was not significantly improved with use of a two-compartment model. The input
parameters for the one-compartment model included Vd, serum half-life, and absorption rate
constant (Ka) and elimination rate constant (Ke) for serum, liver, and kidney. There were slight
differences in the fitted values between males and females for some parameters. The Ke values
were consistently higher in the female mice (Table 2-24). The quantitative measures for liver and
kidney were only available for the 1- and 10-mg/kg/day doses.
Table 2-24. Model Parameters for 1 and 10 mg/kg Single Doses of PFOA to CD1 Mice
Tissue
Serum
Liver
Kidney
Parameter (abbreviation)
Volume of distribution (Vd)
Absorption rate constant (Ka)
Elimination rate constant (Ke)
Half-life (T./2)
Volume of distribution (Vd)
Absorption rate constant (Ka)
Elimination rate constant (Ke)
Volume of distribution (Vd) 1 mg/kg
Volume of distribution (Vd) 10 mg/kg
Absorption rate constant (Ka)
Elimination rate constant (Ke)
Females
0.135 L/kg
0.537 L/hr
0.00185 L/hr
15.6 days
0.161 L/kg
0.5 170 L/hr
0.00161 L/hr
0.822 L/kg
1.092 L/kg
0.527 L/hr
0.00151 L/hr
Males
0.266 L/kg
0.00133 L/hr
21.7 days
0.120 L/kg
0.00129 L/hr
1.280 L/kg
1.170 L/kg
0.001 13 L/hr
Source: Lou et al. 2009
The one-compartment model described above was not able to predict serum concentration in
female mice given a single 60-mg/kg dose, suggesting a change in kinetics with the 60-mg/kg
dose compared to the 1- and 10-mg/kg doses. This conclusion is supported by comparison of the
serum measurements made during the 30-day post-dosing period for all three doses. The serum
PFOA concentration at the 60-mg/kg dose declined more rapidly with time than serum PFOA
concentrations at the 1- and 10-mg/kg doses. For example, a serum concentration of about
0.4 mg/L was reached in about 28 days at the 60-mg/kg dose, 61 days at the 10-mg/kg dose, and
70 days at the 1-mg/kg dose (values estimated from Figure 3, Lou et al. 2009). The one-
Perfluorooctanoic acid (PFOA) - May 2016
2-37
-------
compartment model also produced a poor fit for the serum level measurements taken 24 hours
after the cessation of a 17-day exposure to 20 mg/kg/day. The two-compartment model provided
a better fit with experimental serum concentration data for the single 60-mg/kg dose and the
repeat 20-mg/kg/day dose, but the fit was still unsatisfactory.
Lou et al. (2009) also tried the Andersen et al. renal-resorption model (2006) to determine if
it provided an improved fit for the data. The Andersen et al. model (2006) fit to the data was
superior to that of the one- and two-compartment models of Lou et al. (2009) for the 60-mg/kg
single-dose and the 20-mg/kg/day repeat-dose scenarios.
The Andersen et al. model (2006) includes a second tissue compartment that articulates with
the central compartment but not the filtrate compartment. In addition to values for Vd, Ka, and
Ke, the model includes values for cardiac output, volume for the renal filtrate, renal blood
filtration rate, intercompartmental CL, transport maximum, transport affinity constant (Kt), and
the proportion of free PFOA in serum. With the exception of body weight and cardiac output, the
input parameters for the model were either assumed (i.e., volume of renal filtrate and proportion
of free serum PFOA) or optimized for the model. The wide confidence bounds around the
optimized values are indicative of considerable parameter uncertainty.
The Lou et al. parameter estimates (2009) indicate that there may be several biological
limitations to the Andersen et al. (2006) PK model for adult mice including the fact that it
requires an unreasonably high portion of the cardiac output to pass through the kidneys to
optimize fit to the experimental data. It also does not include excretion via export transporters in
the renal tubular cells or consider that the bound fraction in the serum could vary with the
magnitude of the dose and duration of dosing. Much of the emerging data is consistent with a
variety of tubular transporters functioning in both efflux and resorption from the glomerular
filtrate. In addition, there are opportunities for protein binding within organs that could function
to retard distribution to the cytosol, especially at low doses. The binding of PFOA with L-FABP
is an example. Once binding sites are saturated, the concentration in the cytosol will increase.
A model also has been developed that applied to female CD-I 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-I and 12981/SvlmJ mice, respectively. Exposure
was assumed to be limited by blood flow, and only the experimental doses that did not impact
litter size (i.e., 0.1-1.0 mg/kg/day for CD-I mice and 1-10 mg/kg/day for 12981/SvlmJ mice)
were used in model development.
Lactational exposure was modeled as a dynamic relationship between the dam (n = 10) and
the litter, and they were connected by a milk compartment. Milk yield information was obtained
from the literature. Milk was assumed to be consumed as it was produced without any circadian
impact on consumption patterns. PFOA excreted in pup urine was routed back to the dam.
Both absorption and excretion were assumed to be first order processes as was lactation
transfer from the dam to the litter (Figure 2-6) (Rodriguez et al. 2009). Resorption of a portion of
the PFOA urinary efflux was included in the model. The renal excretion/resorption was
parameterized for cardiac output, kidney blood flow, GFR, urine flow rate, volume of renal
plasma (fraction of body weight), and volume of renal filtrate (fraction of body weight). The
fraction of free PFOA in serum reaching the glomerulus was assumed to be 0.01 based on
Perfluorooctanoic acid (PFOA) - May 2016 2-38
-------
protein binding information. As was the case with the Lou et al. model (2009), Rodriguez et al.
(2009) did not include parameters to adjust for transporter-mediated efflux from the renal tubular
cells into the glomerular filtrate.
2
*-4
1
§
S
!
Qcon
1,
Orel gavage
dose ,
(mg/kg)
a\
co
kad ,
Concept!
Veen Ccon
Dam
Cdam Vdam
Iklac
Uilk
t
/
Qcon
Kidney
fnw ' ' Qur
OOF
Qr-QGF
-">- Filtrate ( " '
. Tm;KI
t Renal 4-
Plasma
"
Qr-Qur
Pup excreta
' recirculation
(frambirtktoPNbU)
Cm Ik Vm
Iklac
Pups
Cpup Vpup
kep
Urine
Figure 2-6. PK Model of Gestation and Lactation in Mice
One of the limitations of the Rodriguez et al. modeling effort (2009) was the limited amount
of laboratory data against which to evaluate projections. Serum measures from the Lau et al.
(2006) and Abbott et al. (2007) studies were available for only a few time points. Nevertheless,
the authors reached several conclusions based on the model projections as follows:
The model had a tendency to overestimate serum levels, suggesting nonlinearity as doses
increased.
Gestation and lactation as a source of exposure contributed about equally to the pups of
129Sl/SvlmJ dams exposed only during gestation.
The contributions to the pups from gestation exceeded those from lactation in the CD-I
mice.
Exposure to the pups via lactation increased over time.
Lactation is a CL pathway for the dam.
A number of uncertainties accompany the model because of the assumptions regarding the
flow limitation on transport to the fetus and to maternal milk: the first order Ke for the pups and,
for the milk, the maternal serum partition coefficient and the limited knowledge regarding the
renal tubular transporters. They caution that the model should not be applied for cross-species or
high-to-low dose extrapolation.
Loccisano et al. (2011) developed a PFOA PBPK model for monkeys based on the Andersen
et al (2006) and Tan et al. (2008) models, and then extrapolated it for use in humans (Figure
2-7). The model reflects saturable renal absorption of urinary PFOA by the proximal tubule of
Perfluorooctanoic acid (PFOA) - May 2016
2-39
-------
the kidney. This is represented in Figure 2-7 by the interactions between the plasma and kidney
plus the interaction of the filtrate compartment with both plasma and kidney.
The fraction of PFOA free in plasma and available for glomerular filtration was based on
data fit and estimated to be less than 10% because of binding to serum proteins, especially
albumin. Lacking the kinetic data on tubular resorption, the rate was based on the best fit to the
plasma/urine data. A storage compartment was added to the model between filtrate and urine.
Tissue plasma partition coefficients were derived from the data by Kudo et al. (2007) following
the disposition pattern of a single intravenous (IV) dose to male Wistar rats.
Oral dose, drinking water
Plasma
Free
fraction
>
QGut
-,
QLiv
-,
QFat
QSkn
>
QR
-
QKid
QFil
Gut
\ t
Liver
Fat
Skin
Rest of body
Kidney
Tm,Kt
Filtrate
1
storage
kurine
urine
Notes'.
Tm = transporter maximum, Kt= affinity constant, and Q = flow in and out of tissues.
Figure 2-7. Structure of the PFOA PBPK Model in Monkeys and Humans
Existing IV and oral data sets from Butenhoff et al. (2004b) for the cynomolgus monkey
were used to develop the monkey model. In the oral study (section 2.2.1), animals were dosed for
6 months and followed for 90 days after dosing. Plasma and urine samples were analyzed
periodically during dosing and recovery. The model projections for the oral study were in good
agreement with the Butenhoff et al. data (2004b) for the 10-mg/kg dose, showing a rapid rise to
plasma steady state and a slow terminal half-life. The model performance for the high dose
(30/20 mg/kg/day) did not fit as well, partially as a consequence of the observed toxicity with the
initial 30 mg/kg/day dose that necessitated cessation of dosing on study day 12, followed by
resumption of dosing at 20 mg/kg/day on study day 22.
The structure of the human model was similar to that used for the monkeys. Human serum
data (means with standard deviations [SDs] or medians) for PFOA are available for occupational
and general populations (Bartell et al. 2010; Calafat et al. 2007a, 2007b; Emmett et al. 2006;
Perfluorooctanoic acid (PFOA) - May 2016
2-40
-------
Holzer et al. 2008; Olsen et al. 2005; Steenland et al. 2009). The fact that the serum data were
the results from measurements made following uncertain routes and uncertain exposure durations
presented a challenge in the assessment of model fit. The human half-lives used for the model
(3.8 and 2.3 years) came from an occupational study (Olsen et al. 2005) and a study of the Little
Hocking, Ohio, population after reduction of the PFOA in drinking water as a result of treatment
(Bartell et al. 2010). See section 2.6.2. Both half-life values were used in evaluating the model's
ability to predict serum concentration at the time the serum samples were collected.
The model produced results that can be characterized as fair to good when compared to the
reported average serum measurements. For the Little Hocking population studied by Emmett et
al. (2006), the model indicated the need for a 30-year exposure to reach steady-state
concentrations. The model indicated that both half-life values provided reasonable results when
compared to the measured serum values. The authors concluded that more data are needed on the
kinetics of renal transporters and intrahuman variability, plus more definitive information on
exposures to further refine the human model.
Fabrega et al. (2014) adapted the Loccisano et al. model (2011) to include compartments for
the brain and lung, and to remove the skin. They applied the adjusted model to humans by using
intake and body burden data from residents in Tarragona County, Spain. Food and drinking
water were the major sources of exposure. Body burden information came from blood samples
from 48 residents; tissue burdens came from 99 samples of autopsy tissues. The adjusted model
overpredicted PFOA serum levels by a factor of about 9, the liver by a factor of 4.5, and the
kidney by a factor of about 18. Model predictions for PFOS were far more consistent with the
tissue concentration experimental data.
The authors also looked at the impact of using data for partition coefficients from human
tissues in place of the Loccisano et al. rat data (2011) for the estimation of steady-state tissue
concentrations. The PFOA simulation values were closer to the human experimental data when
using the human partition coefficient values for the brain and lung, but not for the liver. In the
case of the kidney, the simulated projections were generally equivalent with both the human and
rat partition coefficients. The authors suggested that both saturable resorption and variations in
protein binding are important parameters for PK models. With the exception of serum albumin,
the existing models have not considered protein-binding constants within tissues. Even though
the use of human partition coefficients improved the steady-state predictions for tissues, overall
there were still considerable differences between the experimental values and the predictions
with both models.
Loccisano et al. (2012a) also used the saturable resorption hypothesis when developing a
model for adult Sprague-Dawley rats (Figure 2-8). The structure of the model is similar to that
for the monkey/human model depicted in Figure 2-7, but lacks the fat and skin compartments
and includes a storage compartment to accommodate fecal loss of unabsorbed dietary PFOA as
well as loss from biliary secretions. Oral and IV data used in model development came from
studies by Kemper (2003), Kudo et al. (2007) and Perkins et al. (2004). Partition coefficients for
liverplasma, kidney:plasma, and rest of the body:plasma were derived from unpublished data on
mice by DePierre (2009) through personal communication with the authors (Loccisano et al.
2012a). Most of the other kinetic parameters were based on values providing the best fit to the
experimental data. Because a number of the renal transporters involved with PFOA resorption
are known, available kinetic information was used where appropriate. Model performance was
evaluated primarily based on its ability to predict plasma and liver concentrations from the
Perfluorooctanoic acid (PFOA) - May 2016 2-41
-------
studies identified above. Performance was generally good given the limitations in the primary
data sources, as was the case for the monkey model.
Oral, diet
feoes
Figure 2-8. Structure of the PBPK Model for PFOA in the Adult Sprague-Dawley Rat
Loccisano et al. (2012b) expanded the adult Sprague-Dawley rat model to cover gestational
and lactational exposure to the fetus and pups through their dams. The data from Hinderliter et
al. (2005) were used in model development for both the gestation and lactation periods. The
gestational model structure for the dams is similar to the model structure shown in Figure 2-8.
The model was expanded to include the fetuses linked to the dams by way of the placenta.
Uptake from the placenta was described by simple diffusion; the fetal plasma compartment was
separate from the dams as was distribution to fetal tissues and amniotic fluid. Based on the
transporter data for PFOA, elimination differed for male and female rats and was considered to
be developmentally regulated, resulting in faster elimination for female rats than for male rats
after sexual maturation. The lactation model linked the pups to their dams through mammary
gland secretions. Pup compartments included the gut, liver, kidney, renal filtrate, plasma, and
rest of the body.
Model performance was judged by its ability to predict concentrations in maternal and fetal
plasma, amniotic fluid, and milk. The predictive capability of the model ranged from fair to
good, depending on the medium. The fit of the projections to the data was weakest for the whole
embryo during gestation, for which measured levels were greater than projection for two of three
data points and for neonate plasma during lactation, for which all data points fell below the
predictions.
Loccisano et al. (2013) extended their model development to cover humans during pregnancy
and lactation, building on the work done with rodents and recognizing the limitations of the
human data available for evaluating the model predictions. Figure 2-9 illustrates the structure of
Perfluorooctanoic acid (PFOA) - May 2016
2-42
-------
the model used. The basic structure was derived from the rat model discussed above. Following
are some of the key features of the model:
The fetus is exposed via the placenta through simple bidirectional diffusion.
Transfer rates to the fetus from the amniotic fluid are governed by bidirectional diffusion.
Transfer from the fetal plasma to tissues is flow-limited.
Maternal plasma is directly linked to the milk compartment and considered to be flow-
limited; only the free fraction in plasma is transferred to maternal milk.
The neonate is exposed to PFOA only via maternal milk for the first 6 months
postpartum.
The infant in the model is treated as a single compartment with a Vd.
Plasma
(free
available
for uptake
into
L 1
QGut
c
QLiv
^
QFat
s.
HDIs
QSk
^
j,
^
U,K
^
^
nkHH
^
QFII
Rut
I
Fat
Placenta
^kin
Mammary
Rest of body
Kidney
f
Filtrate
« Oral dose
Drinking f
ktransl
. j
*
ktrans2
Tm.Kt
- *l storage
t
120
plasma
(free
fraction)
j.
Amniotic fluid
/ \
w
^
Rest of fetal body
QRFet
urine
1
IV
Figure 2-9. PBPK Model Structure for Simulating PFOA and PFOS Exposure During
Pregnancy in Humans (Maternal, Left; Fetal, Right)
Limitations to the model are acknowledged and attributed primarily to lack of data to support
a more mechanistic approach. Physiological parameters applicable to a pregnant or lactating
woman, the fetus, and the nursing infant were obtained from a variety of referenced publications.
To obtain a plasma value at the time of conception, the model was run until it reached a
prepregnancy steady-state concentration. The model predicted 30 years as the exposure
necessary to reach steady state for the general female population (1 E-4 to 2 E-3 ng/kg body
weight [bw] /day). The model performance simulations for PFOA were run using an exposure of
1.5 x 10"4 |ig/kg bw/day. Projections were developed for maternal plasma, fetal plasma, infant
plasma, and maternal milk. Agreement between the observed concentrations (|ig/L) and the
predicted values was considered satisfactory if the predicted value was within 1% of the
Perfluorooctanoic acid (PFOA) - May 2016
2-43
-------
observed value. Model output was compared to maternal and fetal plasma values at delivery or at
specific time points, and for the infant plasma and milk data where available. Predicted
maternal:fetal plasma (cord blood) concentration ratios were more consistent for PFOA than for
PFOS when compared to the published data. The projections for fetal internal dose were
reasonable, and there was good agreement between the model and the available human lactation
data. The modeled maternal plasma was 2.4 jig/L at the time of conception; it slowly decreased
across the gestation period and increased slightly at delivery. For the most part, the modeled
results fell within ± 1 SD of the observed data.
During lactation, there was a decline in maternal plasma across the 6 months of lactation (a
change of l|ig/L). Thereafter, plasma values slowly increased and stabilized at about 1.5 jig/L at
6 months postpartum. The fetal plasma concentration was about 2.3 |ig/L at the start of gestation
and declined to about 1.8 |ig/L at the time of delivery. Maternal plasma values are about the
same as those for the fetus. During the lactation period, the infant plasma increased in a linear
fashion to a terminal value of about 5 |ig/L. Milk concentrations declined very slightly across the
lactation period with an initial concentration of 0.07 |ig/L and a final value of 0.05 |ig/L. These
concentrations were estimated from the graphic data presentation. Breast milk appears to be an
important excretory route for the dam.
The projections for PFOA differed from those for PFOS in several respects. Most
importantly, maternal and fetal plasma values were similar for PFOA but for PFOS, maternal
levels were approximately twofold higher than fetal levels. Compared with PFOS, there was a
much greater decline in maternal PFOA plasma values during lactation accompanied by a
comparable decline in the PFOA concentration in milk. The increase in infant plasma across the
lactation period was comparable for PFOA and PFOS, with the concentration at 6 months
postpartum about 2.5 times higher than at 1 month.
Loccisano et al. (2011, 2012a, 2012b, 2013) determined that the human pregnancy lactation
model results, when compared to published data, identified the following important research
needs:
Are there differences in the transporter preferences and transfer rates for the individual
PFASs? Do those differences correlate with half-life differences?
Are there qualitative or quantitative differences between the transporters favored by
PFOA and those favored by PFOS?
What physiological factors influence CL for the mother, the fetus, and the infant during
gestation and lactation?
Are placental transport processes active, facilitated, or passive?
These research needs are more pronounced for PFOS than PFOA because the information
supporting renal resorption and tissue uptake via membrane transporters for PFOS is very
limited. Most models infer that PFOS and PFOA are similar based on their half-lives rather than
on published research on PFOS transporter kinetics.
Building on the work of other researchers, Wambaugh et al. (2013) developed and published
a PK model to support the development of an EPA RfD for PFOA. The model was applied to
data from studies conducted in monkeys, rats, and mice that demonstrated an assortment of
systemic, developmental, reproductive, and immunological effects. A saturable renal resorption
PK model was used. This concept has played a fundamental role in the design of all of the
published PFOA models summarized in this section. In this case, an oral dosing version of the
Perfluorooctanoic acid (PFOA) - May 2016 2-44
-------
original model introduced by Andersen et al. (2006) and summarized early in section 2.6.1 was
selected for having the fewest number of parameters that would need to be estimated. A unique
feature of the Wambaugh et al. approach was to use a single model for all species in the
toxicological studies to examine the consistency in the average serum values associated with
effects and with no effects from nine animal studies of PFOA. The model structure is depicted in
Figure 2-3, with minor modifications.
Wambaugh et al. (2013) placed bounds on the estimated values for some parameters of the
Andersen et al. model (2006) to support the assumption that serum carries a significant portion of
the total PFOA body load. The Andersen et al. model is a modified two-compartment model in
which a primary compartment describes the serum and a secondary deep tissue compartment acts
as a specified tissue reservoir. Wambaugh et al. (2013) constrained the total Vd to a value of not
more than 100 times that in the serum. As a result, the ratio of the two volumes (serum versus
total) was estimated in place of establishing a rate of transfer from the tissue to serum.
A nonhierarchical model for parameter values was assumed. Under this assumption, a single
numeric value represents all individuals of the same species, gender, and strain. The gender
assumption was applied to monkeys and mice while male and female rats were treated separately
because of the established gender-related toxicokinetic differences. Body weight, number of
doses, and magnitude of the doses were the only parameters to vary. In place of external doses,
serum concentrations as measured at the time of euthanasia were used as the metric for dose
magnitude. Measurement errors were assumed to be log-normally distributed. Table 2-25
provides the estimated and assumed PK parameters applied in the Wambaugh et al. model (2013)
for each of the species evaluated.
The PK data that supported the analysis were derived from two PFOA PK in vivo studies.
The monkey PK data were derived from Butenhoff et al. (2004b), and the data for the rats (M/F)
were from Kemper (2003). Two strains of female mice were analyzed separately, with CD1
information derived from Lou et al. (2009) and C57B1/6 information derived from DeWitt et al.
(2008). The data were analyzed within a Bayesian framework using Markov Chain Monte Carlo
sampler implemented as an R package developed by EPA to allow predictions across species,
strains, and genders and to identify serum levels associated with the NOAEL and LOAEL
external doses. The model chose vague, bounded prior distributions on the parameters being
estimated, allowing them to be significantly informed by the data. The values were assumed to
be log-normally distributed, constraining each parameter to a positive value.
The model predictions were evaluated by comparing each predicted final serum
concentration to the serum value in the supporting animal studies. The predictions were generally
similar to the experimental values. There were no systematic differences between the
experimental data and the model predictions across species, strain, or gender, and median model
outputs uniformly appeared to be biologically plausible despite the uncertainty reflected in some
of the 95th percentile credible intervals. The application of the model outputs in deriving a human
RfD is the focus of section 4.0 of this document.
Perfluorooctanoic acid (PFOA) - May 2016 2-45
-------
Table 2-25. Pharmacokinetic Parameters from Wambaugh et al. (2013) Meta-Analysis of
Literature Data
Parameter
bwb
Cardiac
Output"
*a
77
r CC
ku
Rv2:V\
L maxc
^T
Free
gfilc
Vfilc
Units
kg
L/h/kga74
L/h
L/kg
L/h
Unitless
umol/h
umol
Unitless
Unitless
L/kg
CD1 Mouse (f)a
0.02
8.68
290 (0.6 -
73,000)
0.18(0.16-2.0)
0.012 (3.1 xe-10-
38,000)
1.07(0.26-5.84)
4.91(1.75-2.96)
0.037 (0.0057 -
0.17)
0.011(0.0026-
0.051)
0.077(0.015-
0.58)
0.00097 (3.34 x
e-9-7.21)
C57B1/6 Mouse
(f)a
0.02
8.68
340 (0.53 -
69,000)
0.17(0.13-2.3)
0.35(0.058-52)
53(11-97)
2.7 (0.95 - 22)
0.12(0.033-
0.24)
0.034(0.014-
0.17)
0.017(0.01-
0.081)
7.6 x e-5 (2.7 x
e-10-6.4)
Sprague-Dawley
Rat (f)a
0.20(0.16-0.23)
12.39
1.7(1.1-3.1)
0.14(0.11-0.17)
0.098(0.039-
0.27)
9.2(3.4-28)
1.1(0.25-9.6)
1.1(0.27-4.5)
0.086(0.031-
0.23)
0.039(0.014-
0.13)
2.6xe-5(2.9x
e-10-28)
Sprague-Dawley
Rat (m)a
0.24 (0.21 -0.28)
12.39
1.1(0.83-1.3)
0.15(0.13-0.16)
0.028 (0.0096 -
0.08)
8.4(3.1-23)
190(5.5-
50,000)
0.092 (3.4 xe-4-
1.6)
0.08 (0.03 - 0.22)
0.22(0.011-58)
0.0082(1. 3 xe-8
-7.6)
Cynomolgus
Monkey (m/f)a
7 (m), 4.5 (f)
19.8
230 (0.27 -
73,000)
0.4(0.29-0.55)
0.001 1(2.4 xe-10
- 35,000)
0.98(0.25-3.8)
3.9(0.65-9,700)
0.043 (4.3 x e-5 -
0.29)
0.01(0.0026-
0.038)
0.15(0.02-24)
0.002 1(3. 3 xe-9
-6.9)
Notes:
Means and 95% confidence interval (in parentheses) from Bayesian analysis are reported. For some parameters, the distributions
are quite wide, indicating uncertainty in that parameter (i.e., the predictions match the data equally well for a wide range of
values).
m = male, f = female
a Data sets modeled for the GDI mouse were from Lou et al. (2009), for the C57B1/6 mouse were from DeWitt et al. (2008), for
the rat were from Kemper (2003), and for the monkey from Butenhoff et al. (2004b).
b Estimated average body weight for species used except with Kemper study (2003) where individual rat weights were available
and assumed to be constant.
c Cardiac outputs obtained from Davies and Morris (1993).
2.6.2 Half-Life Data
Human. There have been several studies of half-lives in humans and all support a long residence
time for serum PFOA with estimates measured in years rather than months or weeks. Bartell et
al. (2010) determined an average half-life of 2.3 years based on a study of the decreases in
human serum levels after treatment of drinking water for PFOA removal was instituted by the
Lubeck Public Services District in Washington, West Virginia, and the Little Hocking Water
Association (LHWA) in Ohio. Source waters for these systems had become contaminated with
PFAS from the DuPont Works Plant in Washington, West Virginia, between 1951 and 2000.
The Bartell et al. study (2010) was based on a series of serum measurements (eight over
4 years) from 200 individuals who agreed to participate in the study. Inclusion criteria for the
participants included: serum PFOA concentrations > 50 ng/mL, residential water service
provided by one of the two treatment plants, never employed at the DuPont plant, not growing
their own vegetables, and signed acceptance of the study consent form. The participants were
almost equally divided between males and females with an average age of about 50 years (range
of 18-89 years). Most of the participants consumed public tap water (172) as their primary
source, but a small number (28) consumed bottled water as their source.
Perfluorooctanoic acid (PFOA) - May 2016
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The participants were required to report that they primarily used home tap water for cooking,
bathing, and showering for the years between 2005 and 2007. The tap water users had to report
public water as their primary source of residential water consumption, and bottled water users
had to report the use of bottled water as their primary source of residential water consumption.
The initial blood draw for serum occurred in June 2007, with subsequent samples at 1, 2, 3, 6,
and 12 months after the initial sample. Samples were analyzed by the Centers for Disease
Control and Prevention. Nineteen samples from the 2-month blood draw were not analyzed due
to mislabeling.
A linear mixed model was used to determine the decline in serum PFOA concentration over
time. With these models, the decline from baseline by the participants was essentially first order.
The serum PFOA concentration was the only time-varying measurement entered into the model.
Serum concentrations were log-normally distributed, as described by the following equation:
InC = InCO-kt
where:
C= serum concentration at time t
CO = baseline serum concentration
k = elimination rate constant
t = time point for the measurement
The results of this assessment showed a 26% decrease in PFOA concentration per year after
adjustment for covariates and a half-life of 2.3 years [confidence interval (CI) = 2.1-2.4]. The
covariates considered included the water treatment system, the time exposed before and after
filtration, public versus bottled water, gender, age, consumption of local or homegrown
vegetables, and exposure to the public water supply at work. The only potential confounders
determined to be significant were the treatment plant (p = 0.03) and homegrown vegetable
consumption (p<0.001).
Identification of consumption of homegrown vegetables as a significant confounder revealed
a weakness in the study design because it had been an exclusion factor, yet was identified as an
exposure source at the 12-month interview of the study participants. The researchers concluded
that this problem was a result of the way the exclusion question was phrased for the original
interview, "Do you grow your own vegetables?" When the question was asked later in the study,
it was rephrased, "Do you eat any fruits and vegetables grown at your own home?" Some people
who answered "no" to the original questions answered "yes" to the second question.
Changes in the source of drinking water during the study could also have impacted the
results. When baseline interview data were compared with the results from the 12-month
interview, 39% of the bottled water group reported using public water at home. Some of the
public water drinkers (10%) reported using primarily bottled water at the 6-month interview.
In another study, the drinking water supply was contaminated with a mixture of
perfluorinated chemicals when a soil-improver mixed with industrial waste was applied upriver
to agricultural lands in Arnsberg, Germany (Brede et al. 2010). The PFOA levels in the finished
drinking water were measured as 500-640 ng/L in 2006. PFOS and PFHxS also were present.
The plasma PFOA levels in the Arnsberg population were 4.5 to 8.3 times higher than those in a
reference community at the time the problem was discovered. Charcoal filtration was added to
the potable treatment train and succeeded in reducing concentrations in the drinking water.
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The authors used the differences in plasma 2008 PFOA measurements from a subset of the
participants (children and adults) initially exposed in 2006 to determine the PFOA half-life. The
2008 subjects included 66 males, females, and children from Arnsberg and 73 from the reference
community in the evaluation. The drinking water concentration monitoring results (nondetects
estimated as one-half of the limit of detection [LOD], 10 ng/L) and DWI estimates obtained by
questionnaire and interview were used to estimate PFOA exposures. Plasma PFOA samples were
collected during a 2-month period in late 2008. Plasma PFOA had declined in the serum for both
the Arnsberg residents (39.2%) and those from the reference community (13.4%). In Arnsberg,
the decrease was greater for the exposed females and children than for the males when compared
to the reference community, an observation that appeared to reflect the reported lower DWIs of
the Arnsberg females and children (0.3 ± 0.2 and 0.8 ± 0.6 L/day compared to 0.7 ± 0.5 and
1.6 ± 0.8 L/day, respectively). The estimate for the human half-life was 3.26 years (geometric
mean; range 1.03-14.67 years). Regression analysis of the data also suggested that the
elimination rate might have been greater in younger subjects and older subjects.
Seals et al. (2011) determined half-life estimates for 602 residents of Little Hocking, Ohio,
and 971 residents of Lubeck, West Virginia, who were part of the C8 study but had relocated to a
different area of the country. The half-life estimate was based on the decline in serum PFOA
levels after the time of the initial measurement and the years since the change in residential
location occurred. A background estimate (5 ng/mL) was subtracted from the serum
measurements before analysis. On average, the initial serum PFOA concentrations were higher in
Little Hocking (60.6 ng/mL) than in Lubeck (31.0 ng/mL). Due to the nonlinearity in scatter
plots of the natural log for adjusted serum PFOA concentrations versus the years elapsed since
relocation, the authors used a two-segment linear spline regression approach in their analysis of
the data (i.e., Little Hocking4 years, Lubeck9 years). The slope of the line decreased for the
second time segment compared to the first. In former residents of Little Hocking, a -21.4%-
change in serum PFOA was observed in the first 4 years after leaving Little Hocking, and
a -7.6%-change was observed beyond 4 years. In former Lubeck residents, the serum PFOA
change was -7.8% for the first 9 years and 0.2% (a slight increase) afterwards. The half-life
estimates for Little Hocking ranged from 2.5-3.0 years (average 2.9 years) and for Lubeck
ranged from 5.9-10.3 years (average 8.5 years).
Based on their analysis, the authors suggested that, if their assumptions were correct, a
simple first order elimination model might not be appropriate for PFOA given that the rate of
elimination appeared to be influenced by both concentration and time. There was a difference in
the CL for the two locations even though the range of years elapsed since relocation was the
same for both communities. The authors identified three potential limitations of their analysis:
the cross-sectional design, the assumption that exposure was uniform within a water district, and
a potential bias introduced by exclusion of individuals with serum values <15 ng/mL.
3M (Burris et al. 2000, 2002) conducted a half-life study on 26 retired fluorochemical
production workers from their Decatur, Alabama, (n = 24) and Cottage Grove, Minnesota,
(n = 3) plants. Blood was collected from the subjects between 1998 and 2004, a period during
which serum samples were drawn every 6 months over a 5-year period, depending on the facility
at which the subject had worked. Responses on questionnaires determined whether any of the
retirees had occupational exposures after retirement. The average number of years that
participants worked was 31 (range 20-36 years) and they had been retired an average of
2.6 years at study initiation (range 0.4-11.5 years). The mean age of the retirees was 61 years
(range 55-75) at the beginning of the study.
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The initial mean serum PFOA concentration of all of the subjects was 0.691 |ig/ml (range
0.072-5.1 |ig/mL). At the completion of the study, the mean PFOA concentration was
0.262 |ig/mL (range 0.017-2.435 |ig/mL). Two of the retirees died during the study period;
therefore, they were only followed for 4.2 years. The mean serum elimination half-life of PFOA
in these workers was 3.8 years (1378 days, 95% CI, 1131-1624 days) and the median was
3.5 years (Olsen et al. 2005). The range was 1.5-9.1 years (561-3334 days). No association was
reported between the serum elimination half-life and with initial PFOA concentrations, age, or
gender of the retirees, the number of years retired or working at the production facility, or
medication use or health conditions.
Harada et al. (2005) studied the relationship between age, gender, and serum PFOA
concentration in residents of Kyoto, Japan. They found that females in the 20-50-year-old age
group (all with regular menstrual cycles) had serum PFOA concentrations that were significantly
lower than those in females over age 50 (all postmenopausal). Mean serum PFOA concentration
in the younger females was 7.89 ± 3.61 ng/ml versus 12.63 ± 2.42 ng/mL in the older females.
This age difference in serum PFOA concentrations was not seen in males, and serum PFOA
concentrations in males were comparable to those of the older females.
Harada et al. (2005) also estimated the CLR rate of PFOA in humans and found it to be only
about 0.001% of the GFR. There was no significant difference in CLR of PFOA with respect to
gender or age group, and the mean value was 0.03 ± 0.013 ml/day/kg.
Animal. Kemper (2003) examined the plasma concentration profile of PFOA following gavage
administration in sexually mature Sprague-Dawley rats. Male and female rats (four per gender
per group) were administered single doses of PFOA by gavage at DRs of 0.1, 1, 5, and 25 mg
PFOA/kg. After dosing, plasma was collected for 22 days in males and 5 days in females.
Plasma concentration versus time data were then analyzed using noncompartmental PK methods
(see Table 2-26 and Table 2-27). To further characterize plasma elimination kinetics, animals
were given oral PFOA at a rate of 0.1 mg/kg, and plasma samples were collected until PFOA
concentrations fell below quantitation limits (extended time).
Plasma elimination curves were linear with respect to time in male rats at all dose levels. In
males, plasma elimination half-lives were independent of dose level and ranged from
approximately 138 hours to 202 hours. To further characterize plasma elimination kinetics,
particularly in male rats, animals were given oral PFOA at a dose of 0.1 mg/kg, and plasma
samples were collected until PFOA concentrations fell below quantitation limits (2,016 hours in
males). The estimated plasma elimination half-life in this experiment was approximately
277 hours (11.5 days) in male rats.
Plasma elimination curves were biphasic in females at the 5-mg/kg and 25-mg/kg dose
levels. In females, terminal elimination half-lives ranged from approximately 2.8 hours at the
lowest dose to approximately 16 hours at the high dose. The estimated plasma elimination half-
life in the extended time experiment was approximately 3.4 hours in females.
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Table 2-26. PK Parameters in Male Rats Following Administration of PFOA
Parameter
Tmax (hr)
Cmax (ng/mL)
Lambda z (1/hr)
Ti/2 (hr)
AUCiNF (hr ng/mL)
AUCiNF/D
(hr ng/ml/mg/kg)
Clp (mL/kghr)
Dose
0.1 mg/kg
10.25
(6.45)
0.598
(0.127)
0.004
(0.001)
201.774
(37.489)
123.224
(35.476)
1096.811
(310.491)
0.962
(0.240)
1 mg/kg
9.00
(3.83)
8.431
(1.161)
0.005
(0.001)
138.343
(31.972)
1194.463
(215.578)
1176.009
(206.316)
0.871
(0.158)
5 mg/kg
15.0
(10.5)
44.75
(6.14)
0.0041
(0.0007)
174.19
(28.92)
6733.70
(1392.83)
1221.89
(250.28)
0.85
(0.21)
25 mg/kg
7.5
(6.2)
160.0
(12.0)
0.0046
(0.0012)
157.47
(38.39)
25,155.61
(7276.96)
942.65
(284.67)
1.13
(0.31)
1 mg/kg (IV)
NA
NA
0.004
(0.000)
185.584
(19.558)
1249.817
(113.167)
1123.384
(100.488)
0.896
(0.082)
0.1 mg/kg
extended
time
5.5
(7.0)
1.08
(0.42)
0.0026
(0.0007)
277.10
(56.62)
206.38
(59.03)
2111.28
(586.77)
0.51
(0.17)
Source: Kemper 2003
Notes:
Mean (SD)
AUCiNF: area under the plasma concentration time curve, extrapolated to infinity; AUCiNF/D: AUCiNF normalized to dose; Clp:
plasma clearance; Cmax: maximum plasma concentration; Lambda z: terminal elimination constant; Ll/2: terminal elimination
half-life; Tmax: time to Cmax.
Table 2-27. PK Parameters in Female Rats Following Administration of PFOA
Parameter
Tmax (hr)
Cmax (Hg/mL)
Lambda z (1/hr)
Ti/2 (hr)
AUCiNF (hr ng/mL)
AUCmF/D
(hr ng/mL/mg/kg)
Clp (mL/kghr)
Dose
0.1 mg/kg
0.56
(0.31)
0.67
(0.07)
0.231
(0.066)
3.206
(0.905)
3.584
(0.666)
31.721
(5.880)
32.359
(6.025)
1 mg/kg
1.13
(0.63)
4.782
(1.149)
0.213
(0.053)
3.457
(1.111)
39.072
(10.172)
38.635
(10.093)
27.286
(7.159)
5 mg/kg
1.50
(0.58)
20.36
(1.58)
0.15
(0.02)
4.60
(0.64)
114.90
(11.23)
20.78
(2.01)
48.48
(4.86)
25 mg/kg
1.25
(0.87)
132.6
(46.0)
0.059
(0.037)
16.22
(9.90)
795.76
(187.51)
29.54
(6.92)
35.06
(.88)
1 mg/kg
(IV)
NA
NA
0.250
(0.047)
2.844
(0.514)
33.998
(7.601)
30.747
(6.759)
34.040
(9.230)
0.1 mg/kg
extended
time
1.25
(0.50)
0.52
(0.08)
0.22
(0.07)
3.44
(1.26)
O O A
3.34
(0.32)
34.39
(3.29)
29.30
(3.06)
Source: Kemper 2003
Note: Mean (SD)
Gibson and Johnson (1979) administered a single dose of 14C-PFOA averaging 11.4 mg/kg
by gavage to groups of three male 10-week-old CD rats. The elimination half-life of 14C from the
plasma was 4.8 days. NRC ([2005], cited in Butenhoff et al. [2004b]) reported half-lives of 4-6
days for male rats and 2-4 hours for female rats; there was no mention of the strains studied.
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Kemper (2003) reported half-lives of 6-8 days for male Sprague-Dawley rats (Table 2-26) and
3-16 hours for females (Table 2-27).
Lou et al. (2009) determined values of 21.7 days (95% confidence interval: 19.5-24.1) for
male CD1 mice and 15.6 days (95% confidence interval: 14.7-16.5) for females for use in their
pharmacokinetic model (see section 2.6.1). NRC ([2005], cited in Butenhoffet al. [2004b])
provided values of 12 days for males and 20 days for females without any information on strains.
Butenhoffet al. (2004b) looked at the elimination half-life in monkeys treated for 6 months
with 0, 3, 10, and 20 mg/kg/day via capsules. Elimination of PFOA from serum after cessation of
dosing was monitored in recovery monkeys from the 10- and 20-mg/kg dose groups. For the two
monkeys exposed to 10 mg/kg, serum PFOA elimination half-life was 19.5 (R2=0.98) days and
indicated first-order elimination kinetics. For three monkeys exposed to 20 mg/kg, serum PFOA
elimination half-life was 20.8 days (R2=0.82) and also indicated first-order elimination kinetics,
although dosing was suspended at different time points because of weight loss. The data from
NRC (2005), which were provided by Butenhoffet al. (2004b), were about 21 days for females
and 30 days for males.
2.6.3 Volume of Distribution Data
Several researchers have attempted to characterize PFOA exposure and intake in humans
through PK modeling (Lorber and Egeghy 2011; Thompson et al. 2010). As an integral part of
model validation, the parameter for Vd of PFOA within the body was calibrated from the
available data. In the models discussed below, Vd was defined as the total amount of PFOA in
the body divided by the blood or serum concentration.
Two groups of researchers defined a Vd of 170 ml/kg body weight for humans for use in a
simple, single compartment, first-order PK model (Lorber and Egeghy 2011; Thompson et al.
2010). The models developed by these groups were designed to estimate intakes of PFOA by
young children and adults (Lorber and Egeghy 2011) and the general population of urban areas
on the east coast of Australia (Thompson et al. 2010). In both models, the Vd was calibrated
using human serum concentration and exposure data from the NHANES and assumes that most
PFOA intake is from contaminated drinking water. Thus, in using the models to derive an intake
from contaminated water, the Vd was calibrated so that model prediction of elevated blood levels
of PFOA matched those seen in residents.
Butenhoffet al. (2004b) calculated a Vd from noncompartmental PK analysis of data from
cynomolgus monkeys. Three males and three females were administered a single IV dose of
10 mg/kg, and serum PFOA concentrations were measured in samples collected up to 123 days
post-dosing. The Vd of PFOA at steady state (Vdss) were similar for both genders at 181 ± 12
ml/kg for males and 198 ± 69 ml/kg for females.
2.6.4 Toxicokinetic Summary
Uptake and egress of PFOA from cells is largely regulated by transporters in cell membranes.
It is absorbed from the gastrointestinal tract as indicated by serum measurements in humans and
treated animals. In serum, PFOA is electrostatically bound to albumin occupying up to nine to
twelve sites and sometimes displacing other substances that normally would occupy a site.
Linear PFOA chains display stronger binding than branched chains. PFOA binding causes a
Perfluorooctanoic acid (PFOA) - May 2016 2-51
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change in the conformation of serum albumin, altering its ability to bind with some endogenous
and exogenous materials it normally transports.
PFOA is distributed to tissues by a process requiring membrane transporters. Accordingly,
the tissue levels vary from organ to organ. The highest tissue concentrations are usually in the
liver. Liver accumulation in males is greater than in females. Other tissues with a tendency to
accumulate PFOA are the kidneys, lungs, heart, and muscle, plus the testes in males and uterus in
females. Post-mortem studies in humans have found PFOA in liver, lungs, bone, and kidneys,
but only low levels in brain. PFOA is not metabolized, thus, any effects observed in toxicological
studies are the result of parent compound, not metabolites.
Electrostatic interactions with proteins are an important toxicokinetic feature of PFOA.
Studies demonstrate binding or interactions with membrane receptors (e.g., PPARa, T3),
transport proteins, and enzymes. Saturable renal resorption of PFOA from the glomerular filtrate
via transporters in the kidney tubules is a major contributor to the long half-life of this
compound. Branched-chain PFOAs are less likely to be resorbed than the linear molecules based
on half-life information in humans. All toxicokinetic models for PFOA are built on the concept
of saturable renal resorption first proposed by Andersen et al. (2006). Some PFOA is removed
from the body with bile, a process that also is transporter-dependent. Accordingly, the levels in
fecal matter represent both unabsorbed material and that discharged with bile.
During pregnancy, PFOA is present in the placenta and amniotic fluid in both animals and
humans. Post-delivery, PFOA is transferred to offspring through lactation in a dose-related
manner. Maternal serum levels decline as those in the pups increase. This also occurs in humans
as demonstrated in a study of females breast-feeding their infants in Little Hocking, Ohio.
The half-life in humans for occupationally exposed workers was 3.8 years (95% CI, 1.5-9.1).
The average half-life was 2.3 years among people in the Lubeck Public Services District in West
Virginia and the LHWA in Ohio, based on changes in serum levels for the general population
after treatment of drinking water was implemented. This half-life value reflects humans whose
exposure came primarily from their public water system. Half-lives from animals included
21 days (females) and 30 days (males) for monkeys Butenhoff et al. (2004b); 11.5 days (males)
and 3.4 hours (females) in Sprague-Dawley rats (Kemper 2003); and 27.1 days (male) and
15.6 days (female) CD1 mice (Lau et al. 2006). The gender difference between male and female
rats is not seen in mice. In early life, the half-lives are nearly the same for both genders of rats,
but once the animals reach sexual maturity, resorption increases in males, prolonging the half-
life (Hinderliter 2004; Hundley et al. 2006). This change appears to be under the control of
hormones in both males and females (Cheng et al. 2006; Kudo et al. 2002).
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3 HAZARD IDENTIFICATION
This section provides a summary and synthesis of the data from a large number of human
epidemiology studies accompanied by studies in laboratory animals designed to identify both the
dose response and critical effects that result from exposures to PFOA and to examine the MoA
leading to toxicity.
3.1 Human Studies
Epidemiology studies of effects of PFOA have been conducted in three types of populations:
workers exposed in chemical plants producing or using PFOA, high-exposure communities
(i.e., an area in West Virginia and Ohio that experienced water contamination over a period of
more than 20 years), and general population studies with background exposures. These
populations differ with respect to exposure levels. The approximate range in serum PFOA
concentrations is 0.010-> 2.0 (means around 1-4 jig/mL) in the PFOA-exposed workers, and
0.010-0.100 |ig/mL and below LOD to < 0.010 |ig/mL in the high-exposure community and
general population settings, respectively. Although moderate-to-high correlations between PFOA
and PFOS are often seen in general populations (r > 0.5), the correlation is lower in the West
Virginia and Ohio high-exposure area (r=0.3). In evaluating and synthesizing results from these
studies, it is important to consider differences in the exposure range within the study population
and the exposure level within the referent group, as differences (or inconsistencies) can be
expected depending on the shape of the exposure-response curve and the exposure range
encompassed by different studies. In addition, the optimal choice of an exposure metric
(e.g., cumulative or a time-specific) depends on the specific outcome being examined.
Occupational studies. Large-scale production of PFOA occurred in the United States for several
decades. Both 3M (in Alabama and Minnesota) and DuPont (in West Virginia) have been the
primary U.S. producers and users of perfluorinated compounds, and both companies have
offered voluntary fluorochemical medical surveillance programs to workers at plants that
produced or used perfluorinated compounds. The monitoring data collected by 3M and DuPont
were used in conjunction with mortality and health effects information in a number of
epidemiology studies of cancer and noncancer outcomes in the worker populations. 3M
discontinued manufacturing PFOA in 2000, but a subsidiary in Europe (Antwerp, Belgium)
continued to manufacture and sell it through 2008.
High-exposure community studies. Members of the general population living in the vicinity of
the DuPont Washington Works PFOA production plant in Parkersburg, West Virginia, are the
focus of a large-scale, community-based study titled the C8 Health Project. Releases from the
Washington Works plant, where PFOA (C8) was used as a processing aid in the manufacture of
fluoropolymers, contaminated the ground water of six water districts near the plant resulting in
exposures to the general population. The plant began production in the 1950s, with PFOA use
and emission from the plant increasing in the 1980s. Study participants from the affected areas
(n = 69,030; 33,242 males, 35,788 females; <10-70+ years) were identified in 2005-2006, and a
series of studies were conducted. The participants all received compensation and provided a
blood sample and filled out an extensive questionnaire that included information on drinking
water sources, use of home-grown produce, and health information. A variety of approaches to
exposure assessment have been used in these studies, with the most detailed incorporating
individual residential history and water consumption and source data, emissions data,
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environmental characteristics, water pipe installation, PK data, and workplace exposures (Barry
et al. 2013; Shin et al. 2011; Vieira et al. 2013). The methods allow the estimation of cumulative
and of current exposure at different time periods or ages for individual study participants. Details
of the specific analyses undertaken to estimate historical exposures and to ascertain different
types of outcomes (retrospective and prospective analyses) are described in detail below.
Drinking water concentrations were based on PFOA releases from the DuPont plant and
residential address history of the participants (C8 Science Panel 2012).
The C8 Health Project also involved a review of evidence of health effects, considering their
own studies and studies conducted by others and in other populations. The conclusions for each
health endpoint assessed"probable link" or "not a probable link"are available on the C8
Science Panel website in a series of reports completed in 2011-2012 (see
http://www.c8sciencepanel.org/index.html).
General population studies. Studies investigating the association between PFOA levels and
health effects in the U.S. general population have been conducted using the NHANES data set.
NHANES examined representative members of the U.S. population (-5000 adults and
children/year) through surveys focusing on different health topics. The study consists of an
interview (demographic, socioeconomic, dietary, and medical questions) and examination
(medical including blood and urine collection, and dental and physiological parameters).
Biomonitoring included a number of PFAS, predominantly PFOA and PFOS.
A study by Jain (2014) examined the influence of diet and other factors on the levels of
serum PFOA and other PFAS using NHANES 2003-2004, 2005-2006, and 2007-2008 data.
Significantly higher serum PFOA levels were found in males (0.0047 |ig/mL) than in females
(0.035 |ig/mL) and in smokers (0.043 jig/mL) than in nonsmokers (0.040 jig/mL). No significant
differences in PFOA serum concentration were seen during the time periods evaluated. There
was a positive association of PFOA with increases in serum cholesterol (p<0.001), serum
albumin (p<0.001), and body mass index (BMI) (p<0.04) based on the 5,591 records used in the
assessment. Intakes of nonalcoholic beverages were positively associated with serum PFOA
(P<0.001), but no associations were found for other dietary food groupings.
The results of these studies along with other population studies are described in the following
sections. In the studies of worker cohorts, the data collected focused on measures of
cardiovascular risk, signs of organ damage, standard hematological endpoints, and cancer
(primarily cancer-related mortality). Within the general population, data were focused on
cardiovascular risk factors and diabetic or prediabetic conditions as well as reproductive and
developmental endpoints. The following summary focuses on measures of lipids (e.g.,
cholesterol, LDL); liver, kidney, and thyroid effects; reproductive effects (e.g., pregnancy-related
outcomes, specifically pregnancy-related hypertension and preeclampsia, measures of fetal
growth, and pubertal development); and cancer (specifically kidney and testicular cancer). These
outcomes were selected either because of the availability of studies in a variety of settings with
some indications of effects (e.g., as noted in the C8 Science Panel reports), or to allow
comparison with results from studies in animals. Summary tables are included to support
evaluation of the weight of evidence and facilitate comparison of the serum concentrations in the
epidemiology studies to those in the animals studies summarized in section 3.2.
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3.1.1 Noncancer
3.1.1.1 Serum Lipids and Cardiovascular Diseases
Serum Lipids
Occupational studies. Four cross-sectional studies are described in this section and in Table 3-1.
Olsen et al. (2000) analyzed data from voluntary medical surveillance examinations of PFOA
production workers at a 3M plant in 1993, 1995, and 1997. Cholesterol, LDL, HDL, and
triglycerides were measured in male workers (n = 111 in 1993, n = 80 in 1995, and n = 74 in
1997). Multivariable regression analyses, conducted separately by year (cross-sectional), were
adjusted for age, BMI, alcohol consumption, and cigarette use. Employees' serum PFOA levels
were stratified into three categories<1, 1- <10, and >10 ug/mL. The sample size in the highest
category ranged from 11 to 15 in the three examination years. There was little variation by
exposure category in mean or median TC, LDL, HDL, or triglycerides across the workers in
1993, 1995, or 1997.
Olsen and Zobel (2007) examined data from the 2000 medical surveillance program at the
three 3M plants, which is an expanded and refined analysis of the data reported in Olsen et al.
(2003). The fluorochemical workers consisted of males (age 21-67) from the Antwerp, Belgium
(n = 196); Cottage Grove, Minnesota (n = 122); and Decatur, Alabama (n = 188) production
facilities who volunteered to participate in the medical surveillance program and did not take
cholesterol-lowering medication. Blood was collected for fluorochemical concentration
determination and serum lipid parameters including cholesterol, LDL, HDL, and triglycerides.
Analysis of variance (ANOVA), analysis of covariance, logistic regression, and multiple
regression models were used to analyze the data with age, BMI, and alcohol consumption as
covariates. Potential associations with PFOS levels were not evaluated because a previous
analysis had shown no association between PFOS and the selected outcomes. Serum PFOA
concentrations ranged from 0.01 to 92.03 ug/mL for the male workers (all sites combined), with
a mean serum PFOA concentration of'2.21, 1.02, 4.63, and 1.89 |ig/mL for all sites combined,
and the Antwerp, Cottage Grove, and Decatur sites, respectively. Serum PFOA (all sites
combined) was not associated with TC or LDL. A negative association was observed between
serum PFOA concentration (all sites combined) and HDL. Serum triglyceride was positively
associated with serum PFOA at all sites combined and independently at the Antwerp site.
Nonadherence to the fasting requirement for blood collection, especially for night-shift workers,
and potential binding of PFOA to albumin and LDL, were identified by the authors as possible
factors that influenced the triglyceride results.
Sakr et al. (2007a) conducted a cross-sectional analysis of PFOA and lipids among active
employees at the DuPont Washington Works fluoropolymer production plant in West Virginia.
The employees who volunteered to participate in the study (n = 1025, 782 males, 243 females)
each had a physical examination, provided a fasting blood sample, and answered a medical and
occupation history questionnaire in 2004. The association between PFOA and lipid levels was
evaluated by ANOVA, %2 test, student's t-test, and linear regression models. Confounders
including age, BMI, gender, alcohol consumption, and parental heart attack were considered in
the models. Mean serum PFOA concentration in the workers was 0.428 ±0.189 |ig/mL
(interquartile range 0.099-0.381). For those with current occupational exposure to PFOA, the
range was 0.0174-9.55 ug/mL and for workers with intermittent occupational exposure, the
range was 0.0081-2.07 ug/mL. The range was 0.0086-2.59 ug/mL for workers with past
occupational exposure and the 0.0046-0.963 ug/mL for workers with no occupational exposure.
Perfluorooctanoic acid (PFOA) - May 2016 3-3
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Serum PFOA was positively associated with cholesterol, very low-density lipoprotein (VLDL),
and LDL (p<0.03) in the participating workers, whether or not they were taking lipid-lowering
medication. No association was observed between serum PFOA and FIDL or triglycerides. PFOS
was not included in the study.
Costa et al. (2009) examined serum lipid data using 30 years of medical surveillance data
from workers of a PFOA production plant in Italy. The workers (n = 53 males, 20-63 years of
age) participated in the medical surveillance program yearly from 1978 to 2007. The length of
work exposure was 0.5-32.5 years. In 2007, 37 males were active workers and 16 males were
retired or had transferred to other departments and were no longer being exposed. Unexposed
male workers (n = 107, 12 executives and 95 blue collar workers) from different departments
also participated in the medical surveillance program and served as controls. Beginning in 2000,
serum PFOA was monitored yearly except in 2005. Serum PFOA concentrations in the workers
decreased after plant renovations partially automated the PFOA production process and
procedures for the use of protective devices were instituted in 2002. In 2007, the geometric mean
serum PFOA was 4.02 and 3.76 ug/mL, respectively, in currently exposed and retired workers.
Three analyses were conducted: a t-test comparing 34 exposed workers matched to 34 unexposed
workers by age, work seniority, day/shift work, and living conditions; linear regression with 34
exposed workers and 107 unexposed workers adjusting for age, work seniority, BMI, smoking,
and alcohol consumption; and a repeated measures analysis with a total of 56 individuals with
more than one measure, adjusting for age, work seniority, BMI, smoking, alcohol consumption,
and year of observation. TC and uric acid were significantly increased (p<0.05) in relation to
PFOA exposure in each of these analyses. No correlations were observed between serum PFOA
concentration and Apo-A (HDL-associated) or Apo-B (LDL-associated) proteins, HDL, or
triglycerides in any of the analyses. PFOS was not included in this study.
Three other studies included analyses with multiple measures over time (Table 3-1). Olsen et
al. (2003) conducted a longitudinal analysis of the 2000 medical surveillance data from the 3M
workers in conjunction with 1995 and 1997 data. This analysis included 175 male employees
with data from 2000 and at least one of the other survey dates. Only 41 employees were
participants in all three surveillance periods. Mean serum levels for the group sampled in 1995
and 2000 (n = 64) were 1.36 |ig/mL and 1.59 jig/mL, respectively. Mean serum levels for the
group sampled in 1997 and 2000 (n = 69) were 1.22 |ig/mL and 1.49 jig/mL, respectively.
Finally, mean serum levels for the group sampled in 1995, 1997, and 2000 (n = 41) were
1.41 |ig/mL, 1.90 |ig/mL, and 1.77 |ig/mL, respectively. When serum levels were analyzed using
a repeated measures mixed-model multivariable regression, adjusting for age, BMI, smoking,
alcohol consumption, location, year at first entry, years worked (at baseline), and years of
follow-up, there was a statistically significant positive association between PFOA and serum
cholesterol (Beta1 = 0.032, 95% CI 0.013, 0.015) and triglycerides (Beta = 0.094, 95% CI 0.045,
0.144) (p = 0.0002). PFOS levels were not associated with changes in serum lipids over time.
Sakr et al. (2007b) conducted a longitudinal analysis among the workers at the DuPont
Washington Works plant in West Virginia using data from 1979 to 2004. Employee medical
records from the medical surveillance program were used to obtain blood lipid (e.g., TC, LDL,
HDL, triglycerides), height, and weight data. As part of the medical surveillance program, each
employee gave a detailed medical history and had a physical examination at least every 3 years.
Serum PFOA concentration was measured every 1-2 years in PFOA-exposed workers and every
1 The beta coefficient from the regression analysis.
Perfluorooctanoic acid (PFOA) - May 2016 3-4
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3-5 years in non-PFOA-exposed workers on a volunteer basis. This study included 454 workers
who had two or more serum PFOA measurements. The study population included 334 males and
120 females ranging in age from 24 to 66 years who had worked at the plant for at least 1 year
since 1979. A linear mixed effects regression model was used to analyze the data and accounted
for age (and age-squared), gender, BMI, and decade of hire as potential confounders. Serum PFOA
concentrations ranged from 0 to 22.66 ug/mL, with a mean of 1.13 ug/mL over the 23-year
monitoring period in the study population. For employees with two or more PFOA measurements,
the mean of the first and last sample was 1.04 |ig/mL and 1.16 jig/mL, respectively, with an
average of 10.8 years between samples. Serum PFOA concentration was positively associated with
TC after age, BMI, gender, and decade of hire adjustment in the model (Beta = 1.06, 95% CI 0.24,
1.88) per ppm increase in PFOA. Information on lipid-lowering medications and alcohol intake by
the participants was not available. PFOS was not included in this study.
Steenland et al. (2015) conducted an analysis of the incidence of several conditions,
including high cholesterol (based on prescription medication use) among 3,713 workers at the
Washington Works plant in West Virginia who participated in the C8 Health Project. Yearly
serum estimates were modeled from work history information and job-specific concentrations.
Cox proportional hazard models, stratified by birth year, were used to assess self-reported
incidence of high cholesterol in relation to time-varying cumulative estimated PFOA serum
concentration, controlling for gender, race, education, smoking, and alcohol consumption. No
association was seen when analyzed without a lag (HRs by quartile 1.0, 1.11, 1.06, 1.05; trend
p = 0.56 for log cumulative exposure), or when using a 10-year lag (HRs by quartile 1.0, 0.93,
1.01, 0.96; trend p = 0.62).
High-exposure community studies. Several studies examined serum lipids in populations
serviced by water districts contaminated by the Washington Works PFOA production plant in
Ohio and West Virginia (Table 3-2). Emmett et al. (2006) is a small study (n = 371) with limited
analysis (t-tests comparing PFOA levels in people with abnormal versus normal TC); the larger
studies were conducted as part of the C8 Health Project. This collection of studies includes
analyses of current serum PFOA levels in relation to serum lipids in adults (Steenland et al.
2009) and children (Frisbee et al. 2010), longitudinal analysis of the change in lipids seen in
relation to a change in serum PFOA (Fitz-Simon et al. 2013), and analyses of the incidence of
hypercholesteremia in relation to modeled exposure (Winquist and Steenland 2014a). With the
exception of one set of analyses within the Winquist and Steenland study (2014a), these data
provide consistent evidence of positive associations between PFOA exposure (measured directly
in blood or modeled based on environmental and drinking water data) and TC.
Emmett et al. (2006) examined the association of serum PFOA concentration with serum TC
in residents of the Little Hocking water district in Ohio. The study population (n = 371, 2->60
years of age) was a random sample of the population served by LHWA. The subjects completed
questionnaires (e.g., demographic, occupational, health conditions, and so forth) and provided
blood samples. PFOA concentration was determined by HPLC/MS/MS; no other PFASs were
measured. Regression models were used to analyze the data. The median serum PFOA
concentration was 0.354 |ig/mL. No association was observed between serum PFOA and TC.
Steenland et al. (2009) examined the association of PFOA with serum lipids in adult
participants of the C8 Health Project (n = 46,294; 18->80 years). Serum samples were separated
into deciles or quartiles for analysis. TC, HDL, triglycerides, LDL, and non-HDL (TC minus
HDL cholesterol) were measured or calculated from blood samples. The data were analyzed by
linear regression using the log-transformed values for all variables. Covariates of the model
Perfluorooctanoic acid (PFOA) - May 2016 3-5
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included age, gender, quartile BMI, education, smoking, regular exercise, and alcohol
consumption. A logistic model was used to analyze high cholesterol and serum PFOA
concentration (quartiles). The mean serum PFOA concentration was 0.080 jig/mL. All lipid
outcomes, except for FIDL, showed significant increasing trends with increasing serum PFOA
decile. There was a positive association between mean levels of serum PFOA and TC, LDL
cholesterol, triglycerides, TC/HDL ratio, and non-HDL. The predicted increase in TC from
lowest to highest serum PFOA concentration decile was 11-12 milligrams per deciliter (mg/dL).
The odds ratio (OR) for high cholesterol (>240 mg/dL) increased from the lowest to the highest
quartile of serum PFOA concentrations: 1.00, 1.21 (95% CI: 1.12-1.31), 1.33 (95% CI: 1.23-
1.43), and 1.38 (95% CI: 1.28-1.51). No association was observed between mean level of serum
PFOA and HDL cholesterol. PFOS also was positively associated with TC, LDL cholesterol, and
triglycerides. The results of the study were consistent with occupational studies that found a
positive association between PFOA exposure and serum lipids.
The study by Frisbee et al. (2010) used a design and analysis strategy similar to that of
Steenland et al. (2009), but it was conducted among children (n = 6536; 1-11.9 years) and
adolescent (n = 5934; 12.0-17.9 years) participants of the C8 Health Project. The mean serum
PFOA concentration was 0.0777 |ig/mL and 0.0618 jig/mL, respectively, for children and
adolescents. TC, LDL, and triglycerides were positively associated (p<0.02) with serum PFOA
concentration, adjusting for age, gender, BMI, exercise, and length of fast. Assessment of the
quintile trends showed significant differences (p<0.02) between the first and fifth quintile for TC
and LDL for children and adolescents of both genders combined and separated. A significant
difference (p = 0.04) was observed for fasting triglycerides in female children only. An increased
risk of abnormal TC and LDL were positively associated with serum PFOA. The ORs were
1.0 first (reference), 1.1 (95% CI: 1.0-1.3, second), 1.2 (95% CI: 1.0-1.4, third), and 1.2 (95%
CI: 1.1-1.4, fourth and fifth) for TC, and 1.0 (reference, first), 1.2 (95% CI: 1.0-1.5, second),
1.2 (95% CI: 1.0-1.4, third and fourth), and 1.4 (95% CI: 1.2-1.7, fifth) for LDL. An increased
risk of abnormal fasting triglyceride and HDL was not associated with serum PFOA. PFOS also
was positively associated with TC, LDL cholesterol, and HDL cholesterol.
The C8 Science Panel (2012) used data from the C8 general population cohorts as well as
from combined general population and worker cohorts to evaluate the association between PFOA
and a medical diagnosis of high cholesterol. Despite inconsistent evidence between studies, they
concluded that there is a probable link between PFOA and diagnosed high cholesterol. The
worker cohort was not evaluated separately in this analysis.
A cohort of 521 members of the C8 Health Project was evaluated for an association between
changes in serum PFOA levels and changes in serum LDL-cholesterol, HDL-cholesterol, TC,
and triglycerides over a 4.4-year period (Fitz-Simon et al. 2013). Linear regression models were
fit to the logarithm (base 10) of ratio change in each serum lipid measurement in relation to the
logarithm of ratio change in PFOA. Mean serum PFOA concentration decreased by approximately
one-half between baseline (0.140 ± 0.209 |ig/mL) and follow-up (0.068 ± 0.144 |ig/mL). No
corresponding changes in serum lipids were found. However, those individuals with the greatest
declines in serum PFOA had a larger decrease in LDL cholesterol.
Perfluorooctanoic acid (PFOA) - May 2016 3-6
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Table 3-1. Summary of PFOA Occupational Exposure Studies of PFOA and Serum Lipids
Reference and Study Details
PFOA Level
TC
LDL
HDL
Triglycerides
Cross-sectional
Olsen et al. 2000
n= 111 in 1993, 80 in 1995, 74 in 1997;
50-70% participation rate
Mean age: ~ 40 yrs
Mean duration: not reported
ANOVA based on three exposure
categories, adjusted
Olsen and Zobel 2007
3M. Antwerp, Cottage Grove, Decatur
combined; 50-65% participation rate
n = 506 (men, not taking lipid-lowering
medications)
Mean age: 40 yrs
Mean duration: not reported
Linear regression, adjusted
[Related reference: Olsen et al. 2003]
Sakr et al. 2007a
Washington Works (West Virginia)
n = 1025 (782 men, 243 women), 55%
participation rate
Mean age: 46.5 and 44.4 yrs, respectively
for men and women
Mean duration: 19.6 and 15.9 yrs,
respectively for men and women
Linear regression, adjusted
Costa et al. 2009
Italy PFOA production plant
n = 37 currently exposed, 16 formerly
exposed, 107 controls (different areas in
the plant)
Mean age: 42 yrs currently exposed and
controls; 52 yrs formerly exposed
Mean duration (in 2007): 14- 16 yrs
Analysis 1 : t-test, 34 currently exposed
matched to 34 controls, adjusted
Analysis 2: Linear regression, 34 currently
exposed and 107 controls, adjusted
Analysis 3: linear regression (generalized
estimating equation [GEE] modeling), 56
total with concurrent PFOA and lipid
measure, adjusted
(l)0to< l,mean~
0.4 ng/mL
(2) 1 to< 10, mean-
3 ng/mL
(3)> 10, mean -30
Hg/mL
Mean 2. 21 |ig/mL
range 0.01-92.03
Hg/mL
Mean 0.428 ng/mL
range 0.005- 9.55
Hg/mL
Currently exposed:
mean 12.9, geometric
mean 4. 02, range
0.2^7 ng/mL
Formerly exposed:
mean 6. 81 geometric
mean 3. 76, range
0.53-18 ng/mL
1993: mean 215, 219, 232
mg/dl (p = 0.45)
1995: mean 207, 212, 221
mg/dl (p = 0.48)
1997: mean 199, 213, 217
mg/dl (p = 0.08)
Beta = 0.0076
(SE 0.0059)
(p = 0.20)
[log-transformed PFOA
and cholesterol]
All: Beta = 4.036
(SE 1. 284) (p = 0.002)
Excluding workers taking
lipid-lowering
medications:
Beta =5. 519
(SE 1.467)
p = < 0.001)
Analysis 1:
Currently exposed:
237.0 mg/dl
Controls: 206.4 mg/dl
(p = 0.003)
Analysis 2:
Beta = 2 1.7
(95% CI 6.83, 36.6)
(p = 0.005)
Analysis 3:
Beta = 0.028
(95% CI 0.002, 0.055)
(p<0.05)
1993: mean 138, 143, 140
mg/dl (p = 0.84)
1995: mean 131, 133, 130
mg/dl (p = 0.96)
1997: mean 114, 134, 134
mg/dl (p = 0.11)
Beta = 0.0021
(SE 0.0090)
(p = 0.81)
[log-transformed PFOA
and cholesterol]
All: Beta = 2.834
(SE 1. 062) (p = 0.008)
Excluding workers taking
lipid-lowering
medications:
Beta =3. 561
(SE 1.213) (p = 0.003)
Not measured
1993: mean 43, 47, 48
mg/dl (p = 0.32)
1995: mean 42, 43, 41
mg/dl (p = 0.70)
1997: mean 4 1,44, 45
mg/dl (p = 0.40)
Beta =-0.0183
(SE 0.0069)
(p = 0.01)
[log-transformed PFOA
and cholesterol]
All: Beta =-0.1 78
(SE 0.432) (p = 0.68)
Excluding workers taking
lipid-lowering
medications:
Beta = 0.023
(SE 0.058) (p = 0.96)
Analysis 1:
Currently exposed:
56.68 mg/dl
Controls: 57.82 mg/dl
(p => 0.05)
Analysis 2:
Beta = 2.42
(95% CI -2. 30, 7. 13)
(p>0.05)
Analysis 3:
Beta =-0.018
(95% CI -0.047, 0.012)
(p > 0.05)
1993: mean 171, 205,
221 mg/dl (p = 0.77)
1995: mean 152, 123,
1 83 mg/dl (p = 0.07)
1997: mean 21 9, 176,
25 1 mg/dl (p = 0.1 3)
Beta = 0.0711
(SE 0.0169)
(p = 0.0001)
[log-transformed PFOA
and cholesterol]
All: Beta =01 8
(SE 0.021) (p = 0.38)
Excluding workers
taking lipid-lowering
medications:
Beta = 0.030
(SE 0.024 (p = 0.21)
[TG log-transformed]
Analysis 1:
Currently exposed:
150.03 mg/dl
Controls: 155. 35 mg/dl
(p>0.05)
Analysis 2:
Beta=-0. 15
(95% CI -34.6, 34.3)
(p> 0.005)
Analysis 3:
Beta = 0.055
(95% CI -0.036, 0.147)
(p>0.05)
Perfluorooctanoic acid (PFOA) - May 2016
3-7
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Reference and Study Details
PFOA Level
TC
LDL
HDL
Triglycerides
Longitudinal
Olsen et al. 2003
3M, Antwerp and Decatur combined
~5 yr follow-up period
n = 174 (measure in 1995 or 1997, and in
2000)
Mean age: not reported
Mean duration: not reported
Linear mixed effects regression for
repeated measures, adjusted
Sakr et al. 2007b
Washington Works (West Virginia)
n = 454
23-yr follow-up (mean 3.7 PFOA
measures)
Mean age: 27 yrs (at hire)
Mean duration: 27 yrs
Linear mixed effects regression for
repeated measures, adjusted
Steenland et al. 2015
n=3,713 workers
Data collected in 2005-2006 and 2008-
2011
n= 1,298 cases
1995 baseline: 1.36-
1.41 |ig/mL
2000 follow-up:
1.49-1. 77 ng/mL
1.04 ng/mL (first)
1.16 |ig/mL(last)
Declined since 1980
(mean 4.78 |ig/mL in
1980tol.OO|ig/mL
in 2001-2004)
In 2005-2006: mean
0.325 |ig/mL, median
0.113|ig/mL
Beta = 0.032
(95% CI 0.013, 0.051)
[and statistically
significant PFOA-years
follow-up interaction,
Beta = -0.0004]
Beta= 1.06
(95% CI 0.24, 1.88)
Not measured
Beta = 0.46
(95% CI -0.87, 1.79)
Reported as "no
significant changes"
Beta =0.16
(95% CI -0.39, 0.71)
Beta = 0.094
(95% CI 0.045, 0.144)
Beta = 0.79
(95% CI -5.99, 7.57)
HR (95% CI), for self-reported use of cholesterol-lowering medications (incidence based on year of diagnosis).
Cumulative exposure quartile, no lags:
1.00 (referent)
1.11(0.94,1.30)
1.06(0.89,1.27)
1.05(0.87,1.27) (P trend =Q. 56)
Perfluorooctanoic acid (PFOA) - May 2016
3-S
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Table 3-2. Summary of High-Exposure Community Studies of PFOA and Serum Lipids
Reference and Study Details
PFOA level
TC
LDL
HDL
Triglycerides
Cross-sectional
Emmett et al. 2006
n = 371, aged 2-89 yrs, median 50 yrs
(317 from stratified random population sample)
Linear regression (continuous PFOA) and t-test,
PFOA, abnormal versus normal cholesterol
Steenland et al. 2009
n = 46,294, aged 18-80 yrs (not taking
cholesterol-lowering medications)
Linear regression, quartiles PFOA and continuous
PFOA
Frisbeeetal. 2010
6,536 children l-< 12 yrs
5,934 adolescents, 12-18 yrs
Linear regression, adjusted
0.354 ug/mL
Mean 0.08
ug/mL
Quartiles
0-0.0131
0.0132-0.0265
0.0266-0.0669
> 0.067
Mean ug/mL
Children 0.0777
Adolescents
0.0618
Regression: Beta =
0.0055 l(p = 0.27)
p- value oft-test = 0.79
[n=182,49%
abnormal]
Beta = 0.01 112
(SE 0.00076)
[log PFOA and lipids]
By quartiles (OR):
1.0 (referent)
1.21(1.12,1.31)
1.33(1.23,1.43)
1.38(1.28,1.50)
Difference between 1st
and 5th quintile PFOA
(trend p):
Children 5. 8 mg/dl
(p< 0.0001)
Adolescents 4.2 mg/dl
(p< 0.0001)
Not measured
Beta = 0.01499
(SE 0.00121)
[log PFOA and lipids]
Difference between 1st
and 5th quintile PFOA
(trend p):
Children 4. 9 mg/dl
(p = 0.001)
Adolescents 3.2mg/dl
(p = 0.004)
Not measured
Beta = 0.00276
(SE 0.00094)
[log PFOA and lipids]
Difference between 1st
and 5th quintile PFOA
(trend p):
Children 5. 8 mg/dl
(p = 0.88)
Adolescents 4.2 mg/dl
(p = 0.20)
Not measured
Beta =0.00169
(SE 0.00219)
[log PFOA and lipids]
Difference between 1st
and 5th quintile PFOA
(trend p):
Children 2.0 mg/dl
(p = 0.10)
Adolescents 3.8 mg/dl
(p = 0.10)
Longitudinal (Change in Lipid in Relation to Change in PFOA)
Fitz-Simon et al. 2013
Longitudinal; 4.4 yrs n = 521
Linear regression of log of ratio change in serum
lipid to log of ratio change in PFOA, adjusted for
age, gender, interval between measures, fasting
status (change in lipid in relation to change in
PFOA)
0.140 ug/mL
(baseline)
0.068 ug/mL
(follow-up)
Percent decrease (95%
CI) in lipid per halving
PFOA:
1.65(0.32,2.97);
with additional
adjustment for PFOS:
0.63 (-0.88, 2. 12)
Percent decrease (95%
CI) in lipid per halving
PFOA:
3.58(1.47,5.66);
with additional
adjustment for PFOS:
2.92(0.71,5.09)
Percent decrease (95%
CI) in lipid per halving
PFOA:
1.33 (-0.21, 2.85);
with additional
adjustment for PFOS:
1.24 (-0.34, 2.79)
Percent decrease (95%
CI) in lipid per halving
PFOA:
-0.78 (-5.34, 3.58);
with additional
adjustment for PFOS:
-1.16 (-5. 85, 3.33)
Incidence of Hypercholesterolemia
Winquist and Steenland 2014a
n = 32,254 (including 3,713 workers)
Data collected in 2005-2006 and 2008-201 1
n = 9,653 cases in primary analysis (all diagnoses)
n = 1,825 cases in prospective analysis (diagnoses
after 2005-2006)
In 2005-2006:
mean 0.0866
ug/mL, median
0.0261 ug/mL
HR (95% CI), self-reported use of cholesterol-lowering medications, primary analysis
Cumulative exposure quintiles: Year exposure quintiles:
1.00 (referent) 1.00 (referent)
1.24 (1.15,1.33) 1.07 (1.01,1.15)
1.17 (1.09,1.26) 1.11 (1.04,1.19)
1.19 (1.11,1.27) 1.05 (0.99,1.13)
1.19 (1.11, 1.28) (Ptrmd= 0.005) 1.20 (1.12, 1.28) (P trend = 0.001)
Diagnoses after 2005: no association with PFOA with either exposure metric
Perfluorooctanoic acid (PFOA) - May 2016
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More recently, participants in the C8 Health Project were examined for an association
between PFOA levels and incidence of several conditions, including high cholesterol (based on
prescription medication use) (Winquist and Steenland 2014a). The cohort included 28,541
community members and 3,713 workers who had completed study questionnaires during 2008-
2011. The median serum PFOA level at enrollment in 2005-2006 was 0.0261 |ig/mL for the
combined cohort, 0.0242 |ig/mL for the community members, and 0.1127 |ig/mL for the
workers. Retrospective serum levels for the community cohort were estimated from air and water
concentrations, residential history, and water consumption rates. For the workers, yearly serum
estimates were modeled from work history information and job-specific concentrations. Cox
proportional hazard models, stratified by birth year, were used to assess self-reported adult heart
disease hazard in relation to time-varying yearly or cumulative (sum of yearly estimates)
estimated PFOA serum concentration, controlling for gender, race, education, smoking, and
alcohol consumption. Using the cumulative exposure metric, the FIRs for hypercholesterolemia
for quintiles 2-5 versus quintile 1 were 1.24, 1.17, 1.19, and 1.19 (Ptrend = 0.005). Using the
yearly exposure metric, the FIRs for high cholesterol for quintiles 2-5 versus quintile 1 were
1.07, 1.11, 1.05, and 1.20 (Ptrend = 0.001). The strongest association was in males aged 40-59.
No associations were found between PFOA level and hypertension or coronary artery disease
incidence. (The analysis of these data restricted to the worker population by Steenland et al.
[2015] is described in the previous section).
A subset of 290 individuals in the C8 Health Project was evaluated for evidence that PFOA
exposure can influence the transcript expression of genes involved in cholesterol metabolism,
mobilization, or transport (Fletcher et al. 2013). RNA was extracted from whole blood samples
taken from 144 males and 146 females aged 20-60 years; serum collected at the same time was
used to measure PFOA concentration. The association between candidate gene expression levels
and PFOA levels was assessed by multivariable linear regression with adjustments for
confounders. Inverse associations were found between PFOA levels and expressions of
transcripts involved in cholesterol transport (NR1H2, NPC1, and ABCG1; p = 0.002, 0.026, and
0.014, respectively). When genders were analyzed separately, PFOA was negatively associated
with expression of genes involved in cholesterol transport in males (NPC1, ABCG1, PPARa)
and females (NCEH1). Similar associations were found with PFOS.
General population studies. Several studies examined serum lipids in the general population
(Table 3-3). Nelson et al. (2010) examined the relationship between polyfluoroalkyl chemical
serum concentration, including PFOA, and lipid and weight outcomes in the general population
of the United States by analyzing data from the 2003-2004 NHANES. The population (n = 860)
included persons aged 20-80 years with no missing covariate information who were not
pregnant, breast-feeding, taking insulin or cholesterol medicine, or undergoing dialysis.
Cholesterol (TC, HDL, LDL) was measured from serum samples. Data for covariates predicting
cholesterol and body weight including age, gender, race/ethnicity, socioeconomic status,
saturated fat intake, exercise, alcohol consumption at > 20 years of age, smoking, and parity
were obtained from the questionnaires. Regression analyses were performed for gender and the
age groups 12-19 years, 20-59 years, and 60-80 years. The mean PFOA concentration was
0.0046 ± 0.003 |ig/mL. A positive association was found between TC and non-HDL (TC-HDL,
-70-80% TC) cholesterol and serum PFOA (effect estimate 9.8; 95% CI, -0.2-19.7). No
association was found between serum PFOA concentration and HDL, or LDL. No association
was found between serum PFOA concentration and body weight. Similar results were found
with PFOS. A similar analysis using 1999-2008 NHANES data for 815 adolescents (aged 12-18
years) by Geiger et al. (2014a) found an association between serum PFOA and TC
Perfluorooctanoic acid (PFOA) - May 2016 3-10
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(Beta 4.55, 95% CI 0.90, 8.20, per In-unit increase in PFOA) and LDL (Beta 5.75, 95% CI 2.16,
9.33, per In-unit increase in PFOA).
Eriksen et al. (2013) examined the association between plasma PFOA (and PFOS) levels and
TC levels in a middle-aged Danish population. This cross-sectional study included 663 males
and 90 females aged 50-65 years who were enrolled in the Danish Diet, Cancer and Health
cohort. Generalized linear models were used to analyze the association between PFOA and TC
levels, adjusted for age, gender, education, BMI, smoking, alcohol consumption, egg intake,
animal fat intake, and physical activity. The mean plasma PFOA level was 0.0071 |ig/mL. A
significant, positive association was found between PFOA (and PFOS) and TC such that, in the
fully adjusted model, a 4.4-mg/dL (95% CI 0.8, 8.5) higher concentration of TC was found per
interquartile range of plasma PFOA (quartile cut-points were not reported).
Fisher et al. (2013) examined the association of plasma PFAS levels, including PFOA, with
metabolic function and plasma lipid levels. This population-based sample included 2,700
participants aged 18-74 years (-50% male) in the Canadian Health Measures Survey. The
geometric mean PFOA concentration was 0.0025 ± 0.0018 jig/mL. In analyses that included
sampling weights, no associations were found between PFOA (or PFOS) and TC, HDL- and
LDL-cholesterol, and metabolic syndrome and glucose homeostasis parameters. Covariates
considered included age, gender, marital status, income adequacy, race, education, BMI, physical
activity, smoking, and alcohol consumption.
Starling et al. (2014) examined the association between PFOA (and six other PFASs) and
serum lipids in pregnant females in the Norwegian Mother and Child Cohort Study. Most of the
blood samples were drawn during weeks 14-26 of gestation. Weighted multiple linear regression
was used to estimate the association between PFOA level and each lipid level. Covariates
considered included age, prepregnancy BMI, nulliparous or interpregnancy interval, breast-
feeding duration, education, current smoking, gestation week at blood draw, oily fish
consumption, and weight gain during pregnancy. The median plasma PFOA level was
0.00225 |ig/mL. No association was observed between PFOA and triglycerides, TC, or LDL-
cholesterol. PFOA was positively associated with HDL-cholesterol, although the CI was large
for the association. With HDL-cholesterol, each interquartile range- (IQR-) unit increase in In-
PFOA was associated with an increase of 1.28 mg/dL (95% CI: -0.15, 2.71). Five of the seven
PFASs studied were positively associated with HDL cholesterol and all seven had elevated HDL
associated with the highest quartile.
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Table 3-3. Summary of General Population Epidemiology Studies of PFOA with Serum Lipids
Reference and Study Details
PFOA level
TC
LDL
HDL
Triglycerides
AllAduhs
Nelson etal. 2010
United States, NHANES (2003-2004)
n = 860, aged 20-80 yrs
(451 men, 409 women)
Linear regression, PFOA in quartiles and
continuous PFOA, adjusted
PFOA-PFOS correlation Spearman r = 0.65
Eriksen etal. 2013
Denmark
n = 753, aged 50-65 yrs
(663 men, 90 women)
Linear regression, continuous PFOA, adjusted
PFOA-PFOS correlation not reported
Fisher etal. 2013
Canada, Canadian Health Measures Survey
n = 2,700, aged 18-74 yrs
Linear regression, continuous PFOA (log-
transformed PFOA and lipids)
PFOS correlation r = 0.36
Mean 0.0046 |ig/mL,
median 0.0038 |ig/mL
Mean 0.0071 |ig/mL
Mean 0.0025 |ig/mL
Quartiles:
0.00015-0.00185
0.00186-0.00258
0.00259-0.00355
> 0.0036
Beta =1.22
(95% CI 0.04, 2.40)
9.8 mg/dl increase in top
versus bottom quartile
(PFOS results similar)
4.4 mg/dl increase per
IQR
(PFOS results similar)
Beta = 0.03
(95% CI -0.017, 0.07)
[log PFOA and lipids]
Beta =-0.21
(95% CI- 1.9 1,1.49)
(PFOS results similar)
Not measured
Beta = 0.02
(95% CI -0.06, 0.091)
[log PFOA and lipids]
Beta =-0.12
(95% CI -0.41, 0.16)
Different pattern seen with
PFOS
Not measured
Beta = 0.0009
(95% CI -0.04, 0.04)
[log PFOA and lipids]
Not measured
Not measured
Not measured
Pregnant Women
Starling etal. 2014
Norway
n = 891 pregnant women
Plasma PFOA (collected in 2nd trimester)
Linear regression, continuous PFOA (log-
transformed PFOA), adjusted
PFOS correlation Spearman r = 0.64
Median 0.00225
Hg/mL
Beta = 2. 58
(95% CI -4. 32, 9.47)
[per In-unit increase in
PFOA]
Beta = 2.25
(95% CI -3. 97, 8.48)
[per In-unit increase in
PFOA]
Beta = 2.13
(95% CI -0.26, 4.51)
[per In-unit increase in
PFOA]
Beta = 0.00
(95% CI -0.07,
0.06)
[per In-unit increase
in PFOA]
Adolescents
Geiger et al. 2014a
United States, NHANES (1999-2008)
n= 815, aged 12-18 yrs
Linear regression, continuous PFOA (log-
transformed PFOA), adjusted
Mean 0.0042 |ig/mL
Beta = 4. 55
(95% CI 0.90, 8.20)
[per In-unit increase in
PFOA]
(PFOS results similar)
Beta =5. 75
(95% CI 2.16, 9.33)
[per In-unit increase in
PFOA]
(PFOS results similar)
Beta =-1.52
(95% CI -3.02, -0.03
[per In-unit increase in
PFOA]
Attenuated results when
adjusted for PFOS
Beta= 1.74
(95% CI -2.88,
6.36)
Different pattern
seen with PFOS
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The association between PFOA and serum lipids has been examined in several studies in
different populations. Cross-sectional and longitudinal studies in occupational settings (Costa et
al. 2009; Olsen et al. 2000, 2003; Olsen and Zobel 2007; Sakr et al. 2007a, 2007b; Steenland et
al. 2015) and in the high-exposure community (the C8 Health Project study population) (Fitz-
Simon et al. 2013; Frisbee et al 2010; Steenland et al. 2009; Winquist and Steenland 2014a)
generally observed positive associations between serum PFOA and TC in adults and children
(aged l-< 18 yrs); most of the effect estimates were statistically significant. Although exceptions
to this pattern are present (i.e., some of the analyses examining incidence of self-reported high
cholesterol based on medication use in Winquist and Steenland [2014a] and in Steenland et al.
[2015]), the results are relatively consistent and robust. Similar associations were seen in
analyses of LDL, but were not seen with HDL. The range of exposure in occupational studies is
large (with means varying between 0.4 and > 12 jig/mL), and the mean serum levels in the C8
population studies were around 0.08 jig/mL. Positive associations between serum PFOA and TC
(i.e., increasing lipid level with increasing PFOA) were observed in most of the general
population studies at mean exposure levels of 0.002-0.007 |ig/mL (Eriksen et al. 2013; Fisher et
al. 2013; Geiger et al. 2014a; Nelson et al. 2010; Starling et al. 2014). The interpretation of these
general population results is limited, however, by the moderately strong correlations (Spearman
r > 0.6) and similarity in results seen for PFOS and PFOA.
3.1.1.2 Cardiovascular Diseases
Occupational exposure studies. Several studies examined cardiovascular-related cause of death
among PFOA-exposed workers at the West Virginia Washington Works plant (Leonard et al.
2008; Sakr et al. 2009; Steenland and Woskie 2012) and the 3M Cottage Grove plant in
Minnesota (Lundin et al. 2009; Gilliland and Mandel 1993). This type of mortality is of interest
because of the relation between lipid profiles (e.g., LDL) and the risk of cardiovascular disease.
The most recent West Virginia study included 5,791 individuals who had worked at the plant for
at least 1 year between 1948 and 2002, with mortality follow-up through 2008. No associations
were found between cumulative PFOA levels and ischemic heart disease (IHD) mortality
(standardized mortality ratio [SMR] 1.07, 1.02, 0.87, and 0.93 across four quartiles of cumulative
exposure, compared to U.S. referent group). Based on these data from the worker cohorts, the C8
Science Panel (2012) concluded that there is no probable link between PFOA and stroke and
coronary artery disease.
The analysis of the Minnesota plant (n = 3,993 workers who began work between 1983 and
1997, with follow-up through 2002) also found no association between cumulative PFOA
exposure and IHD risk, but an increased risk of cerebrovascular disease mortality was seen in the
highest exposure category (HR 2.1, 95% CI 1.0, 4.6). These studies are limited by the reliance on
mortality (rather than incidence) data, which can result in a substantial degree of under
ascertainment and misclassification.
3.1.1.3 Liver Enzymes and Liver Disease
Cross-sectional studies and longitudinal studies of PFOA and liver enzymes in various
populations are described in this section and summarized in Table 3-4.
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Table 3-4. Summary of Epidemiology Studies of PFOA and Liver Enzymes
Reference and Study Details
PFOA level
Results
Cross-sectional: Occupational Exposure Studies
Olsen et al. 2000
n= 111 in 1993, 80 in 1995, 74 in 1997; 50-70%
participation rate
Mean age: ~ 40 yrs
Mean duration: not reported
ANOVA and linear regression adjusted for age,
BMI, and alcohol and cigarette use
Mean (range)
1993: 5 (0-80)
1995:6.8(0-114)
1997:6.4(0.1-81)
ALT: Year p (±SE) (p value)
1993: 0.89 (2.88) (p = 0.76)
1995: 0.81 (2.62) (p = 0.75)
1997: 2.77(1.27) (p = 0.03)
Change per 10 |ig/mL increase in serum PFOA; stronger association in individuals with
BMI< 30. No associations for ALP, GOT, AST, total bilirubin, direct bilirubin.
Olsen and Zobel 2007
3M. Antwerp, Cottage Grove, Decatur combined;
50-65% participation rate
n = 506 (men, not taking lipid-lowering medications)
Mean age: 40 yrs
Mean duration: not reported
Linear regression adjusting for hi age, hi BMI, hi
alcohol
[Related reference: Olsen et al. 2003)
Mean (range)
2.21(0.01-92.03)
P change per In PFOA (±SE) p value
In ALP: All 0.009 (± 0.008) (p = 0.25)
In AST: All -0.005 (± 0.009) (p = 0.55)
In ALT: All 0.025 (± 0.013) (p = 0.06)
In GGT: All 0.033 (± 0.017) (p = 0.05)
Decatur: 0.08 (0.34) (p = 0.02)
Decatur: 0.011 (0.02) (p = 0.57)
Decatur: 0.08 (0.034) (p = 0.02)
Decatur: 0.08 (0.034) (p = 0.02)
In total bilirubin: All -0.033 (± 0.01) (p = 0.001)Decatur: -0.054 (± 0.021) (p = 0.01)
Replacement of In BMI with triglycerides in the model resulted in reduced associations for
ALT and GGT.
Sakr et al. 2007a
Washington Works (West Virginia)
n = 1025 (782 men, 243 women), 55% participation
rate
Mean age: 46.5 and 44.4 yrs, respectively for men
and women
Mean duration: 19.6 and 15.9 yrs, respectively for
men and women
Linear regression, adjusting for age, BMI, alcohol
consumption, gender, history of heart attack in
parent, use of lipid-lowering medications
0.428 jig/ml
LOQ 0.0005 jig/ml
range 0.005 - 9.55
P (±SE) p value: Full sample
In AST: 0.012 (± 0.012) (p = 0.317)
In ALT: 0.023 (± 0.015) (p = 0.124)
In GGT: 0.048 (± 0.02) (p = 0.016)
Excluding 178 men on lipid-lowering medications
In AST: 0.023 (±0.013) (p = 0.079)
In ALT: 0.031 (±0.017) (p = 0.071)
In GGT: 0.05 (±0.023) (p = 0.03)
In bilirubin: 0.008 (± 0.014) (p = 0.59) In bilirubin: 0.1 (±0.017) (p = 0.637)
Costa et al. 2009
Italy
Cross-sectional
56 male workers (currently and formerly exposed
and unexposed) with concurrent serum PFOA and
clinical parameters measured in last 7 yrs
GEE models adjusting for age, years of exposure,
year of PFOA sampling, BMI, smoking, and alcohol
consumption
Currently exposed:
mean 12.9, geometric
mean 4.02, range 0.2-47
Formerly exposed:
mean 6.81 geometric
mean 3.76, range 0.53-
18
P change per |ig PFOA/mL (95% CI)
AST: 0.038 (-0.003, 0.080)
ALT: 0.116 (0.054, 0.177)
GGT: 0.177 (0.076, 0.278)
ALP: 0.057 (0.007, 0.107)
Total bilirubin: -0.080 (-0.137, -0.024)
Conj. bilirubin: -0.034 (-0.09, 0.031)
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Reference and Study Details
PFOA level
Results
Longitudinal: Occupational Exposure Studies
Olsen et al. 2003
3M, Antwerp and Decatur combined
~5 yr follow-up period
n = 174 (measure in 1995 or 1997, and in 2000)
Mean age: not reported
Mean duration: not reported
Linear mixed effects regression for repeated
measures, adjusted
Sakr et al. 2007b
Washington Works (West Virginia)
n = 454
23-yr follow-up (mean 3.7 PFOA measures)
Mean age: 27 yrs (at hire)
Mean duration: 27 yrs
Linear mixed effects regression for repeated
measures, adjusted
1995 baseline:
1.36-1.41 ug/mL
2000 follow-up: 1.49-
1.77 ug/mL
1.04 ug/mL (first)
1.16 ug/mL (last)
Used PFOA
measurement from same
year as biomarker test or
interpolated using two
surrounding values
No associations observed; however, data not provided
P IU/L change per 1 ug/mL PFOA (95% CI)
ALP: (n = 1327) -0.21 (-0.60, 0.18)
AST: (n= 1326) 0.35 (0.10, 0.60)
ALT: (n = 231) 0.54 (-0.46, 1.54)
GOT: (n = 233) 1.24 (-1.09, 3.57)
Total bilirubin: (n= 1327) 0.008 (-0.0139, -0.0021)
Cross-sectional: High-Exposure Community Studies
Emmett et al. 2006
n= 371, aged 2-89 yrs, median 50 yrs
(317 from stratified random population sample)
Linear regression (continuous PFOA) and t-test,
PFOA, abnormal versus normal enzyme levels
Galloetal. 2012
West Virginia, United States; C8 Health Project,
46,452 of 56,554 (82.1%) adults
Adjusting for age, gender, physical activity, BMI,
average household income, educational level, fasting
status, month of blood sample collection, race,
insulin resistance, alcohol consumption, and cigarette
smoking
Median 0.354 ug/mL
(IQR 0.1 84 -0.571
ug/mL); nonfasting
blood sample
Median 0.028 ug/mL
(IQR 0.1 35 -0.71
ug/mL) nonfasting
blood sample; LOD
0.0005 ug/mL, n = 32
below LOD
Linear regression, Beta (p- value) n (%) abnormal,(t-test p- value)
ALP: -0.00416 (p = 0.65) 6 (2%) (p = 0.63)
AST: -0.0007586 (p = 0.76) 9 (2%) (p = 0.03)
ALT: -0.00183 (p=0.65) 28 (8%) (p= 0.30)
GGT: 0.00057711 (p = 0.89) 11 (3%) (p= 0.50)
Linear regression, Logistic regression of abnormal values
P per 1 unit increase PFOA (95% CI) OR (95% CI)
Ln ALT: 0.022 (0.01 8 -0.025) ALT: 1.10(1.07,1.13)
In GGT: 0.015 (0.01 -0.019) GGT: 1.01(0.99,1.04)
n Direct (conjugated) bilirubin: 0.001 (-0.002 - 0.004) Direct bilirubin 0.97 (0.90, 1 .05)
Analysis of Ln ALT or Ln GGT by decile showed increase from 0.005 to 0.030 ug/mL, then
leveling; p value for trend < 0.001; Direct bilirubin showed a U-shaped relation increasing to
0.030 ug/mL, then declining.
Cross-sectional: General Population Studies
Lin etal. 2010
United States, NHANES (1999-2000; 2003-2004)
1,076 men, 1,140 women of 10,224 enrolled,
excluding < 6-hr fast n = 1 ,802), and missing
covariate or no serum PFOA or liver function of
metabolic syndrome data
Adjusting for age, gender, race/ethnicity, smoking,
alcohol consumption, education level, BMI,
metabolic syndrome, and iron saturation status
Geometric mean
0.00505 ug/mL
0.00406 ug/mL; 0.4%
of samples below LOD
(LOD 0.0002 and
0.0001 ug/mL in 1999-
2000 and 2003-2004)
P per unit (ng/ml) increase in log serum PFOA (95% CI)
ALT:(U/1) 1.86(1.24,2.48)
Log GGT: (U/l) 0.08 (0.05, 0.11)
P per unit (ng/ml) increase in log serum PFOA (95% CI); Same model as above, also
controlling for other PFASs
ALT: (U/l) 2.19(1.4, 2.98)
Log GGT: (U/l) 0.1 5 (0.11, 0.19)
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Liver Enzymes
Occupational exposure studies. Olsen et al. (2000) analyzed alkaline phosphatase (ALP), GOT,
aspartate aminotransferase (AST), ALT, and total- and direct bilirubin data from voluntary
medical surveillance examinations of PFOA production workers at a 3M plant in 1993, 1995,
and 1997. No association was observed between serum PFOA concentration and the parameters
explored in cross-sectional analyses in the workers in 1993 or 1995; although in 1997 increases
in AST per unit increase in serum PFOA concentration were observed. When measurements for
all years were combined in longitudinal analyses (Olsen et al. 2003), the authors reported that no
associations were observed with serum PFOA levels. Other than the analyses of AST, however,
quantitative results were not provided.
A subsequent analysis involving these fluorochemical workers and an additional plant
(Cottage Grove, Minnesota) that used medical surveillance data collected in 2000 examined the
association between serum PFOA concentration and liver enzymes (Olsen and Zobel 2007).
Serum samples were analyzed for ALP, GGT, AST, ALT, and bilirubin concentrations. Ln
serum PFOA was marginally associated with In ALT and In GGT in regression models adjusting
for In age, In BMI, and In alcohol consumption, although the association was reduced when In
triglycerides replaced In BMI in the model. An inverse association between total bilirubin and
serum PFOA concentration (p<0.05) was observed at all sites combined.
Sakr et al. (2007a) examined the relationship between serum PFOA and several clinical
chemistry parameters in workers at the Washington Works plant in West Virginia. A complete
blood count, metabolic panel (glucose, blood urea nitrogen [BUN], creatinine, iron, uric acid,
electrolytes, creatinine kinase, lactic dehydrogenase [LDH], ALP, protein, albumin, C-reactive
protein), liver enzyme panel (AST, ALT, GGT, bilirubin), and serum PFOA concentration were
determined from the blood samples. Serum PFOA was associated (p<0.05) with increasing GGT
in all of the participating workers. It was stated that an association was observed between serum
PFOA concentration and iron, LDH, calcium, and potassium, but quantitative results were not
included and the direction of association was not specified.
Costa et al. (2009) also examined associations between serum PFOA concentration and liver
enzymes in workers at a fluorochemical production plant in Italy. Serum PFOA concentration
was associated with increasing ALT, GGT, and ALP levels (p<0.05), and inversely associated
with total bilirubin (p<0.01) in 56 workers with PFOA and liver enzymes measured concurrently
over the last 7 years. This subset of 56 workers included currently, formerly, and never exposed.
Sakr et al. (2007b) also conducted a longitudinal study of liver enzymes among workers at
the Washington Works plant with two or more PFOA measurements as described previously.
Hepatic clinical chemistry (GGT, AST, ALT, ALP, total bilirubin), height, and weight data were
analyzed. Serum PFOA concentration was associated in the model with increasing AST levels
(p = 0.009) and inversely associated with total bilirubin (p = 0.006) after adjustment for age,
BMI, gender, and decade of hire. No association was observed between serum PFOA
concentration and GGT, ALT, and ALP. The regression models did not adjust for alcohol
consumption, a potential limitation.
High-exposure community studies. A small study (n = 371) of residents of the Little Hocking
water district in Ohio found inconsistent results in different analyses of liver enzymes: An
association with AST but not with ALP or ALT was seen when comparing serum PFOA levels
between groups with abnormal compared to normal enzyme levels, but no association with any
Perfluorooctanoic acid (PFOA) - May 2016 3-16
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enzyme was seen in regression analyses with PFOA as a continuous variable (Emmett et al.
2006). A subsequent study, which included a wider set of communities in the contaminated area,
investigated the correlation between serum PFOA levels and liver enzymes in a total of 47,092
samples collected from members enrolled in the C8 Health Project (Gallo et al. 2012). The
association of ALT, GGT, and direct bilirubin with PFOA was assessed using linear regression
models adjusted for various confounders. The median PFOA level was 0.028 |ig/mL. The In-
transformed values of ALT were significantly associated with In-PFOA (and PFOS). There was a
steady increase in fitted levels of ALT per decile of PFOA, leveling off after approximately
0.030 jig PFOA/mL. Fitted values of GGT by deciles of PFOA showed a slight positive trend
when adjusted for insulin resistance and BMI, but this was not confirmed in the logistic model
analysis of elevated enzyme levels. Direct bilirubin levels appeared to increase at lower
concentrations and then decline in a U-shaped pattern at 0.030 jig PFOA/mL.
General population studies. Lin et al. (2010) investigated the association between serum PFOA
(plus three other PFASs) and liver enzymes in the adult population of the United States by
analyzing data from the 1999-2000 and 2003-2004 NHANES. The study population included
2,216 adults (1076 males, 1140 females) older than 20 years who were not pregnant or nursing;
had fasted more than 6 hours at the time of examination; were negative for hepatitis B or C virus;
had body weight, height, educational attainment, and smoking status data available; and had serum
tests for PFAS, liver function, or the five physiological measures associated with metabolic
syndrome. Regression models were used to analyze the data and adjust for confounders. Mean
PFOA levels were 0.00505 |ig/mL and 0.00406 |ig/mL for males and females, respectively. Serum
PFOA concentration was divided into quartiles (Ql = < 0.0029; Q2 = < 0.0042; Q3 = < 0.00595;
Q4 = > 0.00595 jig/ml). In the univariate regression models, 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 (1) age, gender, and race/ethnicity; (2) age, gender,
race/ethnicity, and lifestyle (smoking status, drinking status, education level), and (3) additional
data for BMI, metabolic syndrome biomarkers, iron saturation status, and insulin resistance. An
association was found between serum log-PFOA concentration and increasing serum ALT and log-
GGT. One unit increase in serum log-PFOA was associated with an increase of 1.86 units in serum
ALT measurements and a 0.08-unit increase in log-GGT measurements. Effect modification was
seen: For example, stronger associations between serum PFOA concentration and serum ALT (or
GGT) were found among non-Hispanic Caucasians. PFOS also was positively associated with
ALT in the fully adjusted model.
The results of the occupational studies provide evidence of an association with increases in
serum AST, ALT, and GGT, with the most consistent results seen for ALT. The associations
were not large and they might depend on the covariates in the models such as BMI, use of lipid-
lowering medications, and triglycerides (Costa et al. 2009; Olsen et al. 2000, 2003; Olsen and
Zobel 2007; Sakr et al. 2007a, 2007b). Two population-based studies of highly exposed residents
in contaminated regions near a fluorochemical industry in West Virginia have evaluated
associations with liver enzymes, and the larger of the two studies reported associations of
increasing serum In ALT and In GGT levels with increasing serum PFOA concentrations
(Emmett et al. 2006; Gallo et al. 2012). A cross-sectional analysis of data from NHANES,
representative of the U.S. national population, also found associations with In PFOA
concentration with increasing serum ALT and In GGT levels. Serum bilirubin was inversely
associated with serum PFOA in the occupational studies. A U-shaped exposure-response pattern
for serum bilirubin was observed among the participants in the C8 Health Project, which might
explain the inverse associations reported for occupational cohorts. Overall, an association of
Perfluorooctanoic acid (PFOA) - May 2016 3-17
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serum PFOA concentration with elevations in serum levels of ALT and GGT has been
consistently observed in occupational and highly exposed residential communities, and the U.S.
general population. The associations are not large in magnitude, but indicate the potential of
PFOA to affect liver function.
Liver Diseases
High-exposure community studies. Few studies of the relationship between PFOA and liver
disease are available, but the C8 Health Project did not observe associations with hepatitis, fatty
liver disease, or other types of liver disease in their initial studies. The most recent update of
disease incidence in the workers identified 35 cases of nonhepatitis liver disease (with medical
validation) (Steenland et al. 2015); no association was seen with cumulative exposure when
analyzed without a lag (FIRs by quartile 1.0, 0.58, 1.43, 1.20; trend p = 0.86 for log cumulative
exposure), but there was a possible trend in the analysis using a 10-year lag (FIRs by quartile 1.0,
1.46, 2.13, and 2.02; trend p = 0.40).
3.1.1.4 Biomarkers of Kidney Function and Kidney Disease
Kidney Function
PFOA has the potential to affect the kidney's function of tubular resorption because of it uses
tubular transporters for excretion and resorption (see section 2.4). Since PFOA is removed from
the blood by the kidney, cross-sectional analyses using serum PFOA as the exposure measure are
problematic if individuals with compromised kidney function are included: PFOA concentrations
could be increased in those individuals and an apparent association with GFR would be
observed, even if one did not exist. Studies examining measures of kidney function are described
in this section and summarized in Table 3-5.
Table 3-5. Summary of Epidemiology Studies of PFOA and Measures of Kidney Function
Reference and Study Details
PFOA Level
Results
Sakr et al. 2007a
Washington Works plant
Cross-sectional; all active,
nonpregnant employees
enrolled over 12 days in 2004
1,025 of 1,863 eligible (55%)
0.428 ug/ml
LOQ 0.0005 ug/ml range
0.005-9.55 ug/mL
Reported association with uric acid but quantified
results were not provided
Costa et al. 2009
Italy
Cross-sectional
56 male workers (currently and
formerly exposed and
unexposed) with concurrent
serum PFOA and clinical
parameters measured in last 7
Currently exposed:
mean 12.9, geometric
mean 4.02, range 0.2-47
ug/mL
Formerly exposed:
mean 6.81 geometric mean
3.76, range 0.53-18 ug/mL
P change per ug PFOA/mL (95% CI)
Uric acid 0.026 (0.001, 0.053)
GEE models adjusting for age, years of exposure,
year of PFOA sampling, BMI, smoking, and
alcohol consumption
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Reference and Study Details
PFOA Level
Results
Steenlandetal. 2010
C8 Health Project, West
Virginia
Cross-section; adult subjects (n
= 53,458; 20->80 years of age)
participating in the C8 Health
Project from 2005-2006
Subjects had consumed water
for at least 1 year prior to 2004
0.0864 ug/mL, measured
in 2005-2006
Increased predicted uric acid of 0.2-0.3 ug/dL with
increasing deciles of PFOA orPFOS
Shankaretal. 2011
United States, NHANES
Uric acid analysis:
1999-2000, 2003-2004 and
2005-2006 cycles
3,883 out of 3,974 participants
> 20 years of age with serum
PFOA measurements; excluded
subjects with missing data (n =
91); 48.3% male, mean age
46.4 years
eGFR analysis:
1999-2000, 2003-2004, 2005-
2006, and 2007-2008 cycles
4,587 out of 5,717 (80%)
eligible 20 years or older with
PFOA measures; excluded serf-
reported CVD (n = 572),
missing data on serum
creatinine or covariates (n =
558)
0.0059 ug/mL; LOD 0.1
ng/mL
Quartiles, ug/mL, n
1< 0.0028 ug/mL, 1,176
20.0028-0.0041 ug/mL,
1,141
3 0.0042-0.0059 ug/mL,
1,141
4 > 0.0059 ug/mL, 1,129
Mean change in uric acid, mg/dL (95% CI) by
quartile
1 referent
20.14(0.04-0.25)
3 0.37 (0.25-0.49)
4 0.44 (0.32-0.56), p trend 0.0001
Mean change in uric acid, mg/dL (95% CI) by In
PFOA: 0.22 (0.15-0.30)
Multivariate regression adjusting for age, gender,
race/ethnicity, education, smoking, alcohol
consumption, BMI, hypertension, diabetes, and
serum total cholesterol
Hyperuricemia risk by quartile, OR (95% CI)
1 referent
21.14(0.78-1.67)
3 1.90 (1.35-2.69)
4 1.97 (1.44-2.70), p trend 0.0001
Hyperuricemia risk per unit increase in In PFOA,
OR (95% CI): 1.43 (1.16-1.76)
Logistic regression adjusting for age, gender,
race/ethnicity, education, smoking, alcohol
consumption, BMI, hypertension, diabetes, and
serum total cholesterol
Chronic kidney disease defined as eGFR < 60
mL/min/1.73 m2
Quartile, OR (95% CI)
1 referent
20.83(0.55-1.24)
3 1.24 (0.75-2.05)
4 1.73 (1.04-2.88)
Logistic regression adjusting for age, gender,
race/ethnicity, education, smoking, alcohol
consumption, BMI, systolic blood pressure,
diastolic blood pressure, diabetes, serum TC, and
glycohemoglobin
Adjustment for PFOS did not alter association with
PFOA
Multivariate regression of association PFOA with
eGFR among subjects with and without chronic
kidney disease
P (SE) with -1.6 (0.8) and without -2.8 (0.6) chronic
kidney disease
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Reference and Study Details
PFOA Level
Results
Watkinsetal. 2013
West Virginia
Cross-sectional population-
based survey, residents near the
Washington Works plant (C8
Health Project)
9,660 (children < 18 yrs) out of
9,783 eligible with complete
data for serum creatinine,
height, and serum PFOA
Median measured PFOA
0.0283 ug/mL; range
0.0007-2.071; yearly
serum PFOA estimated for
each individual from
model used to predict
serum PFOA at time of
enrollment, historical
serum PFOA during the
first 10 years of life, 3
years before enrollment or
at birth
P (95% CI) change in unit eGFR (mL/min/1.73 m2)
per In serum PFOA, -0.75 (-1.410.010)
Linear regression adjusting for age, gender, race,
smoking, and household income; additional
adjustment for regular exercise, BMI z-score, and
TC did not alter association
No associations of predicted serum PFOA
(modeled) with eGFR
Uric Acid (risk factor for hypertension)
Occupational exposure studies. Costa et al. (2009) examined associations between serum
PFOA concentration and uric acid levels in serum in workers at a fluorochemical production
plant in Italy. Serum PFOA concentration was associated with uric acid levels (p<0.05) in 56
workers assessed concurrently over the previous 7 years. This subset of 56 workers included
currently, formerly, and never exposed with relatively high serum PFOA concentrations.
High-exposure community studies. Steenland et al. (2010) examined the association of serum
PFOA concentrations with uric acid levels in adult subjects (n = 53,458; 20->80 years)
participating in the C8 Health Project from 2005-2006. The reference range for uric acid is
2.0-8.5 mg/dL. Serum samples were separated into deciles or quintiles for analysis. The data
were analyzed by linear and logistic regression with uric acid as the outcome and PFOA as the
exposure variable. Covariates of the model included age, gender, BMI, education, smoking,
alcohol consumption, and serum creatinine. The mean serum PFOA concentration was 0.0864
|ig/mL. The mean uric acid level was 5.58 mg/dL with an IQR of 4.5-6.6 mg/dL. The increase in
uric acid from lowest to highest serum PFOA concentration decile was 0.2-0.3 mg/dL. The OR
for high serum uric acid levels increased from the lowest to the highest quintile of PFOA serum
concentrations: 1.00, 1.33 (95% CI: 1.24-1.43), 1.35 (95% CI: 1.26-1.45), 1.47 (95% CI: 1.37-
1.58), and 1.47 (95% CI: 1.37-1.58). The study showed that higher serum PFOA concentrations
were associated with higher incidence of high serum uric acid levels. The serum of C8 study
participants included several PFASs; PFOA appeared to have a greater influence on uric acid
trends than PFOS in the models employed by Steenland et al. (2010).
The C8 Science Panel (2012) combined the data from the C8 general population cohort with
data from worker cohorts and concluded that there is no probable link between PFOA and stroke,
hypertension, and coronary artery disease. The general population cohorts were not evaluated
separately in these analyses.
General population studies. Shankar et al. (2011) investigated the association between serum
PFOA (and PFOS) and uric acid concentration in the adult population of the United States by
analyzing data from the 1999-2000, 2003-2004, and 2005-2006 NHANES evaluations. The
study population included 3,883 adults (48.3% male) older than 20 years with data available for
serum PFOA, plasma uric acid, and important covariates. Regression models were used to
analyze associations with serum PFOA as a continuous variable and in quartiles. Logistic
regression models analyzed risk for hyperuricemia defined as plasma uric acid > 6.8 mg/dL in
males and > 6.0 mg/dL in females. Ln PFOA concentration was associated with increasing uric
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acid concentration in multivariate models adjusting for age, gender, race/ethnicity, education,
smoking, alcohol consumption, BMI, hypertension, diabetes, and serum TC. Mean uric acid
concentration increased by 0.22 (95% CI 0.15-0.30) mg/dL per unit change in In PFOA. A
concentration-response relationship was indicated across all quartiles. In addition, an elevated
hyeruricemia risk was observed with increasing serum PFOA concentration (OR 1.43, 95% CI
1.16-1.76).
Glomerular Filtration Rate
High-exposure community studies. Watkins et al. (2013) evaluated the cross-sectional
association between PFOA exposure and kidney function among children aged 1<18 years (mean
12.4 ±3.8 years) enrolled in the C8 Health Project. A total of 9,660 participants had data
available on serum PFOA (median 0.0283 |ig/mL), as well as serum creatinine and height, which
were used to calculate an estimated glomerular filtration rate (eGFR). Linear regression was used
to evaluate the association between quartiles of measured serum PFOA concentration and eGFR.
A shift from the lowest to the highest quartile of measured, natural log-transformed
concentrations of PFOA in serum [IQR In (PFOA) = 1.63] was associated with a decrease in
eGFR of 0.75 mL/min/1.73 m2 (95% CI: -1.41, -0.1; p = 0.02) adjusting for age, gender, race,
smoking status, and household income. With increasing quartile of serum PFOA concentrations,
eGFR decreased monotonically, although the change was slight and did not attain statistical
significance (p for trend across quartiles = 0.30). PFOS also was associated with a decrease in
eGFR and showed a dose-related trend. Modeled predicted serum PFOA and PFOS
concentrations were not associated with eGFR.
General population studies. Shankar et al. (2011) also used data from the NHANES to
determine whether there was a relationship between serum PFOA levels and chronic kidney
disease defined as eGFR (determined from serum creatinine) of less than 60 mL/min/1.73 m2.
Serum PFOA levels were categorized into quartiles: Ql = <0.0028 |ig/mL; Q2 = 0.0028-0.0041
Hg/mL; Q3 = 0.0042-0.0059 |ig/mL; Q4 = >0.0059 |ig/mL. The adjusted OR for chronic kidney
disease for individuals in Q4 was 1.73 (95% CI: 1.04, 2.88; p for trend = 0.015) compared with
individuals in Ql. The logistic regression model was adjusted for age, gender, race/ethnicity,
education, smoking, alcohol consumption, BMI, systolic blood pressure, diastolic blood pressure,
diabetes, serum TC, and glycohemoglobin. Although a similar increase in OR was seen for
PFOS, additional adjustment for serum PFOS did not alter the association with PFOA. In
addition, the inverse association of eGFR with serum PFOA was observed over all quartiles of
PFOA, as well as among individuals both with and without chronic kidney disease. Although the
possibility of reverse causality could not be excluded, the association between serum PFOA and
eGFR among participants without chronic kidney disease suggests a PFOA-related effect on
kidney function.
Overall, studies of occupational cohorts (Costa et al. 2009), a highly exposed community
(Steenland et al. 2010; Watkins et al. 2013), and the U.S. general population (Shankar et al.
2011) that evaluated uric acid levels or eGFR as a measure of kidney function found associations
with decreased function, although reverse causality as an explanation cannot be ruled out. Since
the URAT transporter functions in the renal resorption of PFOA, the increase in serum uric acid
could be a reflection of systemic transport pharmacodynamics rather than formation
biochemistry.
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Kidney Disease
The occupational mortality studies have produced generally negative results with respect to
the association between PFOA and mortality due to chronic kidney disease (Steenland et al.
2015; Steenland and Woskie 2012; Raleigh et al. 2014). The most recent update of incidence of
chronic kidney disease in the workers in the C8 West Virginia community identified 43 cases
(with medical validation) (Steenland et al. 2015); no association was seen with cumulative
exposure when analyzed without a lag (HRs by quartile 1.0, 0.50, 0.69, 0.67; trend p = 0.92 for
log cumulative exposure), or using a 10-year lag (HRs by quartile 1.0, 1.32, 0.50, and 0.67; trend
p = 0.99).
In 2012, the C8 Science Panel concluded that there is no probable link between PFOA and
chronic kidney disease. Their conclusion was based on findings in combined general population
and worker cohorts, data on children enrolled in the C8 Health Project, and published data from
NHANES.
3.1.1.5 Immunotoxicity
Immune suppression
Immune functionspecifically immune system suppressioncan affect numerous health
outcomes, including risk of common infectious diseases (e.g., colds, flu, otitis media) and some
types of cancer. The World Health Organization (WHO) guidelines for immunotoxicity risk
assessment recommend measures of vaccine response as a measure of immune effects, with
potentially important public health implications (WHO 2012).
Associations between prenatal PFOA exposure and risk of infectious diseases (as a marker of
immune suppression) were not seen in two studies, although there was some indication of effect
modification by gender (i.e., associations seen in females but not in males). Fei et al. (2010a)
examined hospitalizations for infectious diseases in early childhood in a Danish birth cohort.
Mean maternal PFOA concentration was 0.0056 |ig/mL. A slightly higher risk for
hospitalizations was observed in females with higher maternal PFOA concentrations (incidence
rate ration [IRR] = 1.00, 1.20, 1.63, 1.74 for Ql, Q2, Q3, and Q4, respectively), and the risk for
males was below 1.0 for each quartile. Overall, there was no association between hospitalizations
due to infectious diseases and maternal PFOA exposure; similar results were found with PFOS.
Okada et al. (2012) examined history of otitis media (and of allergic conditions) in children
up to the age of 18 months. Mean maternal PFOA concentration was 0.0014 |ig/mL. Cord blood
immunoglobulin E (IgE) level decreased significantly with high maternal PFOA concentration
among female infants, but not male infants. No significant associations were observed between
maternal PFOA levels (and PFOS) with the incidence of otitis media (or specific types of
allergies or wheeze). Two other studies, described below, examined reported history of colds and
gastroenteritis in children up to age 3 years (Granum et al. 2013) or colds and flu in adults
(Looker et al. 2014). Granum et al. (2013) observed associations between prenatal PFOA
exposure and frequency of colds or gastroenteritis episodes, but not with a variable based on
"ever had" this condition in the past year. Looker et al. (2014) did not observe associations
between serum PFOA and "ever had" or frequency of colds or flu in the past year.
In 2012, the C8 Science Panel (2012) concluded that there is no probable link between PFOA
and common infections, including influenza, in children or adults. The panel based their
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conclusions on a subset of adult members of the cohort, a subset of mother-child pairs, and
published data from other researchers.
Three studies have examined response to one or more vaccine (e.g., measured by antibody
titer) in relation to higher exposure to PFOA in children (Grandjean et al. 2012; Granum et al.
2013) or adults (Looker et al. 2014); the latter study was conducted in the high-exposure C8
community population (Table 3-6).
Table 3-6. Summary of Epidemiology Studies of PFOA and Immune Suppression
(Vaccine Response)
Reference and Study Details
PFOA Level
Results
High-Exposure Community: Adults
Looker et al. 2014
C8 Health Project, West Virginia
2005-2006 enrollment and baseline
blood sample and questionnaires; 2010
follow-up n = 41 1 with prevaccination
blood sample - flu vaccination - 21 -day
post vaccination blood sample
Linear regression: antibody titer rise
Logistic regression: seroconversion and
seroprotection
Median 0.032
ug/mL
Ql: 0.0025-0.0137
Q2: 0.0138 -0.0315
Q3: 0.0316 -0.0903
Q4: 0.0904 -2.14
(Percentage positive) OR (95% CI), by influenza strain:
Seroconversion Seroprotection
(fourfold increase in antibody (antibody titer 1 :40 following
titer) vaccine)
Influenza B (62%) (66%)
PFOA continuous 0.80 (0.53, 1.21) 1.04 (0.68, 1.60)
Ql 1.0 (referent) 1.0 (referent)
Q2 1 .43 (0.76, 2.70) 0.76 (0.40, 1 .45)
Q3 1.39(0.73,2.66) 1.13(0.57,2.23)
Q4 0.71(0.38,1.36) 0.77(0.39,1.50)
A/H1N1 (84%) (96%)
PFOA continuous (84%) (96%)
Ql 1.0 (referent) 1.0 (referent)
Q2 0.74(0.34,2.70) 0.74(0.17,3.28)
Q3 1.11 (0.73, 2.66) 1 .59 (0.33, 7.70)
Q4 2.23(0.38,1.36) 6.47(0.91,45.9)
A/H3N2 (65%) (84%)
PFOA continuous 0.76(0.51,1.15) 0.66(0.39,1.12)
Ql 1.0 (referent) 1.0 (referent
Q2 0.90(0.48,1.68) 0.34(0.14,0.83)
Q3 1.13(0.59,2.17) 0.28(0.11,0.70)
Q4 0.62(0.33,1.66) 0.36(0.15,0.99)
General Population: Children
Grandjean et al. 2012
Faroe Islands
Birth cohort, follow-up to age 7 yrs
n=587
Age 5 prebooster (e.g., tetanus,
diphtheria) and 4 weeks after booster
and age 7
PFOA in 3rd trimester blood sample and
in child (age 5)
Linear regression, adjusted for gender,
age, birth weight, maternal smoking,
breast-feeding, and PCBs (and time
since booster for post-booster analysis)
Granum et al. 20 1 3
Norway
Birth cohort, Norwegian Mother and
Child Cohort Study
n = 56 with maternal blood at delivery
and child blood samples at 3 yrs
Linear regression, considered potential
confounders
Geometric mean
Maternal sample
0.0032 ug/mL
Child's sample
0.004 ug/mL
Mean 0.001 ug/mL
Log PFOA and Log antibody Beta (95% CI) [% change in antibody
titer per twofold increase in PFOA]
Maternal PFOA Tetanus Diphtheria
Prebooster -10.5 (-28.2, 11.7) -16.2 (-34.2, 6.7)
Postbooster 14.5 (-10.4, 46.4) -6.2 (-22.4, 13.3)
Year 7
(adjusted for age 5) 12.3 (-8.6, 38.1) -16.8 (-32.9, 3.3)
Child's PFOA Tetanus Diphtheria
Prebooster -13.3 (-31.6, 9.9) -6.8 (-28.3,21.0)
Postbooster - 9.7 (-30.7, 17.7) -6.1 (-23.6, 15.5)
Year 7 -28.2 (-42.7, -10.1) -23.4 (-39.3, -3.4)
(adjusted for age 5)
Similar results seen with PFOS
Beta (95% CI ) (p-value), PFOA and antibody titer
Rubella -0.40 (-0.64, -0.17) (p = 0.001)
Measles -0.13 (-0.35, 0.09) (p = 0.24)
Tetanus 0.01 (-0.009, 0.10) (p = 0.92)
Hib -0.05 (-3.85, 3.74) (p = 0.98)
Simlar results for other PFAS
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A cohort of 411 adult members of the C8 Health Project was evaluated in 2010 for an
association between serum PFOA levels and antibody response following vaccination with an
inactivated trivalent influenza vaccine (Looker et al. 2014). A prevaccination serum sample was
collected at the time of vaccination and the postvaccination serum sample was collected 21 ± 3
days later. The geometric mean serum PFOA level was 0.0337 |ig/mL (95% CI 0.0298, 0.0382)
and participants were divided into quintiles for analyses. PFOA was negatively associated with
geometric mean A/H3N2 antibody titer rise, but no association was found with antibody liters for
A/H1N1 and influenza type B. No association was found between antibody liters and PFOS levels.
Antibody responses to diphtheria and tetanus toxoids following childhood vaccinations were
assessed in context of exposure to five perfluorinated compounds (Grandjean et al. 2012). The
prospective study included a birth cohort of 587 singleton births during 1999-2001 from the
National Hospital in the Faroe Islands. Serum antibody concentrations were measured in children
at age 5 years prebooster, approximately 4 weeks after the booster, and at age 7 years. Prenatal
exposures to perfluorinated compounds were assessed by analysis of serum collected from the
mother during week 32 of pregnancy (PFOA geometric mean 0.0032 |ig/mL; IQR 0.00256-
0.00401); postnatal exposure was assessed from serum collected from the child at 5 years of age
(PFOA geometric mean 0.00406 |ig/mL; IQR 0.00333-0.00496). Multiple regression analyses
with covariate adjustments were used to estimate the percent difference in specific antibody
concentrations per twofold increase in PFOA concentration in both maternal and 5-year serum.
Maternal PFOA serum concentration was negatively associated with antidiphtheria antibody
concentration (-16.2%) at age 5 before booster. The biggest effect was found in comparison of
antibody concentrations at age 7 with serum PFOA concentrations at age 5 where a twofold
increase in PFOA was associated with differences of-36% (95% CI, -52%14%) and -25%
(95% CI, -43%2%) for tetanus and diphtheria, respectively. Additionally at age 7, a small
percentage of children had antibody concentrations below the clinically protective level of
0.1 international unit (IU) /mL. The ORs of antibody concentrations falling below this level were
4.20 (95% CI, 1.54-11.44) for tetanus and 3.27 (95% CI, 1.43-7.51) for diphtheria when age 7
antibody levels were correlated with age 5 PFOA serum concentrations. Maternal and child
PFOS levels also were negatively associated with antibody liters in children.
The effects of prenatal exposure to perfluorinated compounds on vaccination responses and
clinical health outcomes in early childhood were investigated in a subcohort of the Norwegian
Mother and Child Cohort Study (Granum et al. 2013). A total of 56 mother-child pairs, for whom
both maternal blood samples at delivery and blood samples from the children at 3 years of age,
were evaluated. Antibody liters specific to measles, rubella, lelanus, and influenza were
measured as Ihese vaccines are part of Ihe Norwegian Childhood Vaccination Program. Serum
IgE levels also were measured. Clinical heallh oulcomes, including common colds and
gaslroenlerilis, at ages 1, 2, and 3 years were assessed by means of a questionnaire senl lo
participant. Mean maternal plasma PFOA concenlralion was 0.0011 |ig/mL al delivery; Ihe
PFOS level was 0.0056 |ig/mL and PFNA and PFHxS were below Ihe LOQ. PFOA levels in Ihe
children were nol measured. No associations were found wilh PFOA or any perfluorinated
compound and antibody levels lo Ihe vaccines wilh one exception. A slighl, but significanl,
inverse relationship belween maternal PFOS level and anti-rubella antibodies in children al
3 years was found (P = -0.8 [95% CI -0.14, -0.02]). Maternal PFOA levels were nol associated
wilh adverse childhood heallh oulcomes.
In summary, Ihree sludies have reported decreases in response lo one or more vaccines
(e.g., measured by antibody liter) in relation lo higher exposure lo PFOA in children (Grandjean
Perfluorooclanoic acid (PFOA) - May 2016 3-24
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et al. 2012; Granum et al. 2013) and adults (Looker et al. 2014). In the two studies examining
exposures in the background range (i.e., general population exposures, < 0.010 |ig/ml), the
associations with PFOA also were seen with other correlated PFASs. This limitation was not
present in the study in adults in the high-exposure C8 community population. Serum PFOA
levels in this study population were approximately 0.014-0.090 |ig/mL.
Asthma
The association between serum levels of perfluorinated compounds and childhood asthma
was investigated by Dong et al. (2013). The cross-sectional study included a total of 231 children
aged 10-15 years with physician-diagnosed asthma and 225 age-matched nonasthmatic controls.
Between 2009 and 2010, asthmatic children were recruited from two hospitals in Northern
Taiwan, while the controls were part of a cohort population in seven public schools in Northern
Taiwan. Serum was collected for measurement of 10 perfluorinated compounds, absolute
eosinophil counts, total IgE, and eosinophilic cationic protein. A questionnaire was administered
to asthmatic children to assess asthma control and to calculate an asthma severity score
(e.g., frequency of attacks, use of medicine, and hospitalization) during the previous 4 weeks.
Associations of perflourinated compound quartiles with concentrations of immunological
markers and asthma outcomes were estimated using multivariable regression models. Nine of
10 perfluorinated compounds were detectable in >84.4% of all children with levels generally
higher in asthmatic children than in nonasthmatics. Serum concentrations of PFOA in asthmatic
and nonasthmatic children were 0.0015 ± 0.0013 |ig/mL and 0.0010 ± 0.0011 |ig/mL,
respectively; four other compounds were measured at higher concentrations with the highest
levels for PFOS and perfluorotetradecanoic acid. The adjusted ORs for asthma association with
the highest versus lowest quartile levels were significantly elevated for seven of the compounds.
For PFOA, the OR was 4.05 (95% CI: 2.21, 7.42). In asthmatic children, absolute eosinophil
counts, total IgE, and eosinophilic cationic protein concentration were positively associated with
PFOA levels with a significant monotonic trend with increasing serum concentration. None of
these biomarkers were significantly associated with PFOA levels in nonasthmatic children.
Serum PFOA levels were not significantly associated with asthma severity scores among the
children with asthma, although four other compounds did show an association.
Humblet et al. (2014) evaluated a cohort from NHANES to investigate children's PFAS
serum levels, including PFOA, and their association with asthma-related outcomes. Sera were
analyzed for 12 PFASs with focus on PFOA, PFOS, PFHxS, and PFNA. A total of 1,877
children aged 12-19 years with at least one serum sample available were included. Asthma and
related outcomes were self-reported. Median serum PFOA levels were 0.0043 |ig/mL for those
ever having asthma and 0.0040 |ig/mL for children without asthma. In the multivariable adjusted
model, a doubling of PFOA level was associated with an increased odds of ever having asthma
(OR=1.18, 95% CI 1.01, 1.39). PFOS was inversely associated with asthma and no associations
were found between the other PFAS and outcome.
On the basis of epidemiological and other data available, the C8 Science Panel (2012) found
no probable link between PFOA and asthma in children and adults and chronic obstructive
pulmonary disease (COPD) in adults.
Autoimmune conditions
The most recent report on the worker cohort initially described by Leonard et al. (2008)
included 6,026 workers evaluated for disease incidence, not just mortality (Steenland et al.
Perfluorooctanoic acid (PFOA) - May 2016 3-25
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2015). Lifetime serum cumulative dose was estimated by combining occupational and
nonoccupational exposures. Median measured serum level was 0.113 |ig/mL based on samples
collected in 2005. Statistically significant positive trends were found between log of cumulative
exposure and ulcerative colitis and rheumatoid arthritis. Rate ratios for the highest quartile
compared to the lowest quartile were 2.74 (95% CI 0.78, 9.65) for ulcerative colitis and 4.45
(95% CI 0.99, 19.9) for rheumatoid arthritis.
The C8 Science Panel (2012) combined these data with findings from the C8 general
population cohort and concluded that there is a probable link between PFOA and ulcerative
colitis. Using historical estimates for serum PFOA, the C8 Science Panel found a significant
positive, dose-response trend with a relative risk (RR) for the highest quartile compared to the
lowest of 3.18 (95% CI 1.84, 5.51). The panel concluded that there was no probable link between
PFOA and autoimmune diseases, including rheumatoid arthritis, lupus, typel diabetes, Crohn's
disease, or multiple sclerosis. The C8 Science Panel also concluded that there is no probable link
between PFOA and osteoarthritis. These analyses by the panel included both worker and general
population cohorts.
3.1.1.6 Thyroid Effects
Several epidemiology studies have evaluated thyroid function and/or thyroid disease and its
association with serum PFOA concentrations. Thyroid disease is more common in females than
in males. Among the PFOA studies, the three most highly powered studies with the largest
number of participants are one from the general U.S. population (Melzer et al. 2010) and two
from highly exposed individuals within the C8 population (Lopez-Espinosa et al. 2012; Winquist
and Steenland 2014b). Two of these studies are of adults (Melzer et al. 2010; Winquist and
Steenland 2014b) and one is of children/adolescents (Lopez-Espinosa et al. 2012).
Hypothyroidism is characterized by elevated thyroid stimulating hormone (TSH) and low T4;
elevated TSH in conjunction with normal T4 and triiodothyronine (T3) is defined as subclinical
hypothyroidism. Hyperthyroidisim is characterized by elevated T4 and low TSH; low levels of
TSH in conjunction with normal T4 and T3 is defined as subclinical hyperthyroidism. Some
studies focused on the prevalence of clinically defined disease (or the subclinical state), and
others examined variations in TSH, T4, and T3 measurements among people who have not been
diagnosed with a thyroid disease. Both hypothyroidism and hyperthyroidism can result from an
autoimmune pathogenesis involving destruction of thyroid tissue. A summary of the studies on
PFOA's association with thyroid disease or changes in thyroid hormones follows, and is depicted
in Table 3-7 (studies in adults) and Table 3-8 (studies in special populationschildren and
pregnant females).
Occupational exposure studies. Serum PFOA levels were obtained from volunteer workers of
the Cottage Grove, Minnesota, PFOA plant in 1993 (n = 111) and 1995 (n = 80) as part of the
medical surveillance program and analyzed to determine a relationship between TSH and PFOA
concentration (Olsen et al. 1998). Employees were placed into four exposure categories based on
their serum PFOA levels: 0-1 |ig/mL, 1- < 10 |ig/mL, 10- < 30 |ig/mL, and >30 |ig/mL.
Statistical methods used to compare PFOA levels and hormone values included multivariable
regression analysis, ANOVA, and Pearson correlation coefficients. TSH was significantly
(p = 0.002) elevated in 10-<30 |ig/mL exposure category for 1995 only (mean serum TSH level
was 2.9 ppm). However, mean TSH levels for the other exposure categories, including the
>30 |ig/mL category, were all the same (1.7 ppm). In 1993, TSH was elevated in this same
exposure category, but was not statistically significant (p = 0.09) when compared to the other
exposure categories.
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Table 3-7. Summary of Epidemiology Studies of PFOA and Thyroid Effects in Adults
Reference and Study Details
PFOA Level
TSH
T3
T4
Occupational Exposure Studies
Olsen and Zobel 2007
3M. Antwerp, Cottage Grove, Decatur combined;
50-65% participation rate
n=506
Mean age: 40 yrs
Mean duration: not reported
Linear regression adjusting for In age, In BMI, In
alcohol
(Related references: Olsen etal. 1998,2003)
Steenlandetal. 2015
n= 3,713 workers
Data collected in 2005-2006 and 2008-201 1
n = 82 cases in men, 77 cases in women
Mean (range)
2.21 (0.01-92.03)
Hg/mL
In 2005-2006: mean
0.325 |ig/mL, median
0.113|ig/mL
Beta (±SE) (p-value), In
PFOA and In TSH:
0.0360 (± 0.0207) (p = 0.08)
Beta (±SE) (p-value), hi
PFOA and In T3:
0.0105 (± 0.0053) (p = 0.05)
Beta (±SE) (p-value), hi PFOA
and In T4:
-0.0057 (± 0.0054) (p = 0.29)
Beta (±SE) (p-value), hi PFOA
and In FT4:
-0.01 17 (± 0.0043) (p = 0.01)
FIR (95% CI), for self-reported thyroid disease, with medical record validation (incidence based on
year of diagnosis). Cumulative exposure quartile, no lag
In men: In women:
1.0 (referent) 1.0 (referent)
1.64(0.82,3.29) 1.00(0.54,1.87)
1.13(0.50,2.54) 1.02(0.48,2.17)
2.16(0.98,4.77) 0.33(0.08,1.26)
(Arend=0.98) (Ptrend = 0.97)
Adults: High-Exposure Community Studies
Emmett et al. 2006
n = 371, aged 2-89 yrs, median 50 yrs
(317 from stratified random population sample)
t-test, PFOA in abnormal vs normal TSH levels
Winquist and Steenland 2014b
n = 32,254 (including 3,713 workers)
Data collected in 2005-2006 and 2008-201 1
n = 2,008 cases in primary analysis
n = 454 cases in prospective analysis (diagnoses
after 2005-2006)
Stratified by gender; also conducted separate
analyses for hyperthyroidism and hypothyroidism
0.354 |ig/mL
In 2005-2006: mean
0.0866 |ig/mL,
median 0.0261
Hg/mL
6% abnormal; p-value oft-
test comparing PFOA in
abnormal and normal = 0.59
Not measured
Not measured
HR (95% CI), incident thyroid disease (with medical record validation), primary analysis:
Cumulative exposure quintiles Year exposure quintiles:
Full sample Men Women Full sample Men Women
1.0 (referent) 1.0 1.0 1 .0 (referent) 1.0 1.0
1.21 1.12 1.24 1.23 1.13 1.26
1.17 0.83 1.27 1.24 1.11 1.28
1.27 1.01 1.36 1.10 1.06 1.11
1.2 1.05 1.37 1.28 1.04 1.38
(P»w=0.03) (P»w=0.85) (P»w=0.03) (Ptrmd=OM) (Ptrend=0.97) (Ptrend = 0.008)
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Reference and Study Details
PFOA Level
TSH
T3
T4
Diagnoses after 2005:
Cumulative exposure quintiles Year exposure quintiles:
Full sample Men Women Full sample Men Women
1.0 (referent) 1.0 1.0 1 .0 (referent) 1.0 1.0
1.23 1.35 1.23 0.80 1.32 0.74
1.00 1.37 0.93 0.91 2.09 0.76
1.06 1.44 1.00 0.93 1.83 0.82
1.12 1.85 0.96 0.91 1.76 0.80
(Ptrend=0.73) (Ptrend=0.09) (Ptrend=0.55) (Ptrmd=OM) (Plrend=0.54) (Ptrend= 0.53)
Adults: General Population Studies
Bloom etal. 2010
United States (New York; New York State Anglers
Cohort Study)
n = 3 1 (4 women)
Mean age: 39 yrs (31^15 years)
Linear regression, adjusted
PFOA-PFOS correlation r = 0.35
Shrestha etal. 2015
United States (Upper Hudson River Valley)
n = 87 (with serum for analyses); excluded if taking
thyroid medicine
Aged: 55-74 yrs
PFOA-PFOS correlation r = 0.52
Linear regression, adjusted
Melzer etal. 2010
United States, NHANES 1999-2000, 2003-2004,
and 2005-2006
n = 3,974 adults, ages > 20 yrs
Linear regression, stratified by gender, adjusted
Wen etal. 2013
United States, NHANES 2007-2008, 2009-2010
n = 1 1 8 1 , adults, aged > 20 yrs
Linear regression, adjusted, with sampling weights
Geometric mean
0.0013 ng/mL
Geometric mean
(IQR)
0.0092
(0.0071-0.0131)
Hg/mL
Men (ng/mL)
Ql: 0.0001-0.0036
Q2: 0.0037-0.0052
Q3: 0.0053-0.0072
Q4: 0.0073-0.0459
Similar cut-points in
women
Mean 0.004 ng/mL
Log-PFOA and log-TSH:
Beta = -0.06 (-0.78, 0.67)
(p = 0.87)
Log-PFOA and log-TSH:
Beta = 0.102 (-0.047, 0.25)
(p = 0.18)
Not measured
Log-PFOA and log-T3:
Beta 3.03 (-1.73, 7.79)
(p = 0.21)
Log-PFOA and log-T4:
Beta = -0.01 (-0.16, 0.14)
(p = 0.89)
Log-PFOA and log-T4:
Beta = 0.38 (-0.07, 0.83)
(p = 0.97)
Log-PFOA and log-FT4:
Beta = 0.016 (-0.036, 0.069)
(p = 0.54)
Thyroid Disease, self-reported, with medication use:
Men Women
1 (referent) 1 (referent)
1.17(0.64-2.15) 0.98(0.65-1.50)
0.58(0.21-1.59) 1.09(0.66-1.81)
1.58(0.79-3.16) 1.63(1.07-2.47)
Beta (95% CI) (p-value)
Ln-PFOAandln-TSH:
Men
0004 (-0.081, 0.090)
(p = 0.92)
Women
-0.030 (-0.2 157, 0.1 54) (p =
0.73)
Beta (95% CI) (p-value)
Ln-PFOA and ln-T3:
Men
0.775 (-3.048, 4.598) (p =
0.67)
Women
6.628(0.545, 12.7) (p = 0.035)
Ln-PFOA and ln-FT3:
Men
0.016 (0.001, 0.031) (p = 0.04)
Women
0.027 (0.009, 0.044) (p =
0.002)
Beta (95% CI) (p-value)
Ln-PFOA and ln-T4:
Men
0.000 (-0.28, 0.28)
(p=1.0)
Women
0.082 (-0.369, 0.532) (p = 0.71)
Ln-PFOA and ln-FT4:
Men
-0.010 (-0.041, 0.022) (p = 0.53)
Women
-0.004 (-0.047,0.039)
(p = 0.83)
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Table 3-8. Summary of Epidemiology Studies of PFOA and Thyroid Effects in Special Populations
Reference and Study Details
PFOA Level
TSH
T3
T4
Children: High-Exposure Community
Lopez-Espinosa et al. 2012
n = 10,725 children, aged 1-17 yrs C8 Health Project
Reported thyroid disease based on parent- report of
health care provider diagnosis of thyroid disease
(and specific types); also included current use of
thyroid medications
Subclinical disease based on hormone levels
excluding people with self-reported thyroid disease
or taking thyroid medication (subclinical
hypothyroidism = above age-specific reference
range for TSH and total T4 within reference range;
subclinical hyperthyroidism = below age-specific
reference range for TSH and total T4 within or
above reference range)
Modeled in utero
PFOA: median 0.012
ug/mL
Measured in children:
median 0.0293 ug/mL
Beta (95% CI) for % change
inTSHperlQRln-PFOA:
in utero -0.5 (-2.4, 1.5)
in child 1.0 (-0.5, 2.7)
Not measured
Beta (95% CI) for % change in
total T4 per IQR In-PFOA:
in utero -0.1 (-0.8,0.6)
in child 0.1 (-0.5, 0.6)
OR (95% CI) for thyroid disease per IQR In-PFOA:
in utero PFOA: Child's PFOA:
Any thyroid disease (n = 27) 1.45(0.95,2.27) (n=61) 1.44(1.02,2.03)
Hypothyroidism (n = 20) 1.61(0.96,2.63) (n=39) 1.54(1.00,2.37)
Subclinical hypothyroidism (n = 155) 0.94 (0.76, 1.16) (n = 365) 0.98 (0.86, 1.15)
Subclinical hyperthyroidism (n = 31) 1.10 (0.69, 1.74) (n = 78) 0.81 (0.58, 1.15)
Associations with any thyroid disease and hypothyroidism were not seen with PFOS
Children: General Population Studies
de Cock etal. 2014
Netherlands
n = 83 newboms
PFOA in cord blood samples
T4 in heel prick blood
Linear regression, stratified by gender; PFOA
quartiles, adjusted
Lin etal. 2013
Taiwan, Young Taiwanese Cardiovascular Cohort
Study
n = 545 (45 with elevated blood pressure); n = 18
hypothyroid- TSH > normal range
Aged: 12-30 yrs
0.943 ug/L
0.000943 ug/mL
Geometric mean
0.00267 ug/mL
Ql:<0.00364
Q2: 0.00364 -
<0.00066
Q3: 0.00666 -
<0.00971
Q4: >0.009.71
Not measured
Mean (±SE) Ln TSH by
PFOA quartile (adjusted)
Ql: 0.48 (±0.08)
Q2: 0.45 (±0.09)
Q3: 0.36 (±0.11)
Q4: 0.41 (±0.12)
No association with risk of
hypothyroidism
Not measured
Not measured
Beta T4 (nmol/L) (95% CI)
Boys
Ql: Referent
Q2: 7.9 (-18.04, 33.92)
Q3: -2.1 (-20.94, 16.7)
Q4: 6.2 (-16.08, 28.50)
Girls
Ql: Referent
Q2: -5.9 (-26.75, 14.94)
Q3: 11. 8 (-19.08, 42.72)
Q4: 38.6 (13.34, 63.83)
Mean (±SE) free T4 by PFOA
quartile (adjusted)
Ql: 1.07 (±0.01)
Q2: 1.08 (±0.02)
Q3: 1.10 (±0.02)
Q4: 1.06 (±0.02)
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Reference and Study Details
PFOA Level
TSH
T3
T4
Pregnant Women: General Population Studies
Chan etal. 2011
Canada
n = 96 hypothyroxinemia cases (normal TSH with
decreased free T4 - below 10th percentile) and 175
controls (normal TSK and free T4 in SO'-QO*
percentile; matching based on referring physician
and maternal age
2nd trimester blood sample (mean 1 8 weeks)
Conditional logistic regression, adjusted
PFOA-PFOS correlation r = 0.5
Wang etal. 2013
Norway (from case-control study of subfecundity in
the Norwegian Mother and Child Cohort Study;
cases and controls combined)
n = 903 women
2nd trimester blood sample (mean 1 8 weeks)
Linear regression, adjusted
Berg etal. 2015
Norway, Northern Norway Mother and Child
Contaminant Cohort Study
n=375
2nd trimester blood sample (18 weeks)
Thyroid hormones and anti-TPO antibodies
measured at 18 weeks gestation and at day 3 and
week 6 after delivery
Mixed effects linear models
Repeated measures of thyroid hormone levels were
used in model
PFOA-PFOS correlation r = 0.65
Webster etal. 2014
Canada (Vancouver Chemicals Health and
Pregnancy Study)
n = 1 52, not taking thyroid medicine
2nd trimester blood samples (15 and 18 weeks)
Mixed effects linear models, stratified by TPO
antibody levels
PFOA-PFOS correlation r = 0.71
Geometric mean
0.00135 ug/mL
Median 0.00215
ug/mL
Median 0.00153
ug/mL
Median 0.0017
ug/mL
Ln PFOA OR (95% CI):
0.94(0.74-1.18)
With additional adjustment for PFOS and PFHxS:
0.87(0.63-1.19)
PFOA and In-TSH
Beta (95% CI)
-0.0001 (-0.045, 0.044)
Highest quartile PFOA
associated with higher TSH,
but not significant when
adjusted for PFOS
(quantitative results not
reported)
Beta per IQR PFOA and TSH,
(95% CI) (p-value)
Normal TPO antibody
0.07 (-0.1, 0.2) (p = 0.41)
High TPO antibody
0.7(0.09, l)(p = 0.02)
Similar results for PFOS
[IQR PFOA = 0.0014 ug/mL]
Not measured
Quantitative results not
reported (noted as no
association)
not measured
Not measured
Quantitative results not reported
(noted as no association)
Beta per IQR PFOA and FT4,
(95% CI) (p-value)
Normal TPO antibody
-0.03 (-0.3, 0.2) (p = 0.82)
High TPO antibody
-0.4 (-1, 0.5) (p = 0.35)
[IQR PFOA = 0.0014 ug/mL]
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In an expanded and refined analysis of the data reported in Olsen et al. 2003, Olsen and
Zobel (2007) looked at the relationship between serum PFOA concentration and TSH, serum and
free T4, and T3 levels in workers at the Decatur, Antwerp, and Cottage Grove production plants.
The fluorochemical workers consisted of males (aged 21-67) from the Antwerp, Belgium
(n = 196); Cottage Grove, Minnesota (n = 122); and Decatur, Alabama (n = 188) production
facilities who volunteered to participate in the medical surveillance program in 2000. The mean
serum PFOA concentration was 2.21 |ig/mL for all sites combined. No association between
TSH, serum T4, and PFOA concentration was observed. A negative association (p<0.01)
between free T4 and serum PFOA concentration was observed in the unadjusted and adjusted
(age, BMI, and alcohol consumption) models for all locations combined; no association was
observed for the individual locations. A positive association (p<0.05) was observed between T3
and serum PFOA concentration in the unadjusted and adjusted models for all locations
combined, the Antwerp plant, and the Decatur plant. The authors noted that the results were not
considered clinically relevant because the results were within normal reference range. Steenland
et al. (2015) did not find an association between self-reported thyroid disease and PFOA levels
among 3,713 workers at the Washington Works plant in West Virginia who participated in the
C8 Health Project.
Two studies measured thyroid hormones in PFOA-exposed workers, but did not present an
analysis of the relation between PFOA exposure and hormone levels. Both studies noted that the
thyroid hormone values were in the normal range (Costa et al. 2009; Sakr et al. 2007a).
High-exposure community studies. Emmett et al. (2006) examined the association of serum
PFOA with thyroid disease in 371 residents of the Little Hocking, Ohio, water district as
described previously. No association was observed between serum PFOA and thyroid disease.
Serum PFOA was decreased (not significantly different) in subjects with self-reported disease
(e.g., hyperthyroidism, goiter or enlarged thyroid, hypothyroidism) (0.387 |ig/mL; n = 40)
compared to subjects without thyroid disease (0.451 |ig/mL; n = 331). No association was seen
between serum PFOA and TSH when analyzed with linear regression or by t-test comparison of
PFOA in the abnormal TSH (n = 24, 6%) and normal TSH groups (p = 0.59).
Participants in the C8 Health Project were examined for an association between PFOA levels
and thyroid disease (Winquist and Steenland 2014b). The cohort included 28,541 community
members and 3,713 workers who had completed study questionnaires during 2008-2011. The
median serum PFOA level at enrollment in 2005-2006 was 0.0261 |ig/mL for the combined
cohort, 0.0242 |ig/mL for the community members, and 0.1127 |ig/mL for the workers.
Retrospective serum levels for the community cohort, estimated from air and water
concentrations, residential history, and water consumption rates, were used to estimate yearly
intakes. For the workers, yearly serum estimates were modeled from work history information
and job-specific concentrations. Cox proportional hazard models, stratified by birth year, were
used to assess self-reported adult thyroid disease hazard in relation to time-varying yearly or
cumulative (sum of yearly estimates) estimated PFOA serum concentration, controlling for
gender, race, education, smoking, and alcohol consumption. For the combined cohort, quintiles
for yearly exposure were 0.00011-<0.0047, 0.0047-0.00849, 0.00849-O.0216, 0.0216-
O.100, and 0.100-3.303 |ig/mL; quintiles for cumulative exposure were 0.0001-O.115, 0.115-
<0.202, 0.202-<0.497, 0.497-2.676, and 2.676-97.396 |ig/mL-year. As expected, the number of
thyroid disease cases was higher among females than among males. Positive associations were
seen with the cumulative exposure and the per-year exposure metrics for incidence of all thyroid
disease (as well as for specific subtypes), with the observations seen primarily in females
Perfluorooctanoic acid (PFOA) - May 2016 3-31
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(Table 3-7). When limited to disease occurring after the 2005-2006 serum collection, the number
of incident cases was reduced from 2,008 to 454, and the patterns of associations were more
variable. No associations between estimated serum PFOA level and thyroid disease were found
in the analysis limited to workers in this study population (Steenland et al. 2015).
The C8 Science Panel (2012) used data from the C8 general population cohort and concluded
that there is a probable link between PFOA and thyroid disease.
General population studies. Bloom et al. (2010) investigated the associations between serum
PFAS, including PFOA, and TSH and free thyroxine (FT4). The serum samples came from 31
participants (27 males, 4 females; mean age 39 years) of the 1995-1997 New York State Angler
Cohort Study Dioxin Exposure Substudy. The study subjects each completed a questionnaire and
provided a blood sample for serum analysis. The questionnaire contained questions about sport-
fish and game consumption, lifestyle, demographic factors, and medical history. The serum
samples were analyzed for TSH and FT4 in 2003 by immunometric chemiluminescent sandwich
assay and for PFAS in 2006 by ion pair extraction high-performance LC-MS/MS. Regression
models were used to analyze the data and adjust for confounders. No subjects reported use of
thyroid medication or physician-diagnosed goiter or thyroid conditions. Mean TSH concentration
(range 0.43-15.70 |iIU/mL) was within normal range (0.40-5.00 |iIU/mL) with the exception of
one subject. Mean FT4 (0.90-1.55 ng/dL) was within normal range (0.80-1.80 ng/dL) for all
subjects. The mean serum PFOA concentration was 0.00133 |ig/mL and ranged from 0.00057 to
0.00258 |ig/mL. The males had a significantly higher serum PFOA concentration than the
females (0.00147 |ig/mL versus 0.00105 |ig/mL; p = 0.047). There was no association between
serum PFOA concentration (or PFOS) and TSH or FT4.
The relationship between serum levels of PFOA, PFOS and other persistent organic
pollutants and thyroid biomarkers was investigated in older adults (Shrestha et al. 2015). Levels
of TSH, FT4, T4, and T3 were measured in 51 males and 36 females with a mean age of 63.6
years. None of the participants had thyroid disease or were taking thyroid medication. Covariates
in the analysis included age, gender, education level, the sum of poly chlorinated biphenyls
(£PCBs) and polybrominated diphenyl ethers (£PBDEs), smoking status, and alcohol
consumption. The mean PFOA serum level was 0.0104 ± 0.0057 |ig/mL for all participants. In
both unadjusted and adjusted models, PFOA was significantly (p<0.05 or 0.01) and positively
associated with T4 and T3; a possible dose-response was not evaluated in this small sample. A
statistical interaction was detected between age and PFOA for effects on FT4 and T4 suggesting
that the positive associations of PFOA were potentiated by age. PFOS was also positively
associated with FT4 and T4.
Melzer et al. (2010) examined the association between serum PFOA concentration and
thyroid disease in the general population of the United States by analyzing data from the 1999-
2000, 2003-2004, and 2005-2006 NHANES The population included 3,966 adults (2,066
females, 1,900 males) older than 18 years. Each of the participants answered a questionnaire, had
a physical examination, and provided blood and urine samples for analysis. Serum samples were
analyzed for PFOA concentration by solid-phase extraction coupled to isotope dilution/high-
performance LC-MS/MS. Data on diseases diagnosed by a physician and confounding factors,
including year of NHANES, age, gender, race/ethnicity, education, smoking status, BMI, and
alcohol consumption were obtained from the questionnaire. Individuals were considered to have
thyroid disease if they responded on the questionnaire to having a physician-diagnosed disease or
if they were taking medication for either hypothyroidism or hyperthyroidism.
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Regression models were used to analyze the data and adjust for confounders. Serum PFOA
concentration was divided into quartiles for each gender. In females, serum PFOA concentration
ranged from 0.0001-0.123 jig/mL (Ql = 0.0001-0.0026; Q2 = 0.0027-0.004; Q3 = 0.0041-
0.0057; Q4 = 0.0057-0.123), and in males, serum PFOA concentration ranged from 0.0001-
0.0459 |ig/mL (Ql = 0.0001-0.0036; Q2 = 0.0037-0.0052; Q3 = 0.0053-0.0072; Q4 = 0.0073-
0.0459). Females in PFOA Q4 were more likely to report current thyroid disease [OR = 2.24,
95% CI: 1.38-3.65, p = 0.002] compared to females in Ql and Q2. No association between
serum PFOA concentration and thyroid disease was observed in males. With PFOS, the opposite
was found, with males in the highest quartile, but not females, more likely to report thyroid
disease. Data interpretation was limited by the cross-sectional study design, lack of information
on the specific thyroid disorder diagnosis in the questionnaire responses, and single serum
samples for PFOA measurements taken at the same time disease status was ascertained through
the questionnaire. Thus, the possibility of reverse causality cannot be eliminated.
Another study of 1,181 members of NHANES for survey years 2007-2008 and 2009-2010
examined the association between serum PFOA levels (and 12 other PFASs) and thyroid
hormone levels (Wen et al. 2013). Multivariable linear regression models were used with serum
thyroid measures as the dependent variable and individual natural log-transformed PFAS
concentration as a predictor along with confounders. The geometric mean serum PFOA level was
0.00415 jig/mL. A positive association between PFOA level and free T3 (FT3) was found in
females as a 1-unit increase in natural log-serum PFOA increased serum total T3 concentration
by 6.628 ng/dL (95% CI 0.545, 12.712, p = .035). However, the association was no longer
significant when PFOS, PFNA, and PFHxS levels were included in the model.
A different type of examination was undertaken by Pirali et al. (2009). The study measured
intrathyroidal levels of PFOA (and PFOS) in thyroid surgical specimens to determine if a
relationship existed between PFOA and the clinical, biochemical, and histological phenotype of
thyroid disease patients. Serum PFOA concentration also was measured to determine if a
relationship existed between thyroid tissue and serum PFOA levels. Patients (n = 28; 8 males,
20 females; 33-79 years) with benign multinodular goiters (n = 15), Graves' disease (n = 7),
malignant papillary carcinoma (n = 5), and malignant follicular carcinoma (n = 1) were included
in the study. Informed consent, clinical examination, work history, thyroid hormone and
antibody measurements, thyroid ultrasound, fine-needle aspiration of nodules greater than 1 cm,
and serum samples (n = 21) were performed or collected prior to surgery. The control group
consisted of thyroid tissues collected at autopsy from subjects with no history of thyroid disease
(n = 7; 5 males, 3 females; 12-83 years) and serum samples from 10 subjects with no evidence
of thyroid disease. The student's t-test, Mann-Whitney U-test, Pearson and Spearman's
correlation tests, and chi-square test with Fisher's correction were used to compare group results.
Regression analysis was used to test the effect of different variables independently of a covariate.
The median concentration of PFOA in thyroid tissue was 2.0 ng/g (range = 0.4-4.6 ng/g).
The patients were divided into three different groups: group I (toxic and nontoxic multinodular
goiter, n = 12), group II (differentiated thyroid cancer, n = 6), and group III (Hashimoto's
thyroiditis or Graves' disease, n = 10). Thyroid PFOA concentration for the control group, group
I, group II, and group III ranged from 1.0-6.0, 0.4-4.4, 1.4-4.0, and 1.0-4.6 ng/g, respectively.
Serum PFOA concentration for the control group, group I, group II, and group III ranged from
0.004-0.0137, 0.0012-0.0166, 0.0051-0.0096, and 0.0039-0.0125 |ig/ml, respectively. The
concentration of PFOA in the thyroid and serum was similar between control and thyroid
patients at the time of measurement. Age, gender, residence, working activity, malignant /
Perfluorooctanoic acid (PFOA) - May 2016 3-33
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nonmalignant conditions, antibodies, thyroid hormone concentrations, and ultrasound parameters
were not associated with thyroid or serum PFOA concentration. There also was no correlation
between serum and thyroid PFOA concentration. Similar results were obtained with PFOS.
Children. Three studies evaluated thyroid function in children (or children and young adults)
(Table 3-8). In the children from the C8 cohort who were highly exposed to PFOA, Lopez-
Espinosa et al. (2012) observed positive associations between prenatal PFOA (modeled maternal
levels) and any thyroid disease or clinical hypothyroidism; similar results were seen with the
child's PFOA level. Associations were not seen with subclinical hypothyroidism or
hyperthyroidism, or TSH or total T4 levels among children without thyroid disease. In a study
from the Netherlands of 52 males and 31 females, increasing T4 levels in females were
associated with increasing prenatal PFOA concentrations (as measured in cord blood samples)
(de Cock et al. 2014); no associations were reported in males. A study of adolescents and young
adults (aged 12-30 years) from Taiwan did not observe associations between serum PFOA
concentrations and TSH or T4 levels (Lin et al. 2013).
Pregnant females. Several studies of thyroid have been conducted in pregnant females (Table
3-8), mostly reporting null associations between maternal PFOA concentration and thyroid status
during pregnancy (Berg et al. 2015; Chan et al. 2011; Wang et al. 2013). The exception to these
results is the only study that included an analysis stratified by presence of antithyroid peroxidise
(anti-TPO) antibodies (Webster et al. 2014), in which associations between PFOA and TSH were
seen only among females with high autoantibody levels. This finding supports the importance of
further research into the association between PFOA and autoimmunity and autoimmune
conditions.
Chan et al. (2011) examined the association between hypothyroxinemia and serum PFOA
concentration (and PFOS) in pregnant Canadian females (n = 271; 20.1-45.1 years of age,
>22 weeks of gestation) in a matched case-control study. Maternal serum from the second
trimester was collected between December 15, 2005, and June 22, 2006, as part of an elective
prenatal screen for birth defects. Serum samples were analyzed for TSH and FT4 concentrations
and PFOA. The cases of hypothyroxinemia (n = 96) had normal TSH concentrations and FT4
concentrations in the lowest 10th percentile (<8.8 pmol/L). The controls (n = 175) had normal
TSH concentrations and FT4 concentrations between the 50th and 90th percentiles (12-14.1
pmol/L). Maternal age, weight, and gestational age at blood draw and dichotomized at 50th
percentiles were included as confounders, and race was included as a covariate. Chi-square tests
and regression models were used to analyze the data. Overall, the geometric mean PFOA level
was 0.00135 |ig/mL. Statistical comparisons used the geometric mean serum PFOA
concentration in the cases of 3.10 nmol/L and 3.32 nmol/L in the controls. There was no
association between serum PFOA concentration (or PFOS) and hypothyroxinemia in pregnant
females.
A cross-sectional study of 903 pregnant females evaluated the association between plasma
PFOA levels and plasma TSH (Wang et al. 2013). Twelve other PFASs also were quantified and
evaluated. The females were a cohort of the Norwegian Mother and Child Cohort Study and the
blood samples were drawn at approximately week 18 of gestation. The median PFOA
concentration was 0.0022 |ig/mL with an interquartile range of 0.00157-0.00295 jig/mL. No
association was found between plasma levels of PFOA and TSH. PFOS was associated with
higher TSH levels, but plasma levels of other PFASs were unrelated to TSH.
Perfluorooctanoic acid (PFOA) - May 2016 3-34
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Expanding on the above study, Berg et al. (2015) investigated the association between a
number of PFASs, including PFOA, and TSH, T3, T4, FT3, and FT4. A subset of 375 females in
the Norwegian Mother and Child Cohort Study with blood samples at about gestational week 18
and at 3 days and 6 weeks after delivery were included. Seven compounds were detected in more
than 80% of the blood samples, with PFOS present in the highest concentration followed by
PFOA. The median PFOA level was 0.00153 jig/mL, and the females were assigned to quartiles
based on the first blood sample at week 18 of gestation. Females in the highest quartile had
significantly higher mean TSH than females in the first quartile; however, when PFOS
concentration was included as a covariate, the association was not significant.
A study of Canadian females (n = 152) evaluated maternal serum PFOA levels (and PFHxS,
PFNA, and PFOS) for associations with thyroid hormone levels during the early second trimester
of pregnancy, weeks 15-18 (Webster et al. 2014). Mixed effects linear models were used to
examine associations between PFOA levels and FT4, total T4, and TSH; associations were made
for all females and separately for females with high levels of TPO antibody, a marker of
autoimmune hypothyroidism. Median serum PFOA was 0.0017 |ig/mL. No associations were
found between levels of PFOA (or PFOS and PFHxS), and thyroid hormone levels in females
with normal antibody levels. PFNA was positively associated with TSH. Clinically elevated TPO
antibody levels were found in 14 (9%) of the study population. In the females with high antibody
levels, PFOA, as well as PFNA and PFOS, was strongly and positively associated with TSH. An
IQR increase in maternal PFOA concentrations was associated with a 54% increase in maternal
TSH compared to the median TSH level. PFNA and PFOS concentrations were associated with
46% and 69% increases, respectively, in maternal TSH.
As illustrated above, numerous epidemiology studies have evaluated thyroid function and/or
thyroid disease in association with serum PFOA concentrations (Tables 3-7 and 3-8). As noted
previously, thyroid disease is more common in females. Several studies provide support for an
association between PFOA exposure and incidence or prevalence of thyroid disease, and include
large studies of representative samples of the general U.S. population (Melzer et al. 2010) and
the high-exposure C8 community population (Lopez-Espinosa et al. 2012; Winquist and
Steenland 2014b). Two of these studies are of adults (Melzer et al. 2010; Winquist and Steenland
2014b) and one is of children/adolescents (Lopez-Espinosa et al. 2012). The trend for an
association with thyroid disease was seen in females in the C8 population (Winquist and
Steenland 2014b) and the general population (Melzer et al. 2010), and in children (Lopez-
Espinosa et al. 2012); this was most often hypothyroidism. Association between PFOA and TSH
also was seen in pregnant females with anti-TPO antibodies (Webster et al. 2014). In contrast,
generally null associations were found between PFOA and TSH or thyroid hormones (T3 or T4)
in people who have not been diagnosed with thyroid disease.
3.1.1.7 Diabetes and Related Endpoints
Occupational exposure studies. Leonard et al. (2008) examined cause of death among former
workers at the DuPont Washington Works plant in West Virginia. The cohort consisted of 6,027
employees (4,872 males and 1,155 females) who had worked at the plant from 1948 through
2002. The DuPont Epidemiology Registry and U.S. National Death Index were used to obtain
causes of death. SMRs were estimated using three reference populations; the populations of the
United States and West Virginia and the DuPont regional worker reference population excluding
workers at the Washington Works plant. A significant increase in diabetes mortality was
observed for Washington Works plant workers compared to the DuPont regional worker
Perfluorooctanoic acid (PFOA) - May 2016 3-35
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reference population [SMR = 197, 95% CI: 123, 298]. However, no regression analyses were
done with PFOA levels.
The Leonard et al. study (2008) was updated in a cohort mortality study conducted by
Steenland and Woskie (2012) to include 5,791 individuals who had worked at the DuPont West
Virginia plant for at least 1 year between 1948 and 2002. Mean duration of employment was 19
years. Deaths through 2008 were ascertained through either the National Death Index or death
certificate data. Exposure quartiles were assessed by estimated cumulative annual serum levels
based on blood samples from 1,308 workers taken during 1979-2004 and time spent in various
job categories (ppm-years). Referent groups included both nonexposed DuPont workers in the
same region and the U.S. population. Overall, the mean cumulative exposure was 7.8 ppm-years
and the estimated average annual serum level was 0.35 jig/mL. Compared to the referent rates
from other DuPont workers, cause-specific mortality rates were elevated for diabetes (n = 38;
SMR=1.90; 95% CI 1.35, 2.61). These data are limited by the small number of cases and the
restriction to mortality as an outcome.
The most recent report on the above cohort included 6,026 workers evaluated for disease
incidence, not just mortality (Steenland et al. 2015). Lifetime serum cumulative dose was
estimated by combining occupational and nonoccupational exposures. Median measured serum
level was 0.113 |ig/mL based on samples collected in 2005. No association was found between
PFOA level and type II diabetes incidence rate.
High-exposure community studies. MacNeil et al. (2009) examined the association of PFOA
with type II diabetes in adult participants of the C8 Health Project (n = 54,468; age 20 to >80
years). Serum PFOA concentration was divided into deciles using the population distribution.
Other PFAS were not evaluated in this study. Serum PFOA (deciles), BMI, gender, family
history of diabetes, race, use of cholesterol-lowering medicine, and use of blood 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 nonfasting blood samples. The mean serum PFOA concentration for the entire
study population was 0.0868 |ig/mL and 0.0913 |ig/mL for subjects with type II diabetes
validated by medical review (n = 3,539).
There was no association between serum PFOA concentration and fasting serum glucose
level in subjects characterized as nondiabetic. The mean serum PFOA concentration was
0.0929 |ig/mL in subjects who self-reported type II diabetes (n = 4,278) and 0.1227 |ig/mL in
subjects diagnosed in the last 10 years (n = 1,055). No association was observed between type II
diabetes and serum PFOA concentration. The OR by decile was 1.00, 0.71, 0.60, 0.72, 0.65,
0.65, 0.87, 0.58, 0.62, and 0.72. The results of the analysis indicated that PFOA exposure is not
associated with type II diabetes among the population studied. Data interpretation was limited by
the cross-sectional study design, which made it difficult to determine if PFOA exposure preceded
disease.
The C8 Science Panel (2012) combined these data from the C8 general population cohort
with follow-up data and data from worker cohorts, and concluded that there is no probable link
between PFOA and type II diabetes.
General population studies. Preconception serum levels of PFOA (and other PFASs) were
evaluated in females attempting pregnancy in relation to risk of developing gestational diabetes
Perfluorooctanoic acid (PFOA) - May 2016 3-36
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(Zhang et al. 2015). The 258 participants were members of the Longitudinal Investigation of
Fertility and the Environment (LIFE) study with blood samples taken during 2005-2009. The
ORs and 95% CIs of gestational diabetes associated with each SD increment of preconception
serum PFOA concentration (log-transformed) (and six other PFASs) were estimated with the use
of logistic regression after adjusting for confounders. Preconception mean serum PFOA levels
were 0.0033 |ig/mL for the entire cohort, 0.00394 |ig/mL in females with gestational diabetes
and 0.00307 |ig/mL in females without gestational diabetes. A significant positive association
was found between PFOA and risk of gestational diabetes in the fully adjusted model (OR=1.86;
95% CI 1.14, 3.02). Associations for six other PFAS were slightly increased (e.g., PFOS
OR=1.13), but did not attain statistical significance.
Metabolic syndrome is a combination of medical disorders and risk factors that increase the
risk of developing cardiovascular disease and diabetes. Lin et al. (2009) investigated the
association between serum PFOA (plus three other PFASs) and glucose homeostasis and
metabolic syndrome in adolescents (aged 12-20 years) and adults (aged >20 years) by analyzing
the 1999-2000 and 2003-2004 NHANES data. The National Cholesterol Education Program
Adult Treatment Panel III guidelines were used to define adult metabolic syndrome and the
modified guidelines were used to define adolescent metabolic syndrome. The study population
included 1,443 subjects (474 adolescents, 969 adults) at least 12 years of age who had a morning
examination and triglyceride measurement. There were 266 male and 208 female adolescents
and 475 male and 493 female adults. Multiple linear regression and logistic regression models
were used to analyze the data. Covariates included age, gender, race, smoking status, alcohol
intake, and household income. Log-transformed PFOA concentration was 1.51 and 1.48 ng/mL
for adolescents and adults, respectively. In adults, serum PFOA concentration was associated
with increased p-cell function (P coefficient 0.07, p<0.05). Serum PFOA concentration was not
associated with metabolic syndrome, metabolic syndrome waist circumference, glucose
concentration, homeostasis model of insulin resistance, or insulin levels in adults or adolescents.
Both PFOS and PFNA were positively associated with some of the endpoints associated with
metabolic syndrome.
Nelson et al. (2010) examined the relationship between polyfluoroalkyl chemical serum
concentration, including PFOA, and insulin resistance as previously described for data from
NHANES. Fasting insulin and fasting glucose were used to determine the homeostatic model
assessment for insulin resistance. No association was found between serum PFOA concentration,
or any other PFAS, and insulin resistance.
Overall, these studies show a lack of association of PFOA with diabetes, metabolic
syndrome, and related endpoints.
3.1.1.8 Reproductive and Developmental Endpoints
Several studies have examined the relationship between PFOA exposures and reproductive,
gestational, and developmental endpoints as well as postnatal growth and maturation in humans.
Pregnancy-related endpoints include gestational age (Nolan et al. 2009), measures of fetal
growth (Apelberg et al. 2007; Fei et al. 2007, 2008a; Monroy et al. 2008; Nolan et al. 2009; Stein
et al. 2009; Washino et al. 2009), miscarriage or preterm birth (Stein et al. 2009), birth defects
(Stein et al. 2009), hypertension and preeclampsia (Darrow et al. 2013; Savitz et al. 2012a,
2012b; Stein et al. 2009), and fecundity (Fei et al. 2009; Velez et al. 2015). Infant growth and
development during the first 7 years (Andersen et al. 2010, 2013; Fei et al. 2008b, 2010a, 2010b;
H0yer et al. 2015b) and postnatal growth and maturation, including neurodevelopment (Fei and
Perfluorooctanoic acid (PFOA) - May 2016 3-37
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Olsen 2011; Hoffman et al. 2010; H0yer et al. 2015a; Liew et al. 2014; Stein et al. 2013) and risk
of adult obesity (Halldorsson et al. 2012) also have been studied. Male reproductive endpoints
evaluated in humans include sperm count and semen quality (Buck Louis et al. 2015; Joensen et
al. 2009, 2013; Vested et al. 2013). Female pubertal development was examined in three studies
(Christensen et al. 2011; Kristensen et al. 2013; Lopez-Espinosa et al. 2011). As noted
previously, the focus of this review is on pregnancy-related outcomes, specifically pregnancy-
related hypertension and preeclampsia, measures of fetal growth, and pubertal development.
Within each section, the discussion is divided into occupational exposure studies (if applicable),
the C8 high-exposure community studies, and general population studies.
Several analyses are based on the Danish National Birth Cohort (Andersen et al. 2010, 2013;
Fei et al. 2007, 2008a, 2008b, 2009, 2010a, 2010b; Fei and Olsen 2011). The females (n = 1,400)
and their infants were randomly selected, and the study included those who provided their first
blood samples between gestational weeks 4 and 14 and gave birth to a single live-born child
without congenital malformation. The females participated in telephone interviewsat 12 and
30 weeks gestation, when the children were 6 and 18 months of age, and when the children were
7 years of ageand filled out a food frequency questionnaire. As the children aged, more
questionnaires were completed by the mothers with regard to behavioral health and motor
coordination. Highly structured questionnaires were used to gather information on possible
confounders, including infant gender, maternal age, parity, socio-occupational status,
prepregnancy BMI, and smoking during pregnancy. The National Hospital Discharge Register
was used to obtain birth weight, gestational age, placental weight, birth length, head and
abdominal circumference data, Apgar scores based on heart rate, respiratory effort, reflex,
irritability, muscle tone, and skin color. Plasma PFOA concentration was determined from the
first blood samples of 1,399 females, from the second blood samples of 200 females, and from
cord blood samples of 50 infants by solid-phase extraction high-performance LC-MS/MS. PFOA
concentrations were divided into quartiles (Fei et al. 2009, 201 Ob), with the lowest quartile
designated as the reference group, as follows: 0.00697 |ig/mL. Regression models were used to
analyze the data. Results of these studies are included in the following discussion of results for
specific endpoints.
Pregnancy-related hypertension and preeclampsia. There are no occupational exposure and
general population studies examining pregnancy-related hypertension and preeclampsia in
relation to PFOA exposure. The only data available come from the high-exposure C8 Health
Project study population (Table 3-9).
Several studies, using different designs and exposure measures, have examined birth
outcomes, including pregnancy-induced hypertension or preeclampsia in infants born to mothers
in the high-exposure C8 community population in West Virginia and Ohio (information obtained
from questionnaire-based pregnancy histories or from linkage to birth records) (Table 3-9). Stein
et al. (2009) used an exposure measure based on individual serum PFOA levels obtained in the
2005-2006 baseline survey to examine birth outcomes (based on self-report) in 1,845 births in
the 5 years preceding the PFOA measurement. Savitz et al. (2012a, 2012b) included births from
1990 to 2004, modeling exposure based on the serum measurements in 2005, information
obtained in the 2005 baseline questionnaire regarding residential history, information on
historical environmental releases, and PKs. In one of the analyses (study II in Savitz et al.
2012b), linkage with birth records was used to verify the preeclampsia outcome.
Perfluorooctanoic acid (PFOA) - May 2016 3-38
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Table 3-9. Summary of Epidemiology Studies of PFOA and Pregnancy-Induced
Hypertension or Preeclampsia
Study
PFOA Level
Results
Stein et al. 2009
United States (C8 Health Project)
n = 1,845 pregnancies in the 5 years
before enrollment
Exposure based on serum collected in
2005
Outcome based on pregnancy history
collected in 2005
Median 0.0212
ug/mL
OR (95% CI), preeclampsia
per IQR (InPFOA) increase in PFOA:
1.1 (0.9, 1.3) [IQR(lnPFOA)=0.0395 ug/mL]
< 50th percentile 1.0 (referent)
>50* 1.3 (0.9, 1.9)
< 50th percentile 1.0 (referent)
50-75* 1.5 (1.0, 2.3)
75-90* 1.2(0.7,2.1)
>90th 0.9(0.5,1.8)
Savitzetal. 2012a
United States (C8 Health Project)
n = 11,737 pregnancies in 1990-2006
Modeled exposure based on serum
collected in 2005, residential history, and
other data
Outcome based on pregnancy history
collected in 2005
OR (95% CI), preeclampsia
per IQR(lnPFOA) increase in PFOA:
1.13 (1.00-1.28) [IQR (lnPFOA)=0.00219 ug/mL]
per 0.100 ug/mL increase in PFOA:
1.08(1.01-1.15)
< 40th percentile 1.0 (referent)
40-60* 1.2 (1.0, 1.5)
60-80* 1.1 (0.9, 1.4)
>80* 1.2(1.0,1.6)
Also noted stronger associations with Bayesian
calibration of exposure and among women with
highest quality residential history
Savitzetal. 2012b
United States (C8 Health Project)
n = 4,547 pregnancies in 1990-2004
Modeled exposure based on serum
collected in 2005, residential history.
other data
Outcome based on linkage to birth
records
Median 0.0134
ug/mL
and
With Bayesian calibration of exposure:
OR (95% CI)
per IQR (InPFOA) increase in PFOA:
1.13 (0.92, 1.37) [IQR (lnPFOA)=0.00192 ug/mL]
per 0.100 ug/mL increase in PFOA:
0.97(0.85, 1.11)
< 40* percentile 1.0 (referent)
40-60* 1.0 (0.7, 1.4)
60-80* 1.5(1.1,2.1)
>80* 1.2(0.8,1.7)
Darrowetal. 2013
United States (C8 Health Project)
n = 1,330 pregnancies in 2005-2010; 770
first pregnancies after PFOA measures;
947 (pregnancies in 2005-2007)
Exposure based on serum collected in
2005
Geometric mean
0.016 ug/mL
Mean 0.031
OR (95% CI) per log unit increase in PFOA
Full analysis: 1.27 (1.05, 1.55)
(adjusted for PFOS) 1.22 (0.99, 1.51)
By quintile:
Ql up to 0.0069 ug/mL 1.0 (referent)
Q2 0.0069-< 0.0111 2.39(1.05,5.46)
Q3 0.0111-< 0.0189 3.43(1.50(7.82)
Q4 0.0189-< 0.0372 3.12(1.35,7.18)
Q5> 0.0372 3.16(1.35,7.38)
First pregnancy after
PFOA measure 1.23 (0.92, 1.64)
Pregnancies in 2005-2007: 1.35 (1.04, 1.76)
Darrow et al. (2013) examined birth outcomes in births that occurred in the 5 years after the
PFOA measurement. In this study, reproductive history in a follow-up interview in 2010 was
collected from females who had provided serum for PFOA measurement in 2005-2006.
Singleton live births among 1,330 females after January 1, 2005, were linked to birth records to
identify outcomes of pregnancy-induced hypertension and other outcomes (e.g., preterm birth,
low birth weight, and birth weight among full-term infants). Thus there is a progressively greater
refinement and reduction in misclassification (or exposure and outcome) among this set of
Perfluorooctanoic acid (PFOA) - May 2016
3-39
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studies. Each of these studies provides some evidence of an association between PFOA exposure
and risk of pregnancy-induced hypertension or preeclampsia, with the most robust findings from
the methodologically strongest study (Darrow et al. 2013). Maternal serum PFOA levels were
positively associated with pregnancy-induced hypertension, with an adjusted OR per log unit
increase in PFOA of 1.27 (95% CI: 1.15, 1.55). PFOS also was positively associated with
pregnancy-induced hypertension.
The C8 Science Panel (2012) considered both hypertension and preeclampsia together in
determining a link between PFOA and pregnancy-induced hypertension. Some studies conducted
by the panel found no associations while others showed positive associations. Among the studies
with positive associations, no clear dose response was indicated. However, the panel decided that
the evidence was sufficient to conclude that PFOA has a probable link to pregnancy-induced
hypertension.
Fetal growth. Many different measures of fetal growth can be used in epidemiology studies.
Birth weight is widely available (as it is routinely collected in medical records and birth
certificates). Low birth weight (defined as < 2,500 g) can be a proxy measure for preterm birth
(particularly when accurate gestational age dating is not available). Other measures of fetal
growth such as small for gestation age might more accurately reflect fetal growth retardation.
Both birth weight and gestational age are characterized as two-part distributions, with a
larger Gaussian portion representing term births and a longer tail representing preterm births.
Increased risks of complications, including infant mortality, are seen in preterm births (or low
birth-weight births). When analyzed as a continuous measure, changes in birth weight might not
be clinically significant, as small changes in the distribution among term infants do not result in a
shift into the distribution seen in preterm infants (Savitz 2007; Wilcox 2010). This consideration
differs from that of some other types of continuous measures, such as neurodevelopment scales,
blood pressure, or cholesterol, in which shifts in the distribution are expected to move a greater
proportion of the population into an "at risk" or "abnormal" level.
High-exposure community studies. As noted in the previous discussion of preeclampsia,
several studies using different designs and exposure measures have examined birth outcomes in
infants born to mothers in the high-exposure C8 community population in West Virginia and
Ohio (Darrow et al. 2013; Nolan et al. 2009; Savitz et al. 2012a, 2012b; Stein et al. 2009). These
studies include analyses of birth weight and of low birth weight, and have not observed
associations between PFOA and either birth weight among term births or the risk of low birth
weight among all (singleton) births (Table 3-10).
Based on these data, as well as continued follow-up of the community cohort, the C8 Science
Panel (2012) concluded that there is no probable link between PFOA and low birth weight.
General population studies. Two studies examined associations between maternal PFOA levels
and birth weight among term infants (Fei et al. 2007; Monroy et al. 2008). The larger of these is
from the Danish National Birth Cohort by Fei et al. (2007) (Table 3-10). In this study of 1,207
term births, the change in birth weight per log unit increase in PFOA was -9 g (95% CI: -20,
2g).
Perfluorooctanoic acid (PFOA) - May 2016 3-40
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Table 3-10. Summary of Epidemiology Studies of PFOA and Birth Weight
Study
PFOA Level
Results
High-Exposure Community
Darrowetal. 2013
United States (C8 Health
Project)
n = 1,629 pregnancies in 2005-
2010; 770 first pregnancies
after PFOA measures; 947
(pregnancies in 2005-2007)
Exposure based on serum
collected in 2005
Geometric mean
0.0162 ug/mL
Mean 0.031 ug/mL
Change in birth weight per log unit increase (95% CI)
Full analysis: -8 (-28, 12) g
(adjusted for PFOS) -4 (-25, 17) g
First pregnancy after
PFOA measure 5 (-22, 33) g
Pregnancies in 2005-2007: -10 (-34, 14) g
OR (95% CI) for low birth weight (< 2500 g) per log unit
increase
Full analysis: 0.94 (0.75, 1.17)
(adjusted for PFOS) 0.91 (0.72, 1.16)
First pregnancy after
PFOA measure 1.07 (0.78, 1.47)
Pregnancies in 2005 -2007: 0.91 (0.70, 1.17)
[Similar results in Nolan et al. 2009; Savitz et al. 2012a,
2012b; Stein etal. 2009]
General Population: Birth Weight Among Term Births
Fei et al. 2007
Denmark
n = 1,207 (term births)
Blood sample at 4-14 weeks
Monroy et al. 2008
Canada
n=101
Cord blood sample
0.0056 ug/mL
0.0019 ug/mL (cord
blood)
Change in birth weight per unit increase (95% CI)
-8.7 (-19.5,2.1)
Change in PFOA per g change in birth weight:
Beta = 0.000 17 l(p = 0.65)
General Population: Birth Weight or Low Birth Weight Among All Births (by time of blood collection)
Fei et al. 2007
Denmark
Blood sample at 4-14 weeks
n= 1,399 (full sample)
3.8%preterm
Hammetal. 2010
Canada
n = 252
Blood sample at 15-16 weeks
8.3% preterm
Whitworth et al. 2012
Norway
n = 849
Blood sample at around 17
weeks
3. 9% preterm
Maisonet et al. 2012
United Kingdom
n = 395
Blood sample at 10-28 weeks
3.1% preterm
Washino et al. 2009
Japan
n = 428
Blood sample at 23-35 weeks
% preterm not reported
0.0056 ug/mL
0.0021 ug/mL
0.0021 ug/mL
0.0037 ug/mL
0.0014 ug/mL
Change in birth weight per unit increase (95% CI)
-10.6 (-20.8, -0.47) g
OR (95% CI) for low birth weight (< 2500 g) by quartile
Ql up to 0.00390 ug/mL 1.0 (referent)
Q2 0.00391-0.00520 4.3 (0.51, 37)
Q3 0.00521-0.00696 3.7 (0.42, 32)
Q4 > 0.00697 2.4 (0.27, 22) (Trend p = 0.94)
Change in birth weight per In unit increase (95% CI)
-37.4 (-86.0, 11. 2) g
Change in birth weight z-score per unit increase (95% CI)
-0.03 (-0.10,0.04)
Change in birth weight per log unit increase
-34.2 (-54.8, -13) g
Change in birth weight per log unit increase (95% CI)
-75 (-191, 42) g
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Study
PFOA Level
Results
Apelberg et al. 2007
United States (Baltimore)
n = 293
Cord blood sample
13%preterm
0.0016 ug/mL (cord
blood)
Change in birth weight per log unit increase (95% CI)
-104 (-213, 5) g
Chen etal. 2012
Taiwan
n = 429
Cord blood sample
9.3% preterm
0.0018 ug/mL
(cord blood)
Change in birth weight per log unit increase (95% CI)
-19 (-63, 25) g
Fei et al. (2007, 2008a), and other studies in the general population have examined PFOA in
relation to birth weight or risk of low birth weight (or other measures of fetal growth), without
restriction to term births (Table 3-10). These studies vary in size from approximately 250 to
1,400 births, and also in terms of timing of exposure measure. Fei et al. (2007, 2008a) used blood
samples collected early in pregnancy (4-14 weeks), three studies used samples collected in the
second trimester (Hamm et al. 2010; Maisonet et al. 2012; Whitworth et al. 2012), Washino et al.
(2009) used samples collected in the third trimester, and two studies used cord blood samples
(Apelberg et al. 2007; Chen et al 2012). These studies also differed in the percent of births that
were preterm (ranging from approximately 3% to 13%), and presented results using different
types of analyses (i.e., the form of the exposure and outcome variables, continuous, In-
transformed, categorical, etc). Each of the analyses indicates a negative association between
PFOA levels and birth weight (i.e., a decrease in birth weight with increase in PFOA), although
CIs were wide.
In a systematic review based on the Navigation Guide methods (Woodruff and Sutton 2014),
Johnson et al. (2014) identified the general population studies shown in Table 3-10 and the high-
exposure C8 Health Project studies published through 2012. The results from the meta-analysis
showed that a 0.001 |ig/mL increase in serum or plasma PFOA was associated with a -18.9 g
(95% CI -29.8, -7.9) difference in birth weight.
Preeclampsia is a condition that causes the pregnant female to be hypertensive because of
reduced renal excretion associated with a decrease in GFR. Preeclampsia is often accompanied
by low birth weight (Whitney et al. 1987). Morken et al. (2014) used a subset of the Norwegian
Mother and Child Cohort to evaluate the relationship between GFR and fetal size. Participants
included 470 preeclamptic patients and 483 nonpreeclamptic females; plasma creatinine
measured during the second trimester was used to estimate GFR. For the overall cohort, for each
mL/min increase in GFR, infant weight at birth increased 0.73-0.83 g, depending on the method
used to calculate GFR. The increases were greater and statistically significant in females with
preeclampsia. Differences were not statistically significant for the nonpreeclamptic group.
Morken et al. (2014) was not a study of perfluorochemicals and there were no serum
measurements of any PFASs. However, because PFOA/PFOS serum levels are expected to be
higher with a lower GFR, the finding stimulated examination of the GFR as it relates to serum
PFAS levels and the low birth weight identified in the epidemiology studies (Verner et al. 2015;
Vesterinen et al. 2014).
The evidence for an inverse association between PFOA levels and birth weight raised the
question of whether reverse causality linked to maternal GFR played a role in the association of
low birth weight with serum PFOA. PFOA excretion by the kidney is dependent, in part, on the
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GFR. Conditions that result in impairment of GFR (and thus increased serum PFOA) also could
be related to fetal growth restriction, confounding the association between serum PFOA and
decreased birth weight. Vesterinen et al. (2014) examined evidence pertaining to the relationship
between fetal growth and maternal GFR using Navigation Guide systematic review methods.
They identified 35 relevant studies published between 1954 and 2012 that met the Navigation
Guide criteria for inclusion in the analysis. All studies were rated as "low" or "very low" quality
due to inconsistency of findings among studies, small sample sizes resulting in large CIs around
a mean, and high risk of bias in conduct of the study. The quality rating led to the conclusion that
data were "inadequate" to determine an association between fetal growth and GFR. However, a
more recent publication described below, expanded the database on the relationship between
GFR and fetal size.
Verner et al. (2015) modified the human pregnancy/lactation PK model of PFOA by
Loccisano et al. (2013) described in section 2.6.1 to evaluate the association between GFR,
serum PFOA levels, and birth weight. When GFR was accounted for in the model simulations,
the reduction in birth weight associated with increasing serum PFOA was less than that found by
the author's meta-analysis of the same data. This finding suggests that a portion of the
association between prenatal PFOA and birth weight is confounded by maternal GFR differences
within the populations studied. The true association for each 1 ng/mL increase in PFOA could be
closer to a 7-g reduction (95% CI -8, -6) compared to the 14.72-g reduction (95% CI: -
8.92, -1.09) predicted by meta-analysis of the epidemiology data without a correction for low
GFR as observed in individuals with pregnancy-induced hypertension or evidence of
preeclampsia.
Other pregnancy outcomes. Gestational age and preterm birth and risk of miscarriage were not
associated with PFOA in the studies examining pregnancy outcomes in the high-exposure
community (Darrow et al. 2014; Nolan et al. 2009, 2010). In contrast, PFOS was positively
associated with miscarriage (Darrow et al. 2014).
Congenital anomalies were diagnosed in 1.8%, 1.9%, and 2.0% of the mothers with water
provided completely, partially, or not at all by LHWA, respectively (Nolan et al. 2010). When
adjusted for confounders, no statistically significant differences were found. Complications with
labor and delivery were observed in 32.5%, 35.9%, and 41.9% of the mothers with water
provided completely, partially, or not at all by LHWA, respectively. Mothers with water
provided by LHWA did have in increased likelihood of having dysfunctional labor, but the
number of reported cases was low. Mothers with one or more maternal risk factors were 37.5%,
34.4%, and 39.3% of the populations with water provided completely, partially, or not at all by
LHWA, respectively. Adjusted regression models showed no statistical differences across water
service status. An increased likelihood of anemia (crude OR 11, 95% CI: 1.8-64) and
dysfunctional labor (crude OR 5.3, 95% CI: 1.2-24) in mothers with water provided by LHWA
was found, but the number of reported cases was low. No association was found between PFOA
and increased incidence of congenital anomalies, other labor and delivery complications, or
maternal risk factors.
The C8 Science Panel (2012) concluded that there is no probable link between PFOA and
birth defects, miscarriage, preterm birth, or stillbirth. Their conclusion was based on findings in
Nolan et al. (2010), Stein et al. (2009), and other data available to the panel. These other data
included historical estimates of serum PFOA generated by the panel based on amounts released
from the plant and an individual's residential history.
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Fei et al. (2009) examined the association between plasma PFOA concentration and longer
time to pregnancy (TIP) as a measure of fecundity in 1,240 females. TIP was categorized as
follows: immediate pregnancy (<1 month), 1-2, 3-5, 6-12, and >12 months. Having >12 months
TTP or having used fertility treatment to get pregnant were used to define infertility. A total of
620 females had a TTP within the first 2 months of trying to conceive and 379 had a TTP of
>6 months with 188 of those females having a TTP of >12 months. The mean plasma PFOA
concentration was 0.0056 |ig/mL for females who planned their pregnancies, and 0.0054, 0.0060,
and 0.0063 |ig/mL for TTPs <6 months, 6-12 months, and >12 months, respectively. Plasma
PFOA concentration was significantly greater (p<0.001) in females who had TTPs >6 months
than those with TTPs <6 months. The females with TTPs >6 months were more likely to be
older, have middle socio-occupational status, and have a history of spontaneous miscarriage or
irregular menstrual cycles. The adjusted odds for infertility increased 60-154% among females
with >0.00391 |ig/mL plasma PFOA concentration compared to females with <0.00391 |ig/mL
plasma concentration. The fecundity OR was 0.72, 0.73, and 0.60 for the three highest PFOA
concentration quartiles. In the likelihood ratio test, the trend was significant (p<0.001). Both TTP
and infertility also were positively associated with serum PFOS levels in this study. Although the
results of the study suggest that plasma PFOA concentration could reduce fecundity, the authors
noted that selection bias, the unknown quality of the sperm, unknown frequency and timing of
intercourse, and abnormal hormone levels might have an impact on the results and fecundity.
Participants enrolled in the Maternal-Infant Research on Environmental Chemicals Study, a
Canadian pregnancy and birth cohort, were evaluated for an association between serum PFOA
levels (as well as PFOS and PFHxS) and TTP (Velez et al. 2015). A total of 1,743 females,
enrolled between 2008 and 2011 and having a blood sample collected during the first trimester
were included. Infertility was defined as having a TTP of >12 months or requiring infertility
treatment for the current pregnancy. The geometric mean plasma PFOA level was 0.00166
|ig/mL. The crude fecundity OR per one SD increase in log-transformed serum concentration
was significantly lower for PFOA (OR=0.91, 95% CI 0.86, 0.96) (and for PFHxS). In fully
adjusted models, PFOA (and PFHxS) was associated with an 11% reduction in fecundability per
one SD increase in log-transformed serum concentration (OR=0.89; 95% CI 0.83, 0.94). The
adjusted odds of infertility increased by 31% per one SD increase of PFOA (OR=1.31; 95% CI
1.11-1.53) (and of PFHxS). No significant associations were observed for PFOS.
Fei et al. (2010b) reported on the effects of PFOA and PFOS on the length of breast-feeding.
Self-reported data on the duration of breast-feeding were collected during the telephone
interviews with each mother at 6 and 18 months after birth of the child. Higher levels of PFOA
were significantly associated with a shorter duration of breast-feeding. In multiparous females,
the adjusted OR for weaning before 6 months was 1.23 (95% CI, 1.13-1.33) for each 1-ng/mL
increase in PFOA concentration in the maternal blood and the increase was dose-related. A
similar association was observed with PFOS levels. No association was found between length of
breast-feeding and PFOA levels in females having their first child. The authors speculate that the
observed associations might be noncausally related to previous length of breast-feeding or to
reduction of PFOA and PFOS through lactation.
Pubertal development. Two studies examined development of puberty in females in relation to
prenatal exposure to PFOA as measured through maternal or cord blood samples (Christensen et
al. 2011; Kristensen et al. 2013), and another study examined PFOA exposure measured
concurrently with the assessment of pubertal status in females and in males (Lopez-Espinosa et
al. 2011) (Table 3-11).
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Table 3-11. Summary of Epidemiology Studies of PFOA and Pubertal Development
Study
PFOA Level
Results
Prenatal Exposure: General Population
Christensen et al. 2011
United Kingdom
Pregnancy cohort, with case-
control of early menarche in
follow-up
n = 218 cases (menarche before
age 11.5 yrs) and 230 controls
0.0036-0.0039
ug/mL
(maternal)
Median (75th percentile) in
cases: 3.9(5.0)
controls: 3.6 (4.7) (p = 0.15)
OR (95% CI)
above versus below median 1.29 (0.86, 1.93)
per In-unit increase in PFOA 1.01 (0.61, 1.68)
Kristensen et al. 2013
Denmark
Pregnancy cohort, with follow-
up of 343 (79% of eligible)
daughters at age 20
Health questionnaire and
exams/hormone measurements
(for n = 254)
0.0036 ug/mL
(maternal)
Difference in age at menarche (months) by exposure group
low (0.001-0.003 ug/mL ) 0.0 (referent)
medium (0.003-0.0043 ug/mL) 0.9 (-3.0, 4.8)
high (0.0044-0.0198 ug/mL) 5.3 (1.3, 9.3)
continuous 1.01 (0.22, 1.89)
Peripubertal Exposure: High-Exposure Community
Lopez-Espinosa et al. 2011
United States (C8 Health Project)
n = 2,931 girl and 3,076 boy,
aged 8-18 yrs
Serf-reported menarche (girls)
and free or total testosterone
(boys)
Median 0.058
ug/mL
Prevalence of menarche in girls
OR 95% CI days delay
Ql: 1.0 (referent)
Q2: 1.01 (0.65-1.58) -4
Q3: 1.00 (0.64-1.58) -1
Q4: 0.75 (0.49-1.15) 69
Prevalence of delayed puberty in boys
OR 95% CI days delay
Ql: 1.0 (referent)
Q2: 0.54 (0.35-0.84) 142
Q3: 0.50 (0.32-0.87) 163
Q4: 0.57 (0.37-0.89) 130
Results were broadly similar when the analysis was based on estradiol
levels to define menarche or when the models included PFOA and
PFOS jointly, though significance was reduced in some comparisons.
Christensen et al. (2011) used data from a prospective cohort study in the United Kingdom to
perform a nested case-control study examining the association between age at menarche and
gestational exposure to perfluorinated chemicals, including PFOA and PFOS. The study
population from the Avon Longitudinal Study of Parents and Children included single-birth
female subjects who had completed at least two puberty staging questionnaires between the ages
of 8 and 13 years and whose mothers provided at least one analyzable prenatal serum sample. If
more than one serum sample was available, the earliest sample provided was used for analysis.
The study does not provide information as to when samples were collected. The females were
divided into two groups: those who experienced menarche prior to age 11.5 years (n = 218) and a
random sample of those who experienced menarche after age 11.5 (n = 230). Confounders such
as the mother's prepregnancy BMI, age at delivery, age at menarche, educational level, and the
child's birth order and ethnic background were included in linear and logistic regression models
used to analyze the data. The median maternal serum PFOA concentrations were 0.0039 and
0.0036 |ig/mL for the early menarche and nonearly menarche groups, respectively. The authors
noted a modest nonsignificant association between the odds of earlier menarche and prenatal
serum PFOA concentrations above the median. For all models, the CIs included the null value of
1.0. Similar results were obtained for PFOS.
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Effects of prenatal exposure to PFOA (and PFOS) on female and male reproductive function
was evaluated in 343 females and 169 males whose mothers participated in a cohort in 1988-
1989 (Kristensen et al. 2013; Vested et al. 2013). Maternal blood samples were collected during
week 30 of gestation. Follow-up was initiated in 2008 when the offspring were -20 years old.
Median serum PFOA level was 0.0036 |ig/mL for the mothers with daughters evaluated. In
adjusted regression analysis, daughters from mothers in the highest PFOA tertile had a
5.3-month later age at menarche (95% CI 1.3, 9.3) than those in the lowest tertile. No association
was found between prenatal exposure to PFOS and age of menarche. No statistically significant
relationships were found between PFOA (or PFOS) exposure and cycle length, reproductive
hormone levels, or number of follicles assessed by ultrasound (Kristensen et al. 2013).
Lopez-Espinosa et al. (2011) examined the association of serum PFOA concentration and the
age of puberty in exposed children of the Mid-Ohio Valley. Data from the C8 Health project
(e.g., sex steroid hormone levels, self-reported menarche status) along with detailed date of birth
information were used to determine age of puberty in males (n = 3,076) and females (n = 2,931)
aged 8-18 years. Serum PFOA concentrations were divided into quartiles: <0.0114, 0.0114-
0.023, >0.023-0.058, and >0.058 |ig/mL. Confounders such as age at survey, BMI, BMIz-score,
height, family income, ethnicity, smoking status, alcohol consumption, and date and time of
sample collection were included in the logistic regression models used to analyze the data. The
median PFOA concentrations were 0.026 and 0.020 |ig/mL in males and females, respectively.
No association between PFOA concentration and puberty was observed for males. Reduced odds
of having reached puberty was associated with higher PFOA exposure in females (OR=0.57,
95% CI 0.37-0.89). There were 130 days of delay between the highest and lowest quartile.
Reduced odds of experiencing menarche at a younger age (10-15 years) also was observed (OR
0.83, 95% CI 0.74-0.93). The results suggested that PFOA was associated with a later age of
menarche. PFOS was associated with delayed puberty in both males and females. The authors
expressed caution in interpretation of the data because of lack of serum PFOA concentration
prior to puberty, PFOA concentration having been measured after the attainment of puberty, and
lack of secondary sexual maturation data (i.e., physical, Tanner criteria, and biomarker
measurements).
Male reproductive effects. Joensen et al. (2009) examined the association between PFASs,
including PFOA, and testicular function in 105 Danish males who provided semen and blood
samples as part of reporting for the military draft in 2003. The males chosen for the study had the
highest testosterone concentrations (ranging from 30.1 to 34.8 nmol/L; n = 53; 18.2-24.6 years)
and lowest testosterone concentrations (ranging from 10.5 to 15.5 nmol/L; n = 52; 18.2-25.2
years). Regression models were used to analyze associations between PFOA and testicular
function. Median serum PFOA concentration was 0.0044, 0.0050, and 0.0049 |ig/mL in the high
testosterone, low testosterone, and combined groups, respectively. A nonsignificant negative
association was observed between serum PFOA concentration and semen volume, sperm
concentration, sperm count, sperm motility, or sperm morphology. No association was observed
between serum PFOA concentration and testosterone, estradiol, sex hormone-binding globulin
(SHBG), luteinizing hormone (LH), follicle-stimulating hormone (FSH), and inhibin B.
However, significantly fewer (p<0.05) morphologically normal sperm were seen in males with
high combined levels of PFOA/PFOS (6.2 million spermatozoa) than in males with low
PFOA/PFOS levels (15.5 million spermatozoa).
In a slightly expanded study, Joensen et al. (2013) investigated the associations between
PFASs, including serum PFOA concentration, and reproductive hormones and semen quality in
Perfluorooctanoic acid (PFOA) - May 2016 3-46
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247 healthy young Danish males (mean age 19.6 years). Serum samples were analyzed for PFOA
as well as total testosterone (T), estradiol, SHBG, LH, FSH, and inhibin-B. The mean PFOA
level was 0.0035 jig/mL. No associations were found between PFOA levels (or 12 other PFAS)
and any hormone level or semen quality parameters. PFOS levels were negatively associated
with testosterone.
An association between serum levels of seven PFASs and 35 semen quality parameters was
evaluated in 462 males enrolled in the LIFE study cohort (Buck Louis et al. 2015). The males
were from Michigan and Texas with a mean age of 31.8 years and mean PFOA levels 0.00429-
0.00509 |ig/mL. PFOA was significantly associated with a lower percentage of sperm with
coiled tails, an increased curvilinear velocity, and a slightly larger acrosome area of the head. In
total, six PFASs (including PFOA) were associated with changes in 17 semen quality endpoints.
Effects of prenatal exposure to PFOA (and PFOS on male reproductive function was
evaluated in 169 males whose mothers participated in a cohort in 1988-1989 (Vested et al.
2013). Maternal blood samples were collected during week 30 of gestation. Follow-up was
initiated in 2008 when the offspring were -20 years old. Median serum PFOA level was 0.0038
|ig/mL for mothers with sons evaluated. Multivariable regression models showed significant
negative trends for sperm concentration and total sperm count in association with in utero
exposure to PFOA. A 34% reduction in sperm concentration (95% CI 58, 5%) and a 34%
reduction in total count (95% CI 62, 12%) were estimated for the highest exposure tertile
compared with the lowest tertile. Maternal PFOA level also was positively associated with
higher FSH and LH levels in the sons. No associations were found between PFOA level and
percentage of progressive sperm, sperm morphology, semen volume, or testicular volume. PFOS
was not associated with any outcome (Vested et al. 2013).
3.1.1.9 Steroid Hormones
Occupational exposure studies. Olsen et al. (1998) examined several hormones, including
cortisol, estradiol, FSH, dehydroepiandrosterone sulfate, 17 gamma-hydroxyprogesterone (a
testosterone precursor), free testosterone, T, LH, prolactin, and SHBG in male workers at the
Cottage Grove, Minnesota, production plant for 1993 and 1995. This was the same population
used for the thyroid hormone study described above for 111 workers in 1993 and 80 in 1995.
Employees were placed into four exposure categories based on their serum PFOA levels:
0-1 |ig/mL, 1- < 10 |ig/mL, 10- < 30 jig/mL, and >30 jig/mL. Statistical methods used to
compare PFOA levels and hormone values included multivariable regression analysis, ANOVA,
and Pearson correlation coefficients. No association between serum PFOA and any hormone was
observed, but some trends were observed. When the mean measures of the various hormones
were compared by exposure categories, there was a statistically significant elevation in prolactin
(p = 0.01) in 1993 only for the 10 workers whose serum PFOA levels were between 10 and
30 |ig/mL compared to the lower two exposure categories.
Estradiol levels in the >30 |ig/mL PFOA group in both years were 10% higher than in the
other PFOA groups, but the difference was not statistically significant. These results were
confounded by estradiol being correlated with BMI (r = 0.41, p<0.001 in 1993, and r = 0.30,
p<0.01 in 1995). The authors postulated that the study might not have been sensitive enough to
detect an association between PFOA and estradiol because measured serum PFOA levels were
likely below the observable effect levels suggested in animal studies (e.g., 55 |ig/mL PFOA in
the CD rat). Only three employees in this study had PFOA serum levels that high. They also
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suggest that the higher estradiol levels in the highest exposure category could suggest a threshold
relationship between PFOA and estradiol.
In the Sakr et al. study (2007a) of 1,025 workers at the DuPont Washington Works facility in
West Virginia, an association was observed between serum PFOA and serum estradiol
(p = 0.017) and testosterone (p = 0.034) in male workers; however, circadian variations of
hormones were not taken into consideration during analysis. The biological significance of the
results is unknown.
Costa et al. (2009) found no association between serum PFOA concentration and estradiol or
testosterone in 53 male workers at a PFOA production plant in Italy based on medical
surveillance data collected between 2000 and 2007.
High-exposure community studies. Knox et al. (2011) examined the endocrine disrupting
effects of perfluorocarbons in females from the C8 Health Project by analyzing the relationship
between serum PFOA, serum estradiol concentration, and menopause onset. The population
included females over age 18 years (n = 25,957). Serum PFOA and estradiol concentrations were
determined from blood samples. Females who were pregnant; had had full hysterectomies; and
were taking any prescription hormones, selective estrogen receptor modulators, and/or fertility
agents were excluded from estradiol analysis. Serum PFOA concentrations were grouped into
quintiles (natural log-transformation)Ql = 0.00025-0.0112; Q2 = 0.0113-0.0198; Q3 =
0.0199-0.0367; Q4 = 0.0368-0.0849; and Q5 = 0.0850-22.412 |ig/mL. Estradiol analysis was
calculated by age group18-42 years, >42 < 51 years, and >51 < 65 years. Menopause was
determined by questionnaire. Menopause analysis was calculated by age group30-42 years,
>42 < 51 years, and >51 < 65 yearsand excluded those who reported having had
hysterectomies. Logistic regression models were adjusted for smoking, age, BMI, alcohol
consumption, and regular exercise. PFOA concentration in females who had had hysterectomies
was significantly higher than in females who had not had hysterectomies. Serum PFOA and
estradiol concentrations were not associated, while PFOS levels were negatively associated with
estradiol. The odds of attaining menopause analysis in the oldest group of females, showed that
all quintiles were significantly higher for all quintiles than the lowest, and in females between the
ages of 42 and 51 years, Q3, Q4, and Q5 were significantly higher than the lowest. PFOS also
was associated with increased odds of attaining menopause in women 42-51 years and >51
years. Data interpretation was limited by the cross-sectional study design and survey-reported
menopause without age or independent confirmation.
3.1.1.10 Neurodevelopment
High-exposure community studies. A subset of 321 children enrolled in the C8 Health Project
was assessed for neurobehavioral development 3-4 years after enrollment (Stein et al. 2013).
The children had serum samples collected at enrollment in 2005-2006 with the current follow-up
evaluation conducted in 2009-2010, when the children were 6-12 years old. Both the mother and
teacher completed surveys to elicit information on each child's executive function, attention
deficit hyperactivity disorder- (ADHD-) like behavior, and behavioral problems. Information on
family demographics and other health conditions of the child were included as confounders.
Linear regression was used to determine the association between PFOA levels and mother and
teacher reports.
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The median PFOA level was 0.0351 jig/mL with an IQR of 0.0158-0.0941 |ig/mL. When
comparing the highest to the lowest PFOA quartile, survey results from the mother for both
executive function and ADHD showed a favorable association for males, but an adverse
association for females. These findings were not replicated when males and females were
analyzed together or with results from the teacher surveys. No association was found between
PFOA levels and either mother or teacher scores for behavioral problems in females and males.
In 2012, the C8 Science Panel concluded that there is no probable link between PFOA
exposure and neurodevelopmental disorders in children, including attention deficit disorders and
learning disabilities. Their conclusion was based on epidemiology studies conducted by the panel
and other data available.
General population studies. Fei et al. (2008b) examined the association between plasma PFOA
concentration in pregnant females and motor and mental developmental milestones of their
children. The mothers self-reported the infant's fine and gross motor skills and mental
development at 6 and 18 months of age. There was no association between maternal plasma
PFOA concentration and Apgar score or between maternal plasma PFOA concentration and fine
motor skills, gross motor skills, or cognitive skills at 6 and 18 months of age. The children born
to females having higher plasma PFOA concentrations reached developmental milestones at the
same times as children born to females having lower plasma PFOA concentrations. The authors
concluded that there was no association between maternal early pregnancy levels of PFOA and
motor or mental developmental milestones in offspring. However, in children at 18 months,
mothers with higher PFOS levels were slightly more likely to report that their babies started
sitting without support at a later age.
A subset of the Danish National Birth Cohort was evaluated for an association between
prenatal PFAS exposure and the risk of cerebral palsy (Liew et al. 2014). A total of 156 cases of
cerebral palsy were identified and matched to 550 randomly selected controls. Stored maternal
plasma samples were analyzed for 16 PFAS and six compounds were quantifiable in >90% of
the samples. For the cerebral palsy cases and matched controls, median maternal PFOA levels
were 0.00456 and 0.00400 |ig/mL, respectively, for males and 0.00390 and 0.00404 |ig/mL,
respectively, for females. Per natural-log unit increase in maternal PFOA level, the risk of
developing cerebral palsy in males was significantly increased (RR=2.1; 95% CI 1.2, 3.6).
Positive associations were also found with PFOS and perfluoroheptane sulfonate. No association
was found between any PFAS level and risk of cerebral palsy in females.
Fei and Olsen (2011) examined the association between prenatal PFOA (and PFOS) exposure
and behavior or coordination problems in children at age 7. The children and their mothers were
part of the Danish National Birth Cohort. Behavioral problems were assessed using the Strengths
and Difficulties Questionnaire (SDQ), and coordination problems were assessed using the
Developmental Coordination Disorder Questionnaire (DCDQ) completed by the mothers. A total
of 787 mothers completed the SDQ and 537 completed the DCDQ for children aged 7.01-8.47
years (mean age 7.15 years). The mean maternal PFOA concentration was 0.0057 |ig/mL, and
PFOA levels were divided into quartiles:
-------
parental behavior problem as children did not show a positive association between prenatal
PFOA exposure and behavior or coordination problems. Overall, no significant association
between behavioral or coordination problems in children 7 years of age and prenatal PFOA (and
PFOS) exposure was found.
Similar to the above study, the association between maternal PFOA (and PFOS) levels and
offspring behavior and motor development was investigated in a subset of the Biopersistent
Organochlorines in Diet and Human Fertility study (INUENDO) birth cohort (H0yer et al.
2015a). Pregnant females were enrolled between May 2002 and February 2004 with a total of
1,106 mother-child pairs at follow-up between January 2010 and May 2012, when the children
were 7-9 years old. The study population consisted of 526 pairs from Greenland, 89 pairs from
Poland, and 491 pairs from Ukraine. Maternal blood samples for measurement of plasma PFOA
levels were taken any time during pregnancy. Behavior of children was assessed with SDQ
score, and logistic regression models were used in the analyses of PFOA tertile levels and
behavioral problems. Motor development was assessed with DCDQ score, and linear regression
was used for analyses. All analyses were performed on the entire cohort as well as by country,
except that not all analyses could be performed on the Polish subset because of the small number
of cases. The median maternal plasma PFOA level was 0.0014 |ig/mL for the combined
population and 0.0018, 0.001, and 0.0027 |ig/mL for the pregnant females from Greenland,
Ukraine, and Poland, respectively.
No associations were found between PFOA (and PFOS) levels and motor development score.
Total SDQ score was not associated with PFOA levels; however, the OR of having an abnormal
total SDQ score was 2.7 (95% CI 1.2, 6.3) for all groups combined. PFOS levels were associated
with higher total SDQ score only in Greenland. The highest PFOA tertile was associated with a
0.5-point higher hyperactivity score in both the combined analysis and in Greenland, but no
associations were found in Poland and Ukraine. The OR for hyperactive behavior in the
combined analysis was 3.1 (95% CI 1.3, 7.2) for the highest tertile compared to the lowest PFOA
tertile. In Greenland, the ORs for hyperactivity were increased for the middle (OR=5.4, 95% CI
1.1, 25.6) and highest (OR=6.3, 95% CI 1.3, 30.1) tertiles (H0yer et al. 2015a).
Hoffman et al. (2010) examined the associations between perfluorochemicals, including
PFOA, and diagnosis of ADHD using the NHANES data from 1999-2000 and 2003-2004. The
study population comprised 571 children aged 12-15 years, including those who had been
diagnosed as having ADHD (n = 48) and/or were taking ADHD medications (n = 21). Age,
gender, and race/ethnicity were included as covariates; and socioeconomic status, health
insurance coverage and having a routine health care provider, living with someone who smokes,
birth weight, admittance to a neonatal intensive care unit, maternal smoking, and preschool
attendance were confounders. Regression models were used to analyze the data. The median
serum PFOA level was 0.0044 |ig/mL and ranged from 0.0004 to 0.0217 ug/mL. Serum PFOA
was positively associated with parental report of ADHD (OR=1.12, 95% CI 1.01-1.23). The OR
for serum PFOA and parental report of ADHD and ADHD medication use was 1.19 (95% CI
0.95-1.49). Both PFOS and perfluorohexane sulfonate also were positively associated with
parentally reported ADHD. Data interpretation was limited by the cross-sectional study design,
random misclassification error resulting from using current PFOA levels as proxy measures of
etiologically relevant exposures, and other confounders not included in the available data.
Perfluorooctanoic acid (PFOA) - May 2016 3-50
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3.1.1.11 Postnatal Development
General population studies. Andersen et al. (2010) examined the association between maternal
plasma PFOA concentration and offspring weight, length, and BMI at 5 and 12 months of age.
The mothers (n = 1,010) reported the information during an interviews, and weight and length
measurements were used to calculate BMI. The median PFOA level was 0.0052 |ig/mL with a
range of 0.0005-0.0219 |ig/mL. Maternal plasma PFOA concentration was inversely associated
with weight at 5 months (P -30.2, 95% CI -59.3-1.1), BMI at 5 months (P -0.067, 95% CI -
0.1290.004), weight at 12 months (P -43.1, 95% CI -82.9-3.3 ), and BMI at 12 months in
male children (P -0.078, 95% CI -0.1440.011) in models adjusted for maternal age, parity,
prepregnancy BMI, smoking, gestational age at blood draw, socioeconomic status, and breast-
feeding. Similar inverse associations were found with PFOS. No associations were observed
between maternal plasma PFOA concentration and the endpoints for female children in the
adjusted models.
The latest report on the Danish National Birth Cohort evaluated an association between
maternal plasma PFOA levels and the children's BMI, waist circumference, and risk of being
overweight at 7 years of age (Andersen et al. 2013). From the subset of 1,400 females who
provided blood samples during their first trimester, children were included if they had weight
and height information (n = 811) or waist measurements (n = 804) at age 7 years. The median
PFOA level was 0.0053 |ig/mL with a range of 0.0005-0.0219 |ig/mL. Maternal PFOA levels
were inversely associated with all of the children's anthropomorphic endpoints, but statistical
significance was not attained and a dose response was not observed. Maternal PFOA (or PFOS)
did not affect the risk of being overweight in either males or females.
The association between maternal PFOA (and PFOS) levels and prevalence of offspring that
are overweight plus waist-to-height ratio >0.5 was investigated in a subset of the INUENDO
birth cohort (H0yer et al. 2015b). Pregnant females were enrolled between May 2002 and
February 2004 with a total of 1,022 mother-child pairs at follow-up between January 2010 and
May 2012, when the children were 7-9 years old. The study population consisted of 531 pairs
from Greenland and 491 pairs from Ukraine. Maternal blood samples for measurement of plasma
PFOA levels were taken at a mean gestational age of 24 weeks. Each child's weight and height
were measured and BMI calculated. All analyses were performed on the entire cohort as well as
by country.
The median maternal plasma PFOA level was 0.0018 |ig/mL in pregnant females from
Greenland and 0.0010 |ig/mL in pregnant females from Ukraine. No associations were found
between PFOA (and PFOS) levels and risk of being overweight in the combined analysis or in
Ukraine. In Greenland, the risk of being overweight was slightly increased only for females
(RR=1.81, 95% CI 1.04, 3.17). PFOA association with risk of having waist-to-height ratio >0.5
was slightly increased for the combined analysis (RR=1.30, 95% CI 0.97, 1.74), but statistical
significance was not attained. PFOS levels were significantly associated with waist-to-height
ratio >0.5 in the combined analysis (H0yer et al. 2015b).
Halldorsson et al. (2012) examined prenatal exposure to PFASs, including PFOA, and the
risk of being overweight at 20 years of age in a prospective study. A birth cohort consisting of
665 mother-offspring pairs was recruited from a midwife center in Aarhus, Denmark. Maternal
PFOA levels were measured in serum samples collected during week 30 of gestation for
assessment of in utero PFOA exposure and offspring anthropometry at 20 years. The median
PFOA concentration was 0.0037 ± 0.0020 |ig/mL with quartiles of 0.0024 ± 0.0006, 0.0033 ±
Perfluorooctanoic acid (PFOA) - May 2016 3-51
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0.0004, 0.0042 ± 0.0005, and 0.0058 ± 0.0019 |ig/mL. Three PFASs, including PFOS,
perfluorooctane sulphonamide, and perfluorononanoate, increased across quartiles of PFOA
concentration, but eight other PFASs did not. In covariate-adjusted analyses, female offspring
whose mothers were in the highest quartile had 1.6 kg/m2 higher BMI (95% CI: 0.6, 2.6) and
4.3 cm larger waist circumference (95% CI: 1.4, 7.3) than offspring whose mothers were in the
lowest quartile. Female offspring of mothers in the highest versus lowest PFOA quartile were
also more likely to be overweight [RR 3.1 (95% CI: 1.4, 6.9)] and to have a waist circumference
>88 cm at 20 years of age [3.0 (95% CI: 1.3, 6.8)]. Among female participants who provided
blood samples at clinical examination (n = 252), maternal PFOA concentration was positively
associated with insulin, leptin, and the leptin-adiponectin ratio; and inversely associated with
adiponectin levels. PFOA was not associated with being overweight or obesity in male offspring.
The other PFASs were not significantly associated with any endpoint after adjustment for PFOA.
Geiger et al. (2014b) used data from the NHANES to determine whether there was a
relationship between serum PFOA levels and hypertension in children. A total of 1,655
participants (aged 12-18 years) from the 1999-2000 and 2003-2008 cycles of the survey who
had PFOA measurements available were examined. Blood pressure was measured to determine
the presence of hypertension, and linear regression modeling was used to study the association
between increasing quartiles of serum PFOA and mean changes in systolic and diastolic blood
pressures. Mean PFOA level was 0.0044 ± 0.0001 jig/mL. No association was found between
serum PFOA (or PFOS) levels and hypertension in either unadjusted or multivariable-adjusted
analyses. Compared with the lowest quartile, the multivariable-adjusted OR (95% CI) of
hypertension in the highest quartile of exposure was 0.69 (0.41-1.17) (/"-trend >0.30).
3.1.1.12 Summary and Conclusions from the Human Epidemiology Studies
Numerous epidemiology studies have been conducted of workers, a large highly exposed
community (the C8 Health Project), and the general population to evaluate the association of
PFOA exposure to a variety of health endpoints. Health outcomes assessed include blood lipid
and clinical chemistry profiles, thyroid effects, diabetes, immune function, birth and fetal and
developmental growth measures, and cancer.
Serum lipids. The association between PFOA and serum lipids has been examined in several
studies in different populations. Cross-sectional and longitudinal studies in occupational settings
(Costa et al. 2009; Olsen et al. 2000, 2003; Olsen and Zobel 2007; Sakr et al. 2007a, 2007b;
Steenland et al. 2015) and in the high-exposure community (the C8 Health Project study
population) (Fitz-Simon et al. 2013; Frisbee et al. 2010; Steenland et al. 2009; Winquist and
Steenland 2014a) generally observed positive associations between serum PFOA and TC in
adults and children (aged l-< 18 yrs); most of these effect estimates were statistically
significant. Although exceptions to this pattern are present (e.g., some of the analyses examining
incidence of self-reported high cholesterol based on medication use [Steenland et al. 2015;
Winquist and Steenland 2014a]), the results are relatively consistent and robust. Similar
associations were seen in analyses of LDL, but were not seen with HDL. The range of exposure
in occupational studies is large (with means varying between 0.4 and > 12 |ig/mL), and the mean
serum levels in the C8 population studies were around 0.08 |ig/mL. Positive associations
between serum PFOA and TC (i.e., increasing lipid level with increasing PFOA) were observed
in most of the general population studies at mean exposure levels of 0.002-0.007 |ig/mL
(Eriksen et al. 2013; Fisher et al. 2013; Geiger et al 2014a; Nelson et al. 2010; Starling et al.
2014). The interpretation of results for these general population studies is limited, however, by
the moderately strong correlations (Spearman r > 0.6) and similarity in results seen for PFOS and
Perfluorooctanoic acid (PFOA) - May 2016 3-52
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PFOA. Additionally, many of the C8 studies do not appear to have controlled for the impact of
diet on serum lipids.
Liver disease and liver function. Few studies of the relationship between PFOA and liver
disease are available, but the C8 Health Project did not observe associations with hepatitis, fatty
liver disease, or other types of liver disease. In the studies of PFOA exposure and liver enzymes
(measured in serum), positive associations were seen. The results of the occupational studies
provide evidence of an association with increases in serum AST, ALT, and GOT, with the most
consistent results seen for ALT. The associations were not large and might depend on the
covariates in the models, including BMI, use of lipid lowering medications, and triglycerides
(Costa et al. 2009; Olsen et al. 2000, 2003; Olsen and Zobel 2007; Sakr et al. 2007a, 2007b).
Two population-based studies of highly exposed residents in contaminated regions near a
fluorochemical industry in West Virginia have evaluated associations with liver enzymes, and
the larger of the two studies reported associations of increasing serum In ALT and In GOT levels
with increasing serum PFOA concentrations (Emmett et al. 2006; Gallo et al. 2012). A cross-
sectional analysis of data from the NHANES, representative of the U.S. national population, also
found associations with In PFOA concentration with increasing serum ALT and In GGT levels.
Serum bilirubin was inversely associated with serum PFOA in the occupational studies. A
U-shaped exposure-response pattern for serum bilirubin was observed among the participants in
the C8 Health Project, which might explain the inverse associations reported for occupational
cohorts. Overall, an association of serum PFOA concentration with elevations in serum levels of
ALT and GGT has been consistently observed in occupational, highly exposed residential
communities, and the U.S. general population. The associations are not large in magnitude, but
indicate the potential of PFOA to affect liver function.
Immune function. Associations between prenatal, childhood, or adult PFOA exposure and risk
of infectious diseases (as a marker of immune suppression) have not been consistently seen,
although there was some indication of effect modification by gender (i.e., associations seen in
female children but not in male children) (Fei et al. 2010a; Granum et al. 2013; Looker et al.
2014; Okada et al. 2012). Three studies have examined associations between maternal and/or
child serum PFOA levels and vaccine response (measured by antibody levels) in children
(Grandjean et al. 2012; Granum et al. 2013) and in adults (Looker et al. 2014). The study in
adults was part of the high-exposure community C8 Health Project. A reduced antibody response
to one of the three influenza strains tested after subjects received the flu vaccine was seen with
increasing levels of serum PFOA; these results were not seen with PFOS. The studies in children
were conducted in general populations in Norway and in the Faroe Islands. Decreased vaccine
response in relation to PFOA levels was seen in these studies, but similar results also were seen
with correlated PFASs (e.g., PFOS).
Thyroid. Three large studies provide support for an association between PFOA exposure and
incidence or prevalence of thyroid disease in women or children, but not in men (Lopez-
Espinosa et al. 2012; Melzer et al. 2010; Winquist and Steenland 2014b). In addition,
associations between PFOA and TSH were seen in pregnant females with anti-TPO antibodies
(Webster et al 2014). In contrast, generally null associations were found between PFOA and
TSH in people who had not been diagnosed with thyroid disease.
Diabetes. No associations were observed between serum PFOA levels and type II diabetes
incidence rate in general or worker populations with mean serum PFOA up to 0.0913-0.113
|ig/mL (MacNeil et al. 2009; Steenland et al. 2015). PFOA was not associated with measures of
metabolic syndrome in adolescents or adults (Lin et al. 2009). However, one study found an
Perfluorooctanoic acid (PFOA) - May 2016 3-53
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increased risk for developing gestational diabetes in females with mean serum PFOA (measured
preconception) of 0.00394 |ig/mL (Zhang et al. 2015).
Fertility, pregnancy, and birth outcomes. There are no occupational exposure or general
population studies examining pregnancy-related hypertension and preeclampsia in relation to
PFOA exposure. The only data available come from the high-exposure C8 Health Project study
population. Several studies, using different designs and exposure measures, have examined that
outcome in this population (Darrow et al. 2013; Savitz et al. 2012a, 2012b; Stein et al. 2009).
There is a progressively greater refinement and reduction in misclassification (or exposure and
outcome) among this set of studies. Each of the studies provides some evidence of an association
between PFOA exposure and risk of pregnancy-induced hypertension or preeclampsia, with the
most robust findings from the methodologically strongest study (Darrow et al. 2013).
The association between PFOA and birth weight was examined in numerous studies. Most
studies measured PFOA using maternal blood samples taken in the second or third trimester or in
cord blood samples. Studies on the high-exposure C8 community population did not observe
associations between PFOA and either birth weight among term births or the risk of low birth
weight among all (singleton) births (Darrow et al. 2013; Nolan et al. 2009; Savitz et al. 2012a,
2012b; Stein et al. 2009). In contrast, several analyses of general populations indicate a negative
association between PFOA levels and birth weight (Apelberg et al. 2007; Fei et al. 2007;
Maisonet et al. 2012), while others did not attain statistical significance (Chen et al. 2012; Hamm
et al. 2010; Monroy et al. 2008; Washino et al. 2009). A meta-analysis of many of these studies
found a mean birth weight reduction of 19 g (95% CI: -30, -9) per each one unit (ng/mL)
increase in maternal or cord serum PFOA levels (Johnson et al. 2014). It has been suggested that
GFR can impact birth weight (Morken et al. 2014). Verner et al (2015) conducted a meta-
analysis based on PBPK simulations and found that some of the association reported between
PFOA and birth weight is attributable to GFR and that the actual association could be closer to a
7-g reduction (95% CI: -8, -6). Verner et al. (2015) showed that, in individuals with low GFR,
there are increased levels of serum PFOA and lower birth weights. While there is some
uncertainty in the interpretation of the observed association between PFOA and birth weight
given the potential impact of low GFR, the available information indicates that the association
between PFOA exposure and birth weight for the general population cannot be ruled out. In
humans with low GFR (which includes females with pregnancy-induced hypertension or
preeclampsia), the impact on body weight is likely due to a combination of the low GFR and the
serum PFOA.
Two studies examined development of puberty in females in relation to prenatal exposure to
PFOA as measured through maternal or cord blood samples in follow-up of pregnancy cohorts
conducted in England (Christensen et al. 2011) and in Denmark (Kristensen et al. 2013). The
results of these two studies are conflicting, with no association (or possible indication of an
earlier menarche seen with higher PFOA) in Christensen et al. (2011), and a later menarche seen
with higher PFOA in Kristensen et al. (2013). Another study examined PFOA exposure
measured concurrently with the assessment of pubertal status (Lopez-Espinosa et al. 2011). An
association between later age at menarche and higher PFOA levels was observed, but the
interpretation of this finding is complicated by the potential effect of puberty on the exposure
biomarker levels (i.e., reverse causality).
Perfluorooctanoic acid (PFOA) - May 2016 3-54
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Studies found a positive association with ADHD in children in the highly exposed
community (Stein et al. 2013) and the general population (Hoffman et al. 2010). No other
behavior endpoints in children were associated with maternal PFOA levels in either population.
Limited data suggest a correlation between higher PFOA levels (>0.02 jig/mL) in females
and decreases in fecundity and fertility (Fei et al. 2009; Velez et al. 2015), but there are no clear
effects of PFOA on male fertility endpoints (0.0035-0.005 |ig/mL) (Joensen et al. 2009, 2013).
C8 Science Panel conclusions. As part of the C8 Health Project, the C8 Science Panel used
epidemiological and other data available to them to assess probable links between PFOA
exposure and disease (C8 Science Panel 2012). Analyses conducted by the C8 Science Panel
used historical serum PFOA estimates over time, which were developed based on estimated
intake of contaminated drinking water. The panel concluded that a probable link existed between
PFOA exposure and ulcerative colitis, high cholesterol, pregnancy-induced hypertension, and
thyroid disease.
The C8 Science Panel found no probable link between PFOA exposure and multiple other
conditions, including birth defects, other autoimmune diseases (e.g., rheumatoid arthritis, lupus,
type 1 diabetes, Crohn's disease, MS), type II diabetes, high blood pressure, coronary artery
disease, infectious disease, liver disease, Parkinson's disease, osteoarthritis, neurodevelopmental
disorders in children (e.g., ADHD, learning disabilities), miscarriage or stillbirth, chronic kidney
disease, stroke, asthma or COPD, and preterm birth or low birth weight (C8 Science Panel 2012).
3.1.2 Cancer
Occupational exposure studies. Several occupational studies examining cancer mortality have
been conducted at 3M's Cottage Grove facility in Minnesota and at the DuPont Washington
Water Works plant in West Virginia. These studies have focused on kidney, bladder, liver,
pancreatic, testicular, prostate, thyroid, and breast cancers. For cancers with a high survival rate
(i.e., bladder, kidney, prostate, testicular, thyroid, and breast cancer), studies that use mortality
data provide a more limited basis for drawing conclusions than studies that use incidence data.
The discussion in this section summarizes the design and results of the available studies,
focusing on the most recent update of occupational cohorts. Table 3-12 presents results for
studies of kidney and testicular cancer.
Raleigh et al. (2014) is the latest update of the analyses of mortality in the 3M Cottage Grove
workers, previously analyzed in Lundin et al. (2009) and Gilliland and Mandel (1993). Raleigh
et al. (2014) followed 4,668 Cottage Grove workers through 2008, using an improved exposure
reconstruction method and adding a nonexposed worker referent group from a different 3M
plant. In addition to the mortality data, incidence data based on state cancer registries also were
included. Exposure estimates for inhalation exposures were calculated from work history records
and industrial hygiene monitoring data; blood levels were not included. No associations were
found between PFOA exposure and the risk of dying from any cancer type (see Table 3-12 for
bladder, kidney, and testicular cancer results). The mean age of the workers was 29 years at the
start of employment and 63 years at the end of follow-up.
Perfluorooctanoic acid (PFOA) - May 2016 3-55
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Table 3-12. Summary of PFOA Epidemiology Studies of Kidney and Testicular Cancer
Reference and Study Details
Analysis Group
Kidney
Testicular
Occupational Settings
Raleigh etal. 2014
3M, Minnesota
n = 4,668, follow-up through 2008
Mean age: 29 yrs at start of
employment
Mortality and incidence
Comparison based on another (non-
PFOA) 3M plant in Minnesota (n =
4,359)
Cumulative exposure level based on
industrial hygiene data (air
monitoring), and PFOA production
levels
[update of Gilliland and Mandel 1993
and Lundin et al. 2009]
All (Minnesota referent)
By quartile, mortality
analysis
1 up to 0.000026 ug/m3-yr
2 up to 0.00014
3 up to 0.00073
4 maximum not reported
By quartile, incidence
analysis
1 up to 0.000029 ug/m3-yr
2 up to 0.00015
3 up to 0.00079
4 maximum not reported
0.53 (0.20,
Mortality
0.32 (0.01,
0.74 (0.09,
1.66 (0.08,
0.42 (0.01,
Incidence
1.07 (0.36,
1.07 (0.36,
0.98 (0.33,
0.73 (0.21,
1.16)(n=6)
1.77)(n=l)
2.69) (n= 2)
2.38) (n = 2)
2.34) (n=l)
3.16)(n=4)
3.17)(n=4
2.92) (n= 4)
2.48) (n= 4)
n = 5 incident cases
reported; no further
analysis done
Steenland and Woskie 2012
DuPont Washington Water Works,
West Virginia
n = 5,791, follow-up through 2007
Mortality
[update of Leonard et al. 2008, which
did not include analysis by cumulative
exposure]
[Steenland and Woskie 2012 examined
incidence of bladder, colorectal, and
prostate cancers, and of melanoma]
DuPont referent (plants from
8 surrounding states)
U.S. referent
Cumulative exposure (ppm-
yrs)
0 - < 904
904 - < 1,520
15,20 - < 2,720
>2720
1.28(0.66, 2.24) (n =
12)
1.09(0.56, 1.90) (n =
12)
1.07(0.02, 3.62) (n=l)
1.37(0.28, 3.99) (n= 3)
- (0.0, 1.42) (n=0)
2.66(1.15, 5.24) (n = 8)
1.80 (0.05, 10.03) (n
= 1)
High-Exposure Community
Vieiraetal. 2013
C8 Health Project population (Ohio and
West Virginia)
Incidence
Modeled estimates for 1951-2008
using residence at time of diagnosis
and emissions data and environmental
characteristics
Total for 6 water districts
(median serum level ranged
from 5 to 125 ug/1)
Annual serum levels (ug/1);
assumed 10-year residence
and 10-year latency (Ohio)
Unexposed
Low: 3.7 - 12.
Medium: 12.9 - 30.7
High: 30.8 - 109
Very high: > 100
1.1(0.9, 1.4)(n=94)
1.0 (referent)
0.8(0.4, 1.5)(n=ll)
1.2(0.7, 2.0) (n= 17)
2.0(1.3, 3.2) (n = 22)
2.0(1.0,3.9) (n=9)
(0.6, 1.8) (n= 18)
(referent)
0.2 (0.0, 1.6) (n=l)
0.6(0.2, 2.2) (n= 3)
0.3 (0.0, 2.7) (n= 1)
2.8(0.8, 9.2) (n= 6)
Barry etal. 2013
C8 Health Project population (Ohio and
West Virginia)
Case-control (n varies by cancer)
Incidence
Modeled estimates for 1951-2008
using individual-level data on
residential history, drinking water
source, tap water consumption,
emissions data, environmental
characteristics, water pipe installation,
PK data, and workplace water
consumption (and for workers,
workplace exposure based on job
exposure matrix and modeling using
serum samples from 1979-2004 and
job history data.
Full sample
Cumulative exposure,
quartiles (cutpoints based on
cancer-specific case
distribution; approximate
midpoints)
1 (30-50 ug/mL-yr)
2 (90-200 ug/mL-yr)
3 (800-1400 ug/mL-yr)
4 (100,000 ug/mL-yr)
1.09(0.97, 1.21) (n =
105)
(referent)
0.99 (0.53, 1.85)
1.69 (0.3, 3.07)
1.43 (0.76, 2.69)
trend p = 0.34
community cohort; HR
= 1.0,0.94, 1.08,1.50,
trend p = 0.02;
worker cohort HR = 1.0,
1.22, 3.27, 0.99, trend p
= 0.42
1.28(0.95, 1.73) (n =
17)
1.0 (referent)
0.87(0.15,4.88)
1.08 (0.20, 5.90)
2.36 (0.41, 13.7)
trend p = 0.02
15 of the cases from
the community
sample; HR= 1.0,
0.98, 1.54, 4.66,
trend p = 0.02
Perfluorooctanoic acid (PFOA) - May 2016
3-56
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Steenland and Woskie (2012) updated the cohort study by Leonard et al. (2008) of
employees at the DuPont Washington Works plant in West Virginia (see Table 3-12 for bladder,
kidney, and testicular cancer results). This study included 5,791 individuals who had worked at
the DuPont West Virginia plant for at least 1 year between 1948 and 2002. Mean duration of
employment was 19 years. Deaths through 2008 were ascertained through either the National
Death Index or death certificate data. Exposure quartiles were assessed by estimated cumulative
annual serum levels based on blood samples from 1,308 workers taken during 1979-2004 and
time spent in various job categories (ppm-years). Referent groups included both nonexposed
DuPont workers in the same region and the U.S. population. Overall, the mean cumulative
exposure was 7.8 ppm-years and the estimated average annual serum level was 0.35 |ig/mL. A
significant positive trend was found for kidney cancer with the SMR=2.66 (n = 8; 95% CI 1.15,
5.24) for workers in the highest quartile. The most recent report on the same cohort included
6,026 workers evaluated for disease incidence, based on self-report with validation from medical
records (Steenland et al. 2015). Lifetime serum cumulative dose was estimated by combining
occupational and nonoccupational exposures. Median measured serum level was 0.113 |ig/mL
based on samples collected in 2005. Bladder cancer incidence (n = 29 cases) decreased with
increased PFOA levels (RR 1.0, 0.55, 0.47, and 0.31 across quartiles, trend p = 0.03). Prostate
cancer risk increased in Ql compared to Q2 (n = 1.92), and remained at this level in the
remaining quartiles (RR 1.89 and 2.15 in Q3 and Q4, respectively, trend p = 0.10).
Cholecystokinin (CCK) is a peptide hormone that stimulates the digestion of fat and protein,
causes the increased production of hepatic bile, and stimulates contraction of the gall bladder.
Research in rats suggests that pancreas acinar cell adenomas observed in rodents might be the
result of increased CCK levels secondary to blocked bile flow (Obourn et al. 1997). CCK was
measured in male workers (n = 74 males) at the 3M's Cottage Grove plant in 1997 as part of the
medical surveillance program (Olsen et al. 1998, 2000). Employees' serum PFOA levels were
stratified into three categories (<1, 1- <10, and >10 ppm). The mean CCK values for the three
PFOA categories were 33.4, 28.0, and 17.4 pg/mL, respectively. The means in the two serum
categories < 10 ppm were at least 50% higher than in the > 10 ppm category. A statistically
significant negative association between mean CCK levels and the three PFOA categories was
observed (p = 0.03). A multiple regression model of the natural log of CCK and serum PFOA
levels continued to display a negative association after adjusting for potential confounders. As
stated previously (Olsen et al. 2000), no abnormal liver function, hypolipidemia, or cholestasis
was observed in the workers. The authors suggested that the lack of a positive association
between PFOA and CCK in workers could have resulted from serum PFOA levels too low to
cause an increase in CCK provided that the same mechanism that increases CCK levels in
rodents exists in humans.
High-exposure community studies. Vieira et al. (2013) investigated the relationship between
PFOA exposure and cancer among the residents living near the DuPont plant in Parkersburg,
West Virginia. This analysis included incident cases of 18 cancers diagnosed from 1996-2005 in
five Ohio counties and eight West Virginia counties that included public water districts
contaminated with PFOA. The dataset included 7,869 cases from Ohio geocoded to residence
and 17,238 cases from West Virginia linked to water district. Exposure levels and serum PFOA
concentrations were estimated based on residence at time of diagnosis, using modeled data based
on previous work in the C8 study population (Shin et al. 2011). Individual-level exposure was
categorized as very high, high, medium, low, or unexposed based on serum concentrations of
>0.110 |ig/mL, 0.0308-0.109 |ig/mL, 0.0129-0.0307 |ig/mL, 0.0037-0.0129 |ig/mL, and
unexposed (background levels not given), respectively. Logistic regression was applied to
Perfluorooctanoic acid (PFOA) - May 2016 3-57
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individual-level data to calculate ORs and CIs for each cancer category. Data were first analyzed
by water district. The adjusted ORs were increased for testicular cancer and for kidney cancer
(OR: 5.1, 95% CI: 1.6, 15.6; n = 8 and OR: 1.7, 95% CI: 0.4, 3.3; n = 10, respectively) in the
Little Hocking water district and for kidney cancer (OR: 2.0, 95% CI: 1.3, 3.1; n = 23) in the
Tuppers Plains water district. Both districts are in Ohio. Residents of Little Hocking also had
increased OR for non-Hodgkin lymphoma (OR: 1.6, 95% CI: 0.9, 2.8; n = 14) and prostate
cancer (OR: 1.4, 95% CI: 0.9, 2.3; n = 36). The analysis by exposure level for kidney and
testicular cancers is shown in Table 3-12. Kidney cancer was positively associated with very
high and high exposure categories (OR: 2.0, 95% CI: 1.0, 3.9; n = 9 and OR: 2.0, 95% CI: 1.3,
3.2; n = 22, respectively), while ORs for medium and low exposure categories were close to the
null when compared to the unexposed category. The largest OR was for testicular cancer with the
very high exposure category (OR: 2.8, 95% CI: 0.8, 9.2; n = 6), but the estimate was imprecise
because of the small numbers. ORs for the other exposure categories were all <1.0. Ovarian
cancer, non-Hodgkin's lymphoma, and prostate cancer were positively associated with the very
high exposure category, but showed weaker or negative associations for the other exposure
categories (Vieira et al. 2013).
Barry et al. (2013) extended the study of cancer incidence in the C8 Health Project
population in an analysis of data from 32,254 study participants; there is some overlap in the
cases included in Vieira et al (2013) and in Barry et al. (2013). The cohort included 3,713 current
and former DuPont Washington Works employees, but results for this subset were limited by the
small sample size for cancers of interest. Median serum PFOA levels, measured in 2005-2006 at
enrollment in C8, were 0.024 and 0.113 |ig/mL for community and worker populations,
respectively. A proportional hazard regression model was run for each cancer type with the
cancer as the outcome, time-varying cumulative PFOA serum concentration as the independent
variable, and age as the time scale. Cumulative PFOA serum concentrations were estimated
based on historical regional monitoring data and individual residential histories. Self-reported
cancers were validated through a cancer registry or medical record. Confounders included
smoking, alcohol consumption, gender, education, and 5-year birth year period. Testicular cancer
risk was significantly increased with an increase in the log of estimated cumulative PFOA serum
level (HR: 1.34, 95% CI: 1.00, 1.79; n = 17). Using estimated cumulative PFOA serum
concentration quartiles, a significant monotonic trend was found for testicular cancer. Slight
nonsignificant increases were seen for kidney cancer (HR: 1.10, 95% CI: 0.98, 1.24; n = 105)
and for thyroid cancer (HR: 1.10, 95% CI: 0.95, 1.26; n = 86) (Barry et al. 2013).
Members of the C8 Health Project were evaluated for an association between serum PFOA
levels and incidence of colon or rectal cancer (Innes et al. 2014). This cross-sectional study
compared serum PFOA (and PFOS) levels at enrollment with diagnosis of primary colorectal
cancer; 47,151 cancer-free adults and 203 cases were included. Serum PFOA levels ranged from
<0.0005 to 22.4 |ig/mL, with an average of 0.0866 |ig/mL. An inverse relationship was found
between PFOA level and diagnosis of colorectal cancer with OR = 0.64 (95% CI 0.44, 0.94;
highest to lowest quartile, p for trend = 0.002). A concentration-related inverse association also
was found between PFOS and colorectal cancer.
In 2012, the C8 Science Panel concluded that there is a probable link between exposure to
PFOA and testicular and kidney cancer, but no other types of cancers. Their conclusion was
based on the studies presented above, other epidemiology studies on cancer incidence in the mid-
Ohio population, worker cohorts, and published data. Panel studies addressed 21 different
Perfluorooctanoic acid (PFOA) - May 2016 3-58
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categories of cancer and looked for positive trends with increasing exposure as measured by
cumulative serum levels.
General population studies. Eriksen et al. (2009) examined the association between plasma
PFOA concentration and the risk of cancer in the general Danish population. The study
population was chosen from individuals (aged 50-65 years) who had enrolled in the prospective
Danish cohort Diet, Cancer and Health study between December 1, 1993, and May 31, 1997. The
Danish Cancer Registry and Danish Pathology Data Bank were used to identify cancer patients
diagnosed between December 1, 1993, and July 1, 2006. The cancer patients (n = 1,240)
consisted of 1,111 males and 129 females whose median age was 59 years and who had prostate
cancer (n = 713), bladder cancer (n = 332), pancreatic cancer (n = 128), or liver cancer (n = 67).
The individuals (n = 772) in the subcohort comparison group were randomly chosen from the
cohort study and consisted of 680 males and 92 females whose median age was 56 years. The
participants each answered a questionnaire upon enrollment in the cohort study, and data on
known confounders were obtained from the questionnaires. The plasma PFOA concentrations,
based on blood samples provided by cancer patients at enrollment (1993-1997) were as follows:
males 0.0068 |ig/mL, females 0.0060 jig/mL, prostate cancer 0.0069 |ig/mL, bladder cancer
0.0065 |ig/mL, pancreatic cancer 0.0067 |ig/mL, and liver cancer 0.0054 |ig/mL. The plasma
PFOA concentrations for the subcohort comparison group were 0.0069, 0.0054, and 0.0066
|ig/mL for males, females, and combined, respectively. IRRs, crude and adjusted for
confounders, did not indicate an association between plasma PFOA concentration and prostate,
bladder, pancreatic, or liver cancer (see Table 3-12 for bladder cancer results). The plasma
PFOA levels in the population were lower than those observed in occupational cohorts. This
study is novel in that it is the first to examine PFOA levels and cancer in the general population.
A subset of females enrolled in the Danish National Birth Cohort was evaluated for an
association between plasma PFOA levels (as well as 15 other PFASs) measured during
pregnancy and risk of breast cancer during a follow-up period of 10-15 years (Bonefeld-
J0rgensen et al. 2014). A total of 250 females diagnosed with breast cancer were matched for age
and parity with 233 controls. The mean PFOA level in the controls was 0.0052 |ig/mL while
levels in the cases were divided into quintiles ranging from <0.0037 up to >0.0065 |ig/mL. No
association was found between PFOA levels and breast cancer risk. A weak positive association
was found only with perfluorooctane sulfonamide.
Hardell et al. (2014) investigated an association between prostate cancer and levels of
perfluoroalkyl acids (PFAAs) in whole blood. Patients with newly diagnosed prostate cancer
(n = 201) had median PFOA levels of 0.002 |ig/mL while the case-control group (n = 186) had a
median level of 0.0019 jig/mL. PFOA levels were not associated with higher risks of prostate
cancer when compared to controls or when analyzed according to Gleason score (pathology
grade) and prostate-specific antigen. A significantly higher risk for prostate cancer was found for
PFOA levels above the median combined with a first-degree relative with prostate cancer,
indicating a genetic risk factor.
Two studies found no differences in blood and tissue PFOA levels between cancer and
noncancer patients; the types of cancer in the patients were not defined. Vassiliadou et al. (2010)
found that median serum PFOA concentrations among 40 cancer patients (0.00227 |ig/mL in
males; 0.00185 |ig/mL in females) were similar to two control groups (0.00314 and 0.00181
|ig/mL in males; 0.0017 and 0.00171 |ig/mL in females). Yeung et al. (2013) found similar
PFOS levels in serum and liver tissue between controls and those with hepatocellular carcinoma.
Median serum levels in controls (n = 25) and patients with liver cancer (n = 24) were 0.00234
Perfluorooctanoic acid (PFOA) - May 2016 3-59
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and 0.0025 |ig/mL, respectively, and liver tissue were 0.506 (n = 9) and 0.495 (n = 12) ng/g,
respectively.
3.1.2.1 Summary and Conclusions from the Human Cancer Epidemiology Studies
Evidence of carcinogenic effects of PFOA in epidemiology studies is based on studies of
kidney and testicular cancer. These cancers have relatively high 5-year survival rates of 73% for
kidney cancer and 95% for testicular cancer (based on National Cancer Institute [NCI]
Surveillance, Epidemiology, and End Results data for 2005-2011). Thus studies that examine
cancer incidence are particularly useful for these types of cancer. The high-exposure community
studies also have the advantage for testicular cancer of including the age period of greatest risk,
as the median age at diagnosis is 33 years. The two occupational cohorts in Minnesota and West
Virginia (most recently updated, respectively, in Raleigh et al. 2014 and Steenland and Woskie
2012) do not support an increased risk of these cancers, but each of them is limited by a small
number of observed deaths and incident cases. Two studies involving members of the C8 Health
Project showed a positive association between PFOA levels (mean at enrollment of 0.024
|ig/mL) and kidney and testicular cancers (Barry et al. 2013; Vieira et al. 2013). There is some
overlap in the cases included in these studies. None of the general population studies examined
kidney or testicular cancer, but no associations were found in the general population between
mean serum PFOA levels up to 0.0866 |ig/mL and colorectal, breast, prostate, bladder, or liver
cancer (Bonefeld-J0rgensen et al. 2014; Eriksen et al. 2009; Hardell et al. 2014; Innes et al.
2014).
As part of the C8 Health Project, the C8 Science Panel (2012) concluded that a probable link
existed between PFOA exposure and testicular and kidney cancer.
A group of independent toxicologists and epidemiologists critically reviewed the
epidemiological evidence for cancer based on 18 studies of occupational exposure to PFOA and
general population exposure with or without coexposure to PFOS. The project was funded by
3M, but the company was not involved in the preparation or approval of the report. The authors
evaluated the published studies based on the study design, subjects, exposure assessment,
outcome assessment, control for confounding, and sources of bias. They followed the Bradford
Hill guidelines on the strength of the association, consistency, plausibility, and biological
gradient in reaching their conclusion. They found a lack of concordance between community
exposures and occupational exposures one or two magnitudes higher than those for the general
population. The discrepant findings across the study populations were described as likely due to
chance, confounding, and/or bias (Chang et al. 2014).
3.2 Animal Studies
Acute and short-term studies in monkeys, rats, and mice provide data on systemic toxicity
and MoA. Subchronic studies in monkeys and rats found decreased body weight, increased liver
weight accompanied by microscopic lesions, and decreased serum cholesterol. The most
prominent microscopic lesion of the liver in both monkeys and rats was centrilobular
hepatocellular hypertrophy. Data from studies of inhalation and dermal exposures are limited.
Chronic exposure studies were conducted in monkeys, rats, and mice providing information
on tumor incidences for both rats and mice. Effects on development and reproduction were found
in both rats (a 2-generation study) and mice (male fertility) and included developmental delays
and increased neonatal mortality. Many developmental studies focused on the impact of
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gestational/lactational exposure on mammary gland development and effects observed in
offspring at maturity.
3.2.1 Acute Toxicity
Oral Exposure
Dean and Jessup (1978) reported an oral lethal dose for 50% of animals (LDso) of 680 mg/kg
and 430 mg/kg PFOA for male and female CD rats, respectively. Glaza (1997) reported an oral
LDso of greater than 500 mg/kg in male Sprague-Dawley rats and between 250 and 500 mg/kg in
females. Gabriel (1976a) reported an oral LDso of less than 1,000 mg/kg for male and female
Sherman-Wistar rats. According to the Hodge Sterner Scale, these LDso values suggest that
PFOA can be classified as moderately toxic after acute oral exposures.
Rigden et al. (2015) exposed groups of five male Sprague-Dawley rats to doses of 0, 10, 33,
and 100 mg/kg/day for 3 days and maintained them for 4 additional days with daily body weight
measurement and overnight collection of urine. Following the recovery period, the animals were
sacrificed with collection of serum samples for analysis. Major organs were weighed, and the
liver homogenized. The serum samples, liver homogenate, and supernatant were kept frozen at -
80°C until they were analyzed. Phase I and II drug metabolizing enzymes and palmitoyl-
coenzyme A (-CoA) oxidase were measured in the liver homogenate. Urine was analyzed for
malondialdehyde (MDA) and 8-hydroxydeoxyguanine. The results for PFOA were compared
with those for 100-mg/kg/day doses of di(2-ethylhexyl) phthalate (DEHP) and fenofibrate,
known inducers of PPARa.
There was a dose-related statistically significant increase in palmitoyl-CoA oxidase and liver
weight at all PFOA doses. The palmitoyl-CoA increase was not significant for DEHP and
fenofibrate with 100-mg/kg doses; liver weight increased significantly for fenofibrate but not
DEHP. The only serum parameter that showed a significant dose-related response with PFOA
was a decrease in uric acid compared to controls. Serum was analyzed for several minerals,
proteins, enzymes (e.g., ALP, AST, ALT), glucose, cholesterol, and triglycerides. Phase I drug
metabolizing enzyme activities (ethoxyresorufin-O-deethylase [EROD] and pentoxyresorufin-O-
depentilase [PROD]) were significantly increased at the 100-mg PFOA/kg/day dose, and
glutathione-S-transferase (GST) activity was significantly decreased at the two highest doses, but
not in a dose-related fashion. UDP-glucuronyltransferase (UDP-GT) was significantly lower than
controls at all doses, but the changes did not demonstrate a dose-related response. There were no
dose-related significant changes for the other analytes. The 10-mg PFO A/kg/day dose
administered for 3 days was a LOAEL for effects on the liver associated with PPARa activation
and for a decrease in serum uric acid. PFOA at 10 mg/kg/day for 3 days had a stronger impact on
liver weight and palmitoyl-CoA activation than 100 mg/kg/day of DEHP and fenofibrate for the
same exposure duration (Table 3-13). The 10-mg/kg/day dose was a LOAEL for liver effects
usually associated with PPARa activation. The PFOA response was stronger than that for a
100-g/kg/day dose for the two known activators of PPAR-a.
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Table 3-13. Comparison of PPAR-a Related Effects in Rats for PFOA, DEHP, and
Fenofibrate after a 3-day Exposure
Chemical
Control
PFOA
PFOA
PFOA
DEHP
Feno-
fibrate
Dose
mg/kg/
day
0
10
33
100
100
100
Liver wt.
g
4.28 ±0.20
5.73 ±0.29*
6.40 ± 0.20*
6.62 ± 0.47*
4.14 ±0.34
5.73 ±0.24*
Palm. CoA
abs/min/g
prot
1.02 ±0.37
3. 17 ±0.65*
4.89 ±0.79*
6.11 ±1.51*
1.40 ±1.09
1.71 ±0.58
EROD
nmol/min/
mg prot
0.066 ± 0.022
0.096 ± 0.024
0.080 ±0.015
0.113 ±0.025*
0.060 ±0.012
0.060 ±0.016
PROD
nmol/min/
mg prot
0.045 ±0.012
0.080 ± 0.024
0.078 ±0.031
0.107 ±0.029*
0.039 ±0.031
0.046 ± 0.020
UDP-GT
nmol/min/
mg prot
1.69 ±0.12
0.88 ±0.09*
1.00 ±0.14*
1.12 ±0.20*
1.45 ±5031
1.09 ±0.14*
GST
nmol/min/
mg prot
1.21±0.11
1.11 ±0.09
0.88 ±0.12*
0.94 ±0.19*
1.27 ±0.11
0.94 ±0.13*
Notes: Mean ±SD
* Significant ANOVA followed by Dunnett post-hoc test p < 0.05.
Inhalation Exposure
Rusch (1979) reported no mortality in male or female Sprague-Dawley rats following
inhalation exposure to 186,000 mg/m3 PFOA for 1 hour. Kennedy et al. (1986) reported a 4-hour
lethal concentration for 50% of animals (LCso) of 980 mg/m3 for groups of six male rats exposed
to PFOA as a dust in air. As reported in a later publication (Kennedy et al. 2004), body weight
loss, irregular breathing, and red discharge around the nose and eyes were observed. Corneal
opacity and corrosion were seen at concentrations greater than or equal to 810 mg/m3.
Dermal/Ocular Exposure
The dermal LDso in New Zealand White rabbits was determined to be greater than
2,000 mg/kg (Glaza 1995). Kennedy (1985) determined a dermal LDso of 4,300 mg/kg for
rabbits, 7,000 mg/kg for male rats, and 7,500 mg/kg for female rats. The animals lost body
weight and exhibited lethargy, labored breathing, diarrhea, and severe skin irritation (Kennedy et
al. 2004). PFOA is an ocular irritant in rabbits when the compound is not washed from the eyes
(Gabriel 1976b), but is not an irritant in rabbits when washed from the eye (Gabriel 1976c).
Markoe (1983) found PFOA to be a skin irritant in rabbits, while Gabriel (1976d) did not
conclude that PFOA is a skin irritant.
3.2.2 Short-Term Studies
Oral Exposure
Monkey. In a range-finding study, Thomford (2001) administered PFOA to male cynomolgus
monkeys as an oral capsule containing 0, 2, and 20 mg/kg/day PFOA for 4 weeks. There were
three monkeys in the 2- and 20-mg/kg/day groups and one monkey in the control group. Animals
were observed twice daily for mortality and moribundity and were examined at least once daily
for signs of poor health or abnormal behavior. Body weights were recorded weekly and food
consumption was assessed qualitatively. The monkeys were fasted overnight and blood samples
were collected 1 week prior to the start of the study and on day 30 for measurement of serum
PFOA, clinical hematology, and clinical chemistry, plus analysis for hormones (estradiol,
estrone, estriol, TSH, total and FT3, and total and FT4). Blood samples also were collected from
each animal on day 2 (approximately 24 hours after the first dose) for clinical chemistry
measurements.
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At scheduled necropsy, liver samples were collected for palmitoyl-CoA oxidase activity (a
biomarker for peroxisome proliferation) and serum PFOA. Liver, testes, and pancreas were
collected and assayed for cell proliferation using antibodies to proliferating cell nuclear antigen
(PCNA). Bile was collected from each animal for measurement of bile acid. The adrenals, liver,
pancreas, spleen, and testes from each animal were examined microscopically.
All animals survived to scheduled sacrifice. There were no clinical signs of toxicity in the
treated groups and there was no effect on body weight. Low or no food consumption was
observed for one animal given 20 mg/kg/day. There were no effects on the hormones measured
with the exception of estrone, which was notably lower in the 2- and 20-mg/kg/day PFOA
groups. There was no evidence of peroxisome proliferation or cell proliferation in the liver,
testes, or pancreas of the treated monkeys. No adverse effects were noted in either the gross or
clinical pathology evaluations. Under the conditions of this study, the NOAEL was 20 mg/kg and
no LOAEL was established.
Rat. Pastoor et al. (1987) dosed male Crl:CD (SD) BR rats (n = 6 per group) for 1, 3, and 7 days
with 0 and 50 mg PFOA/kg. Liver sections were collected at necropsy and stained with
hematoxylin and eosin. Sections also were examined by electron microscopy. DNA content was
also determined for the livers of rats dosed for 7 days. Treatment with 50 mg PFOA/kg for 7
days caused a 17% decrease (p<0.05) in mean body weight. Pair-fed control rats had a 24%
decrease in body weight. Body weight was no different in the rats treated for 1 and 3 days than in
the control rats. Liver weight of rats treated for 1 day was no different than control liver weight.
The relative liver weight of rats treated for 3 days was significantly increased (p<0.05) compared
to control relative liver weight. Absolute and relative liver weights were significantly increased
(p<0.05) after the 7-day treatment with PFOA. A 57% decrease (p<0.05) was observed in
relative hepatic DNA/g liver, but no difference was observed between total amount of hepatic
DNA/liver and total amount of DNA/liver in control rats.
The hepatocytes of rats treated with PFOA for 3 days were enlarged with partially occluded
sinusoids, and had numerous basophilic granules, eosinophilic granular material in the
cytoplasm, and fewer perinuclear glycogen vacuoles compared to control hepatocytes. Enlarged
hepatocytes with hyperplastic smooth endoplasmic reticulum (ER), increased mitochondria,
increased peroxisomes, decreased rough ER, and increased autophagosomes with electron-dense
material also were observed in the hepatocytes.
Loveless et al. (2008) administered 0, 0.3, 1, 10, and 30 mg linear PFOA/kg by oral gavage
to groups of male CD rats (n = 10 per group) for 29 days. Body weight was recorded on days 0,
3, and 6-28. At necropsy, blood was collected for hematology, clinical chemistry, and
corticosterone (CORT) measurements. Tissues were collected for weight and microscopic
examination. Body weight, weight gain, hematocrit, and hemoglobin were reduced at >10 mg
PFOA/kg/day. Increased reticulocytes and hematopoieses were observed in the rats dosed with
30 mg PFOA/kg/day. Total and non-HDL cholesterol were significantly reduced at 0.3 and
1 mg/kg/day compared to control. HDL cholesterol was significantly decreased at 0.3, 1, and
10 mg/kg/day. Triglyceride levels were significantly decreased at all doses except 1 mg/kg.
Absolute liver weight (>1 mg/kg/day) and relative liver weight (>10 mg/kg/day) were
significantly increased. Hepatocellular hypertrophy was graded as minimum to mild
(0.3-1 mg/kg/day) and moderate (>10 mg/kg/day), and focal necrosis was present at doses
> 10 mg/kg/day. Although not statistically significant, serum CORT was increased at
> 10 mg/kg/day. The decrease in cholesterol and triglycerides at the lowest dose are not
necessarily adverse. The 1-mg/kg/day dose is classified as the NOAEL and the 10-mg/kg/day
Perfluorooctanoic acid (PFOA) - May 2016 3-63
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dose as the LOAEL based on the observations of increased liver weight, hepatocellular
hypertrophy, and hepatic necrosis at that dose. Data on several immunological endpoints were
reported as part of the Loveless et al. (2008) publication. The immunological data from that
study are included in section 3.3.2 of this report.
Cui et al. (2009) exposed male Sprague-Dawley rats (10 per group) to PFOA (96% active
ingredient) at 0, 5, and 20 mg/kg/day for 28 days by gavage once daily. The activity of the rats
was observed over the course of the study. All rats were sacrificed after the final exposure. The
rats dosed with 5 mg/kg/day exhibited hypoactivity, decreased food consumption, cachexia, and
lethargy during the third week of the study. Rats dosed with 20 mg/kg/day also exhibited
sensitivity to external stimuli. The visceral index (i.e., hepatic, renal, gonad weight/animal's
body weight) used to evaluate hyperplasia, swelling, or atrophy was significantly increased in the
treated animals compared to control animals. In the liver, treatment with 5 or 20 mg PFOA/kg
caused hepatic hypertrophy, fatty degeneration, and acidophilic lesions as well as angiectasis
(gross dilation) and congestion in the hepatic sinusoid or central vein. In the lung, treatment with
5 or 20 mg PFOA/kg caused pulmonary congestion and focal or diffuse thickened epithelial
walls. No histopathologic lesions were observed in the kidneys of the low-dose animals, but
turbidness and swelling in the epithelium of the proximal convoluted tubule were observed at
20 mg PFOA/kg. Under the conditions of this study, the LOAEL was 5 mg/kg/day based on
increased visceral indices, and liver and pulmonary lesions; no NOAEL was established.
Male Sprague-Dawley rats (n = 10 per group) were fed diets containing 0 and 300 ppm
PFOA for 1, 7, and 28 days in two studies (Elcombe et al. 2010). The mean daily intake for study
1 and study 2 were 19 and 23 mg/kg/d, respectively. A group of rats was fed diets containing
50 ppm Wyeth 14,643, a PPARa agonist, as a positive control. The animals were observed daily
and body weights and food consumption were recorded. At necropsy, day 2, day 8, or day 29, the
organs were weighed, examined for gross pathology and preserved for histopathology. In study
1, liver DNA content and concentration were determined, and plasma was collected for analysis
of liver enzymes, cholesterol, triglycerides, and glucose. Hepatic cell proliferation and apoptosis
also were determined.
In both studies, body weight significantly decreased (p<0.05) after 7 and 28 days on the
PFOA diet. Body weight was not affected by Wyeth 14,643. Absolute liver weight was
significantly increased (p<0.05) in rats fed PFOA diets for 7 days in the first study and in rats
treated for 7 and 28 days in the second study (Table 3-14). The liver-to-body-weight ratio was
significantly higher in rats fed PFOA diets for 7 and 28 days in both studies. Absolute liver
weight and liver-to-body-weight ratios were significantly increased in rats fed the Wyeth 14,643
diet in both studies.
After 1 day of eating the PFOA diet, subjects' plasma AST was significantly decreased and
triglycerides were significantly increased. After 7 and 28 days on the PFOA diet, TC,
triglycerides, and glucose levels were significantly decreased. The AST response did not show a
duration-related response because there was a significant decrease at 1 and 28 days, but not at
7 days on the PFOA diet. Liver DNA concentration was significantly decreased (p<0.05) in all
PFOA-exposed rats except those treated for 1 day in the second study, but liver DNA content
was not altered by PFOA, suggesting that the increase in volume was responsible for the change
in concentration.
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Table 3-14. Hepatic Effects of Rats Exposed to PFOA
Liver
weight (g)
Liver-to-
bw (g/kg)
Labeling
index (%)
Day
1
7
28
1
7
28
1
7
28
Study 1
Control
13.6 ±1.3
15. 3 ±1.3
18.3 ±2.5
4.25 ±.34
4.10 ±0.26
3.96 ±0.36
0.22 ±0.17
1.42 ±0.65
ND
300 ppm
PFOA
(19
mg/kg/day)
14.1 ±2.4
19.2 ±3.1*
20.8 ±3.2
4.39 ±0.44
5.83 ±0.55*
5.83 ±0.56*
0.74 ±0.55*
5.94 ±2. 12*
2.08 ±1.03
50 ppm
Wyeth
14,643
15.7 ±1.2
23.1±3.1*
30.6 ±3.2*
4.64±0.17*
6.26 ±0.48*
7.09 ±0.42*
2.10 ±1.10*
12.56 ±6.42*
10.15 ±2.69
Study 2
Control
15.2 ±1.9
16.6 ±1.7
17.2 ±2.0
4.39 ±0.36
4.28 ± 0.24
3.70 ±0.21
1.02 ±0.37
2.57 ±1.31
0.66 ± 0.45
300 ppm
PFOA
(32
mg/kg/day)
14.4 ±0.9
22.8 ±2.6*
24.6 ±2.2*
4.27 ±0.14
6.56 ±0.38*
6.13 ±0.53*
2.18 ±0.73*
13. 18 ±3. 18*
1.74 ±0.96*
50 ppm
Wyeth 14,643
15.8 ±1.4
23.4 ±2.5*
29.2 ±4.0*
4.49 ±0.23
6.34 ±0.33*
6.65 ±0.59*
4.54 ±1.03*
23. 85 ±7.02*
5.34 ±2.79*
Source: Elcombe et al. 2010
Notes: 'Significantly different from control (p < 0.05); ND = No Data.
After 1 day on the Wyeth 14,463 diet, subjects'AST and TC were significantly decreased.
After 7 and 28 days on the Wyeth 14,643 diet, subjects' ALT, TC, triglycerides, and glucose
levels were significantly decreased. AST was not significantly decreased in rats fed Wyeth
14,643 diets for 28 days, but it was after 7 days. In the Wyeth 14,643 rats, liver DNA
concentration was significantly decreased after 1 and 7 days in the first study and 7 and 28 days
in the second study. Total liver DNA content in the Wyeth 14,643-treated rats was significantly
increased after 7 and 28 days in both studies.
Labeling indices for hepatic cell proliferation, as measured by bromodeoxyuridine (5-bromo-
2-deoxyuridine, BrdU) incorporation, was significantly increased after day 1 and 7 in study 1 in
both PFOA (p<0.05) and Wyeth 14,643 (p<0.01) diet-fed rats. Samples from control livers at
day 29 were not available for comparison. In study 2, labeling was significantly increased
(p<0.05) at all time points in both groups of rats compared to labeling in control rats (Table
3-14). Apoptosis of hepatic cells was not altered by treatment with PFOA at any time point. In
rats fed diets containing Wyeth 14,643 for 28 days, hepatic apoptosis was significantly decreased
(p<0.01) compared to apoptosis observed in control livers.
Histological examination of the livers of PFOA and Wyeth 14,643 diet-fed rats showed
decreased glycogen after 1, 7, and 28 days. An increase in hepatocellular hypertrophy was
observed after 7 and 28 days on the diets, fatty vacuolation was observed after 7 days on the
diets, and increased hepatocellular hyperplasia was observed after 28 days on the diets. The
hepatic observations were similar in both studies, and findings in Wyeth 14,643 diet-fed rats
were generally more pronounced or severe than those in PFOA diet-fed rats. Although there
were many similarities in response to the PFOA and Wyeth 14,463 diets, the body weight and
apoptosis responses differed.
Mouse. Kennedy (1987) fed male and female Crl:CD-l mice diets containing 0, 30, 300, and
3,000 ppm PFOA for 14 days. At necropsy body weight and liver weight were recorded and
analyzed. No histological evaluations were conducted. All mice died at 3,000 ppm. At 300 ppm,
body weight was decreased and one female died. Both male and female mice had significantly
increased absolute and relative liver weights at all doses (p<0.05) compared to the control. The
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LOAEL was 30 ppm based on increased liver weight, and no NOAEL was established. Kennedy
(1987) used lower doses in a follow-up study lasting 21 days. Male and female mice were fed
diets containing 0, 0.01, 0.03, 0.1, 0.3, 1, 3, 10, and 30 ppm PFOA. Absolute and relative liver
weights for male and female mice were significantly increased (p<0.05) at > 3 ppm PFOA. The
LOAEL was 3 ppm based on increased liver weight, and the NOAEL was 1 ppm.
Loveless et al. (2008) administered 0, 0.3, 1, 10, and 30 mg linear PFOA/kg by oral gavage
to groups of male CD-I mice (n = 20 per group) for 29 days. Body weight was recorded on days
0, 3, and 6-28. At necropsy, blood was collected for hematology, clinical chemistry, and CORT
measurements. Tissues were collected for weight and microscopic examination. Body weight
was significantly reduced at 10 and 30 mg/kg/day. An increase in neutrophils and monocytes
was observed at >10 mg/kg/day along with a decrease in eosinophils. Serum CORT levels were
significantly increased in mice dosed with 10 mg/kg and elevated in those dosed with
30 mg/kg/day. Total serum cholesterol and triglycerides were significantly decreased at
>10 mg/kg/day. HDL was significantly reduced at >1 mg/kg/day. In mice treated with 30 mg/kg
and water, triglycerides and HDL levels were significantly decreased compared to control levels.
Absolute and relative liver weights were significantly increased at >1 mg PFOA/kg/day.
Increased incidences of microscopic lesions in the liver included mild hepatocellular hypertrophy
at 0.3 mg/kg/day, moderate-to-severe hypertrophy and individual cell necrosis at>l mg/kg/day,
and increased hepatocellular mitotic figures, fatty changes, and bile duct hyperplasia at
>10 mg/kg/day. The LOAEL for this study was 1 mg/kg/day based on increased liver weight,
hepatocellular hypertrophy, and cell necrosis; the NOAEL was 0.3 mg/kg/day. Data on several
immunological endpoints were reported as part of the Loveless et al. (2008) publication. The
immunological data are included in section 3.3.2 of this report.
Some of the epidemiology studies report that evaluated serum lipids demonstrate a positive
correlation between total serum cholesterol and triglycerides and serum PFOA. Tan et al. (2013)
used male C57BL/6N mice to determine if dietary fat content could be an important variable
influencing the impact of PFOA on serum lipids. Groups of seven or eight 4-month-old male
mice were given either a liquid regular fat diet (RFD) or a high-fat diet (HFD), with or without
PFOA, for 3 weeks. The RFD provided 12% and the HFD provided 35% of their calories from
fat. Calories from protein (18%) were equivalent in both diets. The RFD provided 60% and the
HFD provided 40% of their calories from carbohydrate. The fats were primarily
monounsaturated (olive oil) or polyunsaturated (safflower and corn oil). PFOA was added to
both diets for 3 weeks at a level that maintained a dose of 5 mg/kg/day to the mice. The PFOA-
treated groups were fed ad libitum, and the control groups were given the amount consumed by
the PFOA-treated groups the previous day. Body weight; liver weight; plasma ALT, AST, and
ALP; total and direct bilirubin; free fatty acids and liver triglycerides; as well as subcutaneous
and epididymal white adipose tissue were monitored. Statistical differences between groups
(p<0.05) were determined using one-way ANOVA. Liver and epididymal white fat tissue
samples were examined histologically.
The fat content of the diets alone resulted in significant differences in body weight and
subcutaneous white adipose tissue, but not in liver weight. The addition of PFOA to the RFD
resulted in significant increases in body weight, liver weight, ALT, ALP, and plasma free fatty
acids, but not in AST or bilirubin. The addition of PFOA to both the RFD and HFD resulted in
decreases in the mass of both epididymal and subcutaneous white fat deposits.
The HFD alone did not result in definitive alterations in liver histopathology. When PFOA
was added to the RFD, indications of hepatocyte hypertrophy, necrosis, and inflammatory cell
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infiltration were observed. The liver damage in the animals being fed the HFD with PFOA was
increased more than in the RFD-PFOA animals, as indicated by higher levels of necrosis and
inflammation accompanied, in this case, by lipid droplet accumulation and significantly
increased liver triglycerides, but not liver cholesterol or free fatty acids. In the epididymal
adipose tissues, adipocyte size was increased in the HFD control compared to the RFD control
but decreased with the addition of PFOA compared to both the RFD and HFD controls.
Inflammatory cell infiltration was observed in the epididymal adipose tissues when PFOA was
added to the HFD but not the RFD. No data for the subcutaneous white fat tissues was provided.
The authors evaluated the hepatic expression of 84 genes involved in the regulation of fatty
acid metabolism using RT2 Profiler PCR Arrays. HFD and/or PFOA altered the expression of
33 genes (> 1.5 fold). PFOA alone upregulated 13 genes (>1.5) and downregulated 4 (>1.5)
genes with fatty acid and triglyceride catabolism. Eight fatty acid transport-related genes were
upregulated by PFOA and one was downregulated. The study demonstrates the importance of the
fat content of the diet as a modulator of the effects of PFOA on the liver in animals. Damage to
the liver tissues was intensified in the presence of the HFD.
Son et al. (2008) administered 0, 2, 10, 50, and 250 mg/L PFOA (0, 0.49, 2.64, 17.63, and
47.21 mg/kg PFOA, respectively) in the drinking water to 4-week-old male imprinting control
region (ICR) mice for 21 days. Food and water consumption, and body weight were recorded
daily. At sacrifice, blood was collected and the liver and kidneys were removed and weighed.
Plasma from the blood was used to determine levels of ALT, AST, BUN, and creatinine.
Sections of the liver and kidney were processed and stained with hematoxylin and eosin or
stained for caspase 3 (a biomarker for apoptosis). Expression of mRNA for tumor necrosis
factor-a, interleukin-lp, and transforming growth factor-p were determined using reverse
transcription polymerase chain reaction (RT-PCR).
The mice exposed to 250 mg/L PFOA (47.21 mg/kg/day) had significantly reduced food and
water consumption (p<0.05), and body weight gain (p<0.05) compared to the control mice. Body
weight gain also was significantly reduced (p<0.05) in mice receiving 50 mg/L PFOA in the
drinking water. In all PFOA-exposed mice, relative liver weight was significantly increased in a
dose-dependent manner (p<0.05) compared to liver weight of the control mice. Relative kidney
weight was not affected by PFOA exposure. At > 10 mg/L PFOA (2.64 mg/kg/day), plasma ALT
activity was significantly increased, and at > 50 mg/L PFOA (17.63 mg/kg/day), plasma AST
activity was significantly elevated compared to the activity level in the control mice. Exposure to
PFOA did not affect BUN or creatinine.
The livers of mice exposed to >50 mg/L PFOA were characterized by enlarged hepatocytes
with acidophilic cytoplasm and the presence of eosinophils. No apoptotic bodies were observed
in the liver with staining for caspase 3. Exposure to PFOA did not affect kidney morphology and
did not cause toxic damage or necrosis in the kidney. In the liver, tumor necrosis factor-a
expression was significantly reduced at > 50 mg/L PFOA, interl eukin-lp expression was
significantly reduced at 250 mg/L PFOA, and transforming growth factor-p expression was
significantly elevated at > 50 mg/L PFOA. Under the conditions of this study, the LOAEL was
2 mg/L (0.49 mg/kg/day) based on increased liver weight, and no NOAEL was established. The
LOAEL for increased plasma ALT was 2.64 mg/kg/day.
Wolf et al. (2008a) gavage-dosed wild-type 129Sl/SvlmJ mice (n = 7-8 per group) and
PPARa-null mice (129S4/SvJae-PPARatmlGonz/J, n = 6-8 per group) with 0, 1, 3, or 10 mg
PFOA/kg or 50 mg Wyeth 14,643, a PPARa agonist, and wild-type CD-I (n = 7-8 per group)
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with 0, 1, and 10 mg PFOA/kg for 7 days to characterize hepatic effects resulting from exposure.
The mice were sacrificed 24 hours following the last dosing. Blood was collected for serum, and
the livers were removed and weighed. Liver sections were stained with hematoxylin and eosin
for examination by light microscopy and with uranyl acetate for transmission electron
microscopy. Liver sections were also processed for immunohistochemistry of PCNA. Hepatocyte
hypertrophy and vacuolation, observed in both strains of wild-type mice, were assigned a score
from 0 to 4 based on severity, with 0 being no lesions observed and 4 being panlobular
hypertrophy with cytoplasmic vacuolation. Hepatic lesions in PPARa-null were assigned a score
(0-4) based on cytoplasmic vacuolation as no hypertrophy was observed. The percentage
labeling index was obtained by counting the number of positive PCNA cells in 900-1,000
hepatocyte nuclei per animal. Slides were read blind to treatment but with knowledge of genetic
status.
Compared to control values, the absolute and relative liver weights, lesion score, and labeling
index were significantly increased (p<0.05) in a dose-dependent manner in both strains of wild-
type mice exposed to PFOA and also were significantly increased (p<0.05) in the wild-type
129Sl/SvlmJ mice exposed to Wyeth 14,643 (see Table 3-15). The absolute and relative liver
weights and lesion score were significantly increased (p<0.05) in a dose-dependent manner in all
PFOA-exposed PPARa-null mice. The labeling index was significantly increased (p<0.05) in
PPARa-null mice exposed to 10 mg PFOA/kg. Absolute and relative liver weights, lesion score,
and labeling index of PPARa-null mice exposed to Wyeth 14,643 were no different from control
values.
Table 3-15. Hepatic Effects in PFOA-Treated Mice
Group
Liver Weight (g)
Relative Liver
Weight (%)
Lesion Score
Labeling Index
Wild-type CD-I Mice
Control
1 mg/kg/day PFOA
10 mg/kg/day PFOA
1.53 ±0.14
2.26 ±0.24*
3.48 ±0.54*
4.5 ±0.4
6.5 ±0.5*
10.5 ±0.8*
0.3 ±0.5
2.1 ±0.9
3.0 ±0*
0.6 ±0.4
0.7 ±0.5
7.7 ±3.0*
Wild-type 12981/SvlmJ Mice
Control
1 mg/kg/day PFOA
3 mg/kg/day PFOA
10 mg/kg/day PFOA
50 mg/kg/day Wyeth
14,643
0.87 ±0.08
1.22 ±0.22*
1.70 ±0.12*
2.20 ±0.23*
1.5±0.13*
3.3 ±0.4
1.6 ±0.2*
6.4 ±0.4*
8.3 ±0.2*
5.6 ±0.1*
0.3 ±0.5
2.0 ±0.0*
2.0 ±0.0*
4.0 ±0.0*
3.3 ±0.5*
0.3 ±0.2
0.7 ±0.6
1.0 ±0.4
2.4 ±0.9*
2.1 ±1.2*
PPARa-null Mice
Control
1 mg/kg/day PFOA
3 mg/kg/day PFOA
10 mg/kg/day PFOA
50 mg/kg/day
Wyeth 14,643
0.92 ±0.08
1.2 ±0.14*
1.46 ±0.21*
2.8 ±0.18*
1.07 ±0.24
3.4 ±0.4
4.5 ±0.2*
5. 8 ±0.3*
9.4 ±0.6*
3.9 ±0.5
1.1 ±0.4
1.9 ±0.6*
3.0 ±0.0*
4.0 ±0.0*
1.4 ±0.5
0.2 ±0.2
0.6 ±0.4
0.6 ±0.3
7.7 ±3.0*
0.6 ±0.5
Source: Wolf etal. 2008a
Note: * Statistically different from control, p < 0.05.
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Ultrastructure evaluations were done on liver sections from wild-type 12981/SvlmJ mice and
PPARa-null mice, but not from CD-I 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 PPARa-null mice, the responses of the
control and Wyeth 14,643-dosed animals were similar, but the response of the PFOA-dosed
animals differed. Table 3-16 summarizes the cellular characteristics of the hepatocytes for the
control, POFA-treated, and Wyeth 14,643-treated wild-type and PPARa-null mice on the basis
of their glycogen content, Golgi bodies and associated rough ER, mitochondria, peroxisomes,
and lipid-like cytoplasmic vacuoles.
Table 3-16. Mouse Hepatocyte infrastructure After PFOA or Wythe 14,643 Treatment
Mou se/Treatment
Wild-type/Control
Wild-type/PFOA
(10 mg/kg)
Wild-type/Wyeth
PPARa-null/Control
PPARa-null/PFOA
(10 mg/kg)
PPARa-null/Wyeth
Characteristics
Glycogen
Prominent
Negative
Negative
Prominent
Limited
Prominent
Golgi/ Rough
ER
Prominent
Nominal/ scarce
ER
Nominal/ scarce
ER
Prominent
Limited
Prominent
Mitochondria
Numerous
Numerous
Numerous
Numerous
Not reported
Numerous
Peroxisomes
Few
Numerous
Numerous
Absent
Not reported
Absent
Lipid-like
Vacuoles
Rare
Scattered
Scattered
Scattered
Numerous3
Scattered
Source: Wolf etal. 2008a
Note: a Described as electron-dense, nonmembrane-bound spaces morphologically consistent with lipids ranging from the size of
mitochondria to the size of nuclei. The vacuoles were believed to be an accumulation of PFOA.
It is apparent from the data in Table 3-16 that PFOA and Wyeth 14,643 behaved similarly in
the wild-type strains but differently in the PPARa-null mice. The hepatocytes of PFOA-dosed
PPARa-null mice exhibited lower glycogen content, Golgi bodies, and associated rough ER than
both the control and Wyeth 14,643 PPARa-null mice. In addition, the PFOA-dosed PPARa-null
mice had numerous large nonmembrane-bound lipid-like vacuoles throughout the cytoplasm. At
the high dose (10 mg/kg/day), there was an increase in the labeling index that was not observed
with Wyeth 14,643. The authors concluded that the large lipid-like vacuoles in the hepatocytes of
PFOA-dosed PPARa-null mice were likely accumulations of PFOA. Under the conditions of this
study, the LOAEL was 1 mg/kg/day based on increased absolute and relative liver weight and
hepatic morphology changes; no NOAEL was established.
Nakamura et al. (2009) investigated the functional difference in PFOA response between
mice and humans using a humanized PPARa transgenic mouse strain (hPPARa). Humanized
PPARa mice express a high level of human PPARa protein in the liver. Male 8-week-old wild-
type (mPPARa) mice, PPARa-null mice, and hPPARa mice were gavage-dosed with 0, 0.1, and
0.3 mg/kg/day PFOA (n = 4-6 per group) for 2 weeks and sacrificed 18-20 hours following the
last dose. Blood was collected and analyzed for triglyceride and cholesterol concentrations, and
ALT measurements. Livers were collected and analyzed for triglyceride and cholesterol
concentrations, plus histopathological changes. The differences in the observations for the three
strains of mice are summarized in Table 3-17.
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Table 3-17. Relative Response of hPPARa, mPPARa, and PPARa-null Mice to PFOA
Parameter
Liver weight
Liver/body weight ratio
Hepatocyte hypertrophy
ALT
Plasma cholesterol
Liver cholesterol
Plasma triglyceride
Liver triglyceride
hPPARa
ND
ND
Mild (0.3 mg/kg/day)
ND
t compared to mPPARa
(all doses)
4 compared to PPARa-null
(0.1, 0.3 mg/kg/day),
mPPARa (0.3 mg/kg/day)
ND
4 compared to PPARa-null
(0.3 mg/kg/day)
mPPARa
t compared to control (0.3
mg/kg/day)
t compared to control (0.3
mg/kg/day)
Mild (0.3 mg/kg/day)
ND
ND
t compared to control (0.3
mg/kg/day)
ND
J, compared to PPARa-null
(0. 1,0.3 mg/kg/day; t
compared to control (0.3
mg/kg/day)
PPARa-null
4 compared to control (0.1
mg/kg/day)
ND
ND
ND
ND
ND
ND
t compared to mPPARa
(all doses)
Source: Nakamura et al. 2009
Notes:
hPPARa: transgenic mice (that express a high level of human PPARa protein in the liver); mPPARa: wild-type mice.
t = significant increase (p < 0.05).
4 = significant decrease (p < 0.05)
ND = no differences.
Body weight of the hPPARa mice was slightly lower than the mPPARa and PPARa-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 hPPARa
mice differed from the wild-type mice in that their plasma cholesterol was significantly increased
and their liver cholesterol and triglycerides significantly decreased at the highest dose (Table
3-17). In addition, the increases in absolute and relative liver weights were less than those
observed in the wild-type mice. The PPARa-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 PPARa-null mice than in the wild-type and hPPARa
mice. There were no other significant differences between PPARa-null mice and wild-type mice.
Under the conditions of the study, the LOAEL for mPPARa mice was 0.3 mg/kg/day of
PFOA based on increased liver weight and increased liver triglyceride and cholesterol
concentrations. The NOAEL for mPPARa mice was 0.1 mg/kg/day of PFOA. The NOAEL for
PPARa-null mice was 0.3 mg/kg/day because the changes in absolute liver weight were not
dose-related and the increase in relative liver weight was not significantly different from the
control. The NOAEL for hPPARa mice was 0.3 mg/kg/day of PFOA, the highest dose tested.
However, a nonsignificant but dose-related increase was observed in plasma cholesterol.
Minata et al. (2010) examined hepatobiliary injury in mice treated with PFOA. Male wild-
type 12984/SvlmJ mice (n = 39) and PPARa-null (129S4/SvJae-PparatmlGonz/J mice (n = 40)
were orally dosed with 0, 12.5, 25, and 50 umol/kg/day of PFOA (equivalent to ~0, 5.4, 10.8,
and 21.6 mg/kg/day of PFOA) for 4 weeks. At the end of 4 weeks, animals were sacrificed and
blood, liver, and bile were collected for clinical chemistry analysis and determination of PFOA
concentration. Sections of the liver were processed for histological examination, oxidative DNA
damage, and multidrug resistance protein 2 (Mdr2) and tumor necrosis factor a (TNF-a) mRNA
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expression. Bile acid and phospholipid contents in bile were determined as well as the protein
expression of canalicular bile salt export pump (BSEP) and canalicular MRP2.
Absolute and relative liver weights in all PFOA treated wild-type and PPARa-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 umol/kg/day (equivalent to 10.8 and
21.6 mg/kg/day) and in PPARa-null mice at 50 umol/kg/day compared to the concentrations of
their respective controls. Plasma ALT was no different from control in the treated mice. In wild-
type mice, total bilirubin was significantly decreased at 12.5 umol/kg/day and significantly
increased at 50 umol/kg/day. In PPARa-null mice, total bilirubin was significantly increased at
50 umol/kg/day. Total bile acid was significantly increased at 50 umol/kg/day in PPARa-null
mice. TC was significantly decreased in wild-type mice at 25 and 50 umol/kg/day, and total
triglyceride was significantly increased at 12.5 and 25 umol/kg/day. TC was significantly
decreased at 12.5 and 25 umol/kg/day and significantly increased at 50 umol/kg/day in PPARa-
null mice. In PPARa-null mice, total triglycerides were significantly increased at all doses.
Hepatocellular hypertrophy was observed in wild-type mice treated with 12.5, 25, and
50 umol/kg/day (equivalent to 5.4, 10.8 and 21.6 mg/kg/day). A dose-dependent increase in
eosinophilic cytoplasmic changes consistent with peroxisome proliferation was observed in liver
parenchyma, but no fat droplets or focal necrosis were observed in wild-type mice. An increase
in bile duct epithelium thickness suggested slight cholangiopathy in wild-type mice at 25 and
50 umol/kg/day. Increased apoptosis in hepatic cells, hepatic arterial walls, and bile duct
epithelium was observed at 25 and 50 umol/kg/day in wild-type mice. Ultrastructure examination
of livers from PFOA-treated wild-type mice showed decreased glycogen granules, degranulated
or disrupted rough ER, nuclear vacuoles, extensive peroxisome proliferation, and slight
mitochondria proliferation.
In PPARa-null mice treated with 12.5, 25, and 50 umol/kg/day of PFOA (equivalent to 5.4,
10.8 and 21.6 mg/kg/day), hepatocellular hypertrophy, cytoplasmic vacuolation, and increased
microvesicular steatosis were observed. These observations are consistent with Wolf et al.
(2008a). At 50 umol/kg/day, focal necrosis was observed. Areas of bile fibrosis and bile plaque
and few inflammatory cells were observed in the bile ducts of PPARa-null mice at 25 and
50 umol/kg/day. Increased apoptosis was observed in bile duct epithelium at 25 and
50 umol/kg/day in PPARa-null mice. Ultrastructure examination of livers from PFOA-treated
PPARa-null mice showed decreased glycogen granules, degranulated or disrupted rough ER,
increased cytoplasmic lipid accumulation, mitochondria proliferation, and mitochondrial changes
(e.g., swelling and decreased matrix density). Peroxisome proliferation was not observed.
Ultrastructure of bile duct showed degradation of cytoplasmic structure, vacuolization,
disintegration of nuclei and organelles, periductal infiltration of fibroblasts and macrophages,
and fibrosis.
The marker for oxidative damage, 8-hydroxydeoxyguanosine (8-OH-dG), and TNF-a were
not elevated or upregulated in wild-type mice. In PPARa-null mice, 8-OH-dG was elevated in
the liver at 21.6 mg/kg/day and TNF-a mRNA was significantly increased at 10.8 and 21.6
mg/kg/day. The transporter Mdr2 moves biliary phospholipids from hepatocytes to bile and was
significantly upregulated in wild-type mice at all doses, but only at 5.4 mg/kg/day in PPARa-null
mice. The BSEP transports bile acid from hepatocytes to bile and was significantly decreased in
wild-type mice at 21.6 mg/kg/day, significantly increased in PPARa-null mice at 5.4 mg/kg/day,
and significantly decreased at 21.6 mg/kg/day. The transporter MRP2 also transports bile acid
and was significantly decreased at 21.6 mg/kg/day in both groups of mice.
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Under the conditions of the study, the LOAEL for male wild-type and PPARa-null mice was
5.4 mg/kg/day of PFOA based on increased liver weight. A NOAEL was not established. At the
LOAEL, the difference between the PPARa-null mice and the wild-type mice was the presence
of cytoplasmic vacuoles and microvesicular steatosis in addition to hypertrophy in the PPARa-
null mice.
The effects of gavage exposure on groups of six male Klunming mice (8 weeks old) to doses
of 0, 2.5, 5, and 10 mg PFOA/kg/day for 14 days on the testes and epididymis was examined by
Liu et al. (2015). The lowest dose tested was a LOAEL for dose-related effects on decreased
sperm count, testicular superoxide dismutase (SOD), catalase, nuclear respiratory factor 2
(NRF2), and BAX expression (0<0.05) plus increases in MDA, hydrogen peroxide, BAX and
BCL expression (p<0.05). There was no effect on relative testes weight at any dose. Some effects
were observed on testicular morphology at the lowest dose, including atrophy of the
seminiferous tubules, depletion of spermatogonial cells, detachment of germ cells from the
seminiferous epithelium, and decreased sperm production. The severity of the testicular
morphological changes increased with dose. Six animals per dose group were used for the
evaluation of testicular weight and 4 animals per dose group were used for the other assays. The
increase in MDA and hydrogen peroxide accompanied by the decrease in SOD and carnitine
acyltransferase (CAT) activity and NRF2 expression indicate that oxidative stress played a major
role in the observed toxicity. NRF2 plays an important role as a messenger that upregulates
genes involved in response to oxidative stress.
Lu et al. (2015) reported on the testicular effects of PFOA on the blood testes barrier after a
28-day exposure of BALBL/c male mice (14 days old) to gavage doses of 0. 1.25, 5, and
20 mg/kg/day (3-5 animals per dose group). The blood testes barrier divides the seminiferous
epithelium into apical and basolateral compartments and plays an important role in germinal cell
development and male fertility. The barrier prevents the passage of large molecules from one
compartment to the other. At termination of the exposure, the animals were sacrificed and the
testes recovered for analysis. A second component of this study examined the impact of the
PFOA treatment on male fertility and is reported in section 3.2.6.
The blood testes barrier integrity was weakened at the lowest dose tested and in a dose-
dependent manner as indicated by the passage of a red fluorescent dye injected into the
interstitium and concentrations of IgG measure in gel electrophoresis columns visualized by
chemiluminescence (three per dose group). Membrane integrity is dependent on coexisting tight
junctions, basal ectoplasmic specializations, and gap junctions. Accompanying in vitro assays of
cultured sertoli cells demonstrated downregulation of key proteins associated with the tight
junction and gap junction intercellular communication (GJIC) and regulation of the ectoplasmic
specialization protein N-cadhedran. Tumor necrosis factor actin protein in the testes increased in
a dose-related fashion at 5 and 20 mg/kg/day on observation of three per dose group. The authors
identified the 5-mg/kg/day dose as a LOAEL for PFOA effects on the blood testes barrier and the
1.25-mg/kg/day dose as a NOAEL, apparently based on the results for the key protein
biomarkers for cellular intercommunication rather than the IgE and fluorescence results where
1.25 mg/kg/day was a LOAEL.
Li et al. (2011) investigated the involvement of mouse and human PPARa in PFOA-induced
testicular toxicity. Wild-type, PPARa-null, and humanized PPARa male 129/Sv mice were given
PFOA daily by gavage at doses of 0, 1, and 5 mg/kg/day for 6 weeks. Body weight and testis
weight were not affected by treatment in any group. Absolute and relative-to-body weights of the
epididymis and seminal vesicle plus prostate gland were decreased only in high-dose wild-type
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mice compared to the wild-type controls. No effects on sperm count and motility were seen in
any group. Sperm abnormalities were significantly increased in both treated groups of wild-type
and humanized PPARa mice, but not in the PPARa-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 PPARa mice compared to the control groups.
Testosterone levels were slightly reduced in a dose-related manner in the PPARa-null mice, but
statistical significance was not attained.
Using real-time quantitative PCR, the mRNA levels for several genes associated with
testicular cholesterol synthesis, transport, and testosterone biosynthesis were examined. Levels
of 3-hydroxy-3-methylglutaryl coenzyme A (HMG-CoA) synthase, HMG-CoA reductase, and
aromatase were not changed after treatment in any group. Expression of steriodogenic acute
regulatory protein (which transports cholesterol into mitochondria) was inhibited in wild-type
mice at the high dose and in humanized PPARa mice at both doses; peripheral benzodiazepine
receptor level was decreased only in high-dose humanized PPARa mice; cytochrome P450 side-
chain cleavage enzyme was decreased in both groups of wild-type mice; cytochrome P450 17a-
hydroxylase/C 17-20 lyase was inhibited at the high dose in both wild-type and humanized
PPARa mice; and 3p-hydroxysteroid dehydrogenase was decreased in both treated groups of
humanized PPARa mice. Decreased expression of l?p-hydroxysteroid dehydrogenase was the
only change found in treated PPARa-null mice.
In the mitochondria, carnitine palmitoyltransferase (CPT) was decreased in both groups of
wild-type and high-dose humanized PPARa mice, and SOD levels were reduced in all treated
wild-type and humanized PPARa 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 PPARa mice. No morphological changes were observed in the testes
from PFOA treatment in PPARa-null mice. The 1-mg/kg/day dose was the LOAEL for
significant (p<0.05) sperm abnormalities, decreased testosterone, and several biochemical
alterations in the PPARa and hPPARa mice, but not in the PPARa-null mice. There were dose-
related decreases in testosterone in the PPARa-null mice, but they did not achieve statistical
significance.
Inhalation Exposure
No data on the effects of short-term inhalation exposures to PFOA were identified in the
literature.
Dermal Exposure
Fairley et al. (2007) investigated the role of dermal exposure to PFOA in an experiment to
evaluate toxicity in BALB/c mice. The mice were exposed to 0, 0.01%, 0.1%, 0.25%, 0.5%,
1.0%, and 1.5% PFOA (equivalent to 0, 0.25, 2.5, 6.25, 12.5, 25, and 50 mg/kg PFOA). It was
applied to the dorsal surface of both ears daily for 4 days. The mice were sacrificed 6 days later.
Dermal PFOA exposure did not cause reductions in body weight or signs of inflammation at the
application site. A significant increase in liver weight was observed in mice dosed with
> 6.25 mg/kg PFOA (p<0.01) compared to control liver weight. Under the conditions of the
study, the LOAEL was 6.25 mg/kg PFOA based on increased liver weight, and the NOAEL was
2.25 mg/kg PFOA.
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3.2.3 Subchronic Studies
Oral Exposure
Monkey. Goldenthal (1978) administered rhesus monkeys (two per gender per group) doses of
0, 3, 10, 30, and 100 mg/kg/day PFOA by gavage for 90 days. Animals were observed twice
daily and body weights were recorded weekly. Blood and urine samples were collected once
during a control period, and at 1 and 3 months for hematology, clinical chemistry, and urinalysis.
Organs and tissues from animals that were sacrificed at the end of the study and from animals
that died during the treatment period were weighed, examined for gross pathology, and
processed for histopathology.
All monkeys in the 100-mg/kg/day group died between weeks 2-5 of the study. Signs and
symptoms that first appeared during week 1 included anorexia, frothy emesis, swollen face and
eyes, decreased activity, prostration, and body trembling. Three monkeys from the 30-mg/kg/day
group died during the study. Beginning in week 4, all four animals showed slight to moderate,
and sometimes severe, decreased activity. One monkey had emesis and ataxia, swollen face,
eyes, and vulva. Beginning in week 6, two monkeys had black stools and one monkey had 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 versus 3.78 kg for the control).
The results of the urinalysis, and hematological and clinical chemistry analyses were comparable
for the control and the 3- and 10-mg/kg/day groups at 1 and 3 months.
At necropsy, there were significant decreases in the absolute heart and brain weight and
relative liver weight in 10-mg/kg/day females. At 3 mg/kg/day, the relative pituitary weight in
males was significantly increased. The biological significance of these weight changes is
difficult to assess, as they were not accompanied by morphologic changes.
In animals that died, one male and two females from the 30-mg/kg/day group and all animals
from the 100-mg/kg/day group had marked diffuse lipid depletion in the adrenal glands. All
males and females from the 30- and 100-mg/kg/day groups also had slight to moderate
hypocellularity of the bone marrow and moderate atrophy of lymphoid follicles in the spleen.
One female from the 30-mg/kg/day group and all animals in the 100-mg/kg/day group had
moderate atrophy of the lymphoid follicles in the lymph nodes.
The one male in the 30-mg/kg/day group that survived until terminal sacrifice had slight-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/day based on decreased heart and brain weight, and the NOAEL was 3 mg/kg/day.
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Rat. In a dietary study reported by Perkins et al. (2004), male ChR-CD rats (45-55 per group)
were administered concentrations of 1, 10, 30, and 100 ppm PFOA for 13 weeks. These doses
are equivalent to 0.06, 0.64, 1.94, and 6.50 mg/kg/day. There were two control groupsa
nonpair-fed control group and a pair-fed control group for the 100-ppm dose group); both were
fed the basal diet. Following the 13-week exposure period, 10 animals per group were fed basal
diet for an 8-week recovery period. The animals were observed twice daily for clinical signs of
toxicity, and body weights and food consumption were recorded weekly. Food consumption was
recorded daily for the pair-fed animals.
A total of 15 animals per group were sacrificed following 4, 7, and 13 weeks of treatment;
10 animals per group were sacrificed after 13 weeks of treatment and an 8-week recovery period.
Serum samples collected from 10 animals per group at each scheduled sacrifice during treatment
and from five animals per group during recovery were analyzed for estradiol, T, LH, and PFOA.
The level of palmitoyl-CoA oxidase was analyzed from a section of liver that was obtained from
five animals per group at each scheduled sacrifice. Weights of the brain, liver, lungs, testis,
seminal vesicle, prostate, coagulating gland, and urethra were recorded, and these tissues also
were examined histologically. In addition, the brain, liver, lungs, testis, seminal vesicle, and
prostate were preserved in glutaraldehyde for electron microscopic examination.
In the analysis of the data, animals exposed to 1, 10, 30, and 100 ppm PFOA were compared
to the control animals in the nonpair-fed group, while the data from the pair-fed control animals
were compared to animals exposed to 100 ppm PFOA. At 100 ppm, significant reductions in
body weight and body weight gain were seen compared to the pair-fed control group during
week 1 and the nonpair-fed control group during weeks 1-13. Body weight data in the other
dosed groups were comparable to controls. At 10 and 30 ppm, mean body weight gains were
significantly lower than for the nonpair-fed control group at week 2. These differences in body
weight and body weight gains were not observed during the recovery period. Animals fed
100 ppm consumed significantly less food during weeks 1 and 2 than the nonpair-fed control
group. Overall, there was no significant difference in food consumption. There were no
significant differences among the groups for any of the hormones evaluated, although there was
some indication of elevated estradiol for the 100 ppm group at week 4. The elevated estradiol for
the high-dose group at week 4 should be interpreted with caution because most of the
measurements for control and treated groups were below the level of detection at all other
timepoints (Perkins et al. 2004).
Significant dose-related increases in absolute and relative liver weights and hepatocellular
hypertrophy were observed at weeks 4, 7, and 13 in the 10, 30, and 100 ppm groups
(Table 3-18). There was no significant evidence of any dose-related degenerative changes.
Hepatic palmitoyl-CoA oxidase activity was significantly increased at weeks 4, 7, and 13 in the
30 and 100 ppm groups. At 10 ppm, hepatic palmitoyl-CoA oxidase activity was significantly
increased at week 4 only. At 13 weeks, the palmitoyl-CoA oxidase activity was lower than it was
at weeks 4 and 7 for the 10, 30, and 100 ppm dose groups, possibly suggesting attenuation of the
peroxisomal response. Histologically, liver effects were limited to minimal or slight coagulative
necrosis observed in 0/45, 1/45, 0/45, 2/45, and 3/44 in the control, 1, 10, 30, and 100 ppm
groups, respectively.
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Table 3-18. Liver Effects in Male Rats
Parameter
Palmitoyl-CoA
Oxidase (lU/g)
Hepatocellular
Hypertrophy
Hepatocellular
Necrosis,
Coagulative
Absolute Liver
Weight (g)
Mean Body
Weight (g)
Liver/Body
Weight (%)
Week
4
7
13
4
7
13
4
7
13
4
7
13
4
7
13
4
7
13
Dose (mg/kg/day)
Oa
8 ±0.5
7 ±1.5
8 ±0.9
0/15
0/15
0/15
0/15
0/15
0/15
16.34 ±2.14
17.78 ±2.12
19.73 ±2.01
388 ±21
457 ± 29
541 ±41
3.97± 0.37
3.75 ±0.29
3.53 ±0.28
Ob
8 ±0.4
7 ±1.5
5±1.1
0/15
0/15
0/15
1/15
0/15
0/15
15. 83 ±1.13
16.91 ±2.22
16.30 ±1.62
365 ±11
434 ±19
508 ± 22
4.07 ±0.27
3.76 ±0.37
3.24 ±0.23
0.06 (1 ppm)
9 ±1.7
7 ±0.8
8 ±1.9
0/15
0/15
0/15
0/15
0/15
1/15
15.45 ±1.71
17.68 ± NA
18.03 ±2.81
388 ±23
461 ±30
548 ±37
3.73 ±0.23
3.64 ±0.33
3.24±0.30C
0.64 (10 ppm)
14±3.6C
16 ±5.5
10 ±2.1
12/15
12/15
13/15
0/15
0/15
0/15
17.89 ±2.13
19.42 ±2.10
20.44 ± 2.87
383 ±25
458 ± 30
551 ±42
4.49 ±0.32C
4.12 ±0.37
3.69 ±0.32
1.94 (30 ppm)
24±11.4C
32±12.2C
14±3.4C
15/15
15/15
14/15
1/15
0/15
1/15
23.23 ±2.83C
27.76 ±3.51C
22.74 ±4.21
380 ± 27
448 ±31
531 ±46
5.77 ±0.60C
5.14±0.53C
4.21±0.56C
6.5 (100 ppm)
37 ± 14.8c'd
54±35.3c'd
17±4.5cd
14/15
15/15
15/15
2/14
1/15
0/15
25.44 ± 1.89C
27.76 ±3.51C
26.78 ± 5.47C
356 ± 27C
432±39C
494 ± 64C
6.73 ±0.49C
6.06 ± 0.72C
5.50±0.84C
Source: Perkins et al. 2004
Notes: Mean ± SD; NA= not available.
a Nonpair-fed controls.
b Pair-fed controls.
c Statistically significant (p < 0.05) with nonpair-fed control.
d Statistically significant (p < 0.05) with pair-fed control.
Under the conditions of this study, the authors identified the LOAEL as 10 ppm
(0.64 mg/kg/day) based on increases in absolute and relative liver weight and hepatocellular
hypertrophy (Perkins et al. 2004). The NOAEL identified was 1.0 ppm (0.06 mg/kg/day).
However, the liver weight and palmitoyl-CoA responses were associated with the activation of
PPARa and were not accompanied by significant dose-related changes that would classify them
as adverse for humans (e.g., fibrosis, macrovesicular steatosis, inflammation) as enumerated by
Hall et al. (2012). Therefore, for the current assessment, the LOAEL is identified as
1.94 mg/kg/day based on a slight increased incidence of coagulative necrosis in the liver. The
NOAEL is 0.64 mg/kg/day.
Serum samples were collected from 8 to 10 animals prior to each sacrifice. PFOA
concentrations in serum increased with the dose, but all dose levels appeared to have reached
steady state by the first sacrifice at week 4. Following the 8-week recovery period, serum levels
were below detection for many animals and consistent with a half-life of about seven days in
male ChR-CD rats.
Inhalation and Dermal Exposure
No data on the effects of subchronic inhalation or dermal exposures to PFOA were identified
in the literature.
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3.2.4 Neurotoxicity
Johansson et al. (2008) gave male neonatal Naval Medical Research Institute (NMRI) mice
(3-4 litters, -5-6 male pups per litter) a single gavage dose of 0, 0.58, and 8.7 mg PFOA/kg in a
lecithin/peanut oil emulsion on PND 10, the approximate peak time of rapid brain growth in
mice. Spontaneous behavior (e.g., locomotion, rearing, and total activity) and habituation in
response to a placement in unfamiliar environment were tested in 10 mice in each group at ages
2 and 4 months. Each test period was divided into three 20-min periods. The habituation ratio
was determined by dividing the activity for the third 20-min period by the activity for the first
period. A high habituation ratio indicated that movement patterns of the exposed animals when
placed in an unfamiliar test chamber differed from control by displaying comparatively low
activity for the first 20 mins and comparatively higher activity for the last 20 mins.
Exposure to PFOA did not affect body weight or body weight gain in male NMRI mice
following treatment. Compared to controls, the habituation ratio for rearing and locomotion in
the high-dose animals was elevated compared to controls at 2 and 4 months, with a significantly
higher ratio (p<0.01) at 4 months than at 2 months. At 4 months, the changes in activity patterns
for the high dose were significant (p<0.01) compared to controls for locomotion, rearing, and
total activity. The results at the low dose were less pronounced, with a significant impact on
locomotion and slight changes in rearing behavior.
At 4 months of age, mice were tested for nicotine-induced behavior and behavior in the
elevated plus maze. Increased activity is the expected response to nicotine injection (80 ug) as a
result of stimulation of the cholinergic receptors in the brain. The activity responses of the
PFOA-exposed animals to nicotine stimulation were significantly less than the response of the
controls, but the differences were most pronounced in the high-dose animals.
The mice also were tested in an elevated plus maze, which determined whether they would
select an enclosed environment (the expected response) over an open environment. No
significant differences were observed in the PFOA-exposed mice in this test. Under the
conditions of this study, the clear LOAEL was 8.7 mg/kg based on locomotion, rearing, and total
activity; habituation ratio; and response to nicotine at 2 and 4 months after receiving a single
gavage dose on PND 10. There were significant differences in locomotion and total activity at 4
months in the low-dose animals, which supports identifying the 0.58-mg/kg dose as a marginal
LOAEL. However, the data at the low dose are less compelling than those at the high dose.
In a follow-up to their original study, Johansson et al. (2009) gave male neonatal NMRI mice
(3-4 litters, -5-6 male pups per litter) a single gavage dose of 0 and 8.7 mg PFOA/kg on PND
10. Protein levels of calcium/calmodulin-dependent protein kinase II (CaMKII), 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 unchanged in the cerebral cortex. Levels
of GAP-43 protein in the hippocampus were significantly higher (17%, p<0.05) in PFOA-
exposed mice than levels in control mice, but unchanged in the cerebral cortex. Synaptophysin
levels in mice exposed to PFOA were significantly increased in the hippocampus (52%) and
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cerebral cortex (82%). Tau protein levels in PFOA-exposed mice were increased 92% and 142%
(p<0.001) in the hippocampus and cerebral cortex, respectively, above levels in the control mice.
The authors concluded that alterations of these proteins could be a factor in the altered behavior
of adult mice that were exposed to PFOA as neonates because they are required for normal brain
development.
Onishchenko et al. (2011) exposed pregnant C57BL/6/Bkl mice (n = 6 per group) to 0 and
0.3 mg PFOA/kg/day in the diet from GD 1 to the end of pregnancy. The behavior of the weaned
offspring was analyzed in locomotor, circadian activity, elevated plus maze, and forced swim
tests at 5-8 weeks of age. Muscle strength and motor coordination tests were given at 3-4
months of age. The distance traveled over 30 mins was registered in 5-min intervals in the
locomotor test. For the circadian activity test, the activity of the mice in social groups was
monitored for 48 hours after placement in new cages. Anxiety-like behavior was determined
using the elevated plus maze. Depression-like behavior was determined in the forced swim test
by tracking the time spent floating passively for 2 seconds or longer. Muscle strength (three
trials) was measured by how long within 60 seconds it took the mouse to fall off an upside-down
lid onto the cage floor. Motor coordination (four trials) was measured by how long the mice
remained on a rotating drum as a rotarod accelerated from 4 to 40 rpm over 5 mins.
Prenatal exposure to PFOA did not alter offspring locomotor activity, anxiety-related
behavior, depression-like behavior, or muscle strength. In the circadian activity tests, male
offspring exposed to PFOA were significantly more active (p = 0.013) and the female offspring
were significantly less active (p = 0.036) than control offspring during the first hour of the test.
PFOA-exposed male offspring were significantly more active (p<0.05) than control males from
the dark phase of day 1 through the dark phase day 2. Both male and female offspring exposed to
PFOA had significantly less inactive periods (p<0.05) during the light phase compared to their
respective controls. In the accelerating rotarod test, female offspring exposed to PFOA exhibited
decreased fall latency over the four trials compared to control females, but no effect of treatment
was observed in male offspring. The authors concluded that prenatal exposure to 0.3 mg/kg/day
of PFOA resulted in gender-related postnatal alterations in offspring behavior and motor function
at 3-4 months of age.
In vitro. Slotkin et al. (2008) characterized the neurotoxicity of PFOA using PC12 cells. The
cells were derived from a neuroendocrine tumor of the rat adrenal medulla and serve as a model
for neuronal development and differentiation. Exposure to nerve growth factors causes PC 12
cells to differentiate into cells expressing either dopamine or acetylcholine phenotypes. The cells
were incubated with 10, 50, 100, and 250 jimol PFOA. Synthesis of DNA, cell viability, cell
growth, and lipid peroxidation were measured to determine if PFOA targets specific events in
neural cell differentiation. Differentiation shifts towards or away from the dopamine and
acetylcholine phenotypes were measured by assessing the activities of tyrosine hydroxylase (TH,
dopamine) and choline acetyltransferase (ChAT, acetylcholine). The undifferentiated cells were
evaluated after a 24-hour exposure to PFOA, and differentiating cells were evaluated after 4-6
days of exposure to PFOA.
Significant inhibition of DNA synthesis (p<0.0001) occurred in the undifferentiated cells
after exposure to 250 jimol PFOA with no change in DNA content. Lipid peroxidation was
significantly increased (p<0.02) after exposure to 10 jimol PFOA, and cell viability was
significantly decreased (p<0.03) after a 24-hour exposure to 100 and 250 jimol PFOA.
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In differentiating PC12 cells, exposure to 250 jimol PFOA caused decreased DNA content
with no change in total protein/DNA content ratio or the membrane/total protein ratio. The
lowest and highest PFOA concentrations caused a significant increase in lipid peroxidation
(p<0.007), but no effect was observed in cell viability. TH activity was decreased (p<0.05) after
exposure to 10 and 250 |imol PFOA, and the TFt/ChAT ratio was decreased (p<0.05) at 10 jimol
PFOA. The results suggest that PFOA exposure caused the differentiating cells to shift slightly to
favor the acetylcholine phenotype.
3.2.5 Developmental/Reproductive Toxicity
Reproductive Effects
A comprehensive two-generation reproductive toxicity study was conducted in Sprague-
Dawley Rats with publication of the results by Butenhoff et al. (2004a). A subsequent study by
York et al. (2010) provided details of male reproductive organ histopathology. One study in mice
examined the impact of mating exposed males with unexposed females on fertility and neonatal
body weight (Lu et al. 2015).
Rat. A standard oral two-generation reproductive toxicity study of PFOA in Sprague-Dawley
rats was conducted (Butenhoff et al. 2004a). Five groups of male and female SD rats (30 per
gender per group) were administered PFOA by gavage at doses of 0, 1, 3, 10, and 30 mg/kg/day.
The parental generation (FO) rats (n = 30 per gender per group) were dosed for 10 weeks prior to
mating and until sacrificed (after mating for males; after weaning for females). Fl generation rats
(n = 60 per gender per group) were dosed similarly, beginning at weaning. The F2 generation
rats were maintained through LD 22. Reproductive parameters evaluated in the FO and Fl
generations included estrus cyclicity, sperm number and quality, mating, fertility, natural
delivery, and litter viability and growth. Age at sexual maturation in Fl pups, anogenital distance
in F2 pups, and presence of nipples (males) in F2 pups also were determined. Food consumption,
body-weight gain, selected organ weights, gross pathology, and appropriate histopathology of
reproductive organs were evaluated.
FO Male Rats
One FO male rat in the 30 mg/kg/day dose group was sacrificed on day 45 of the study
because of adverse clinical signs. Statistically significant increases in clinical signs also were
observed in male rats in the high-dose group, including dehydration, urine-stained abdominal fur,
and ungroomed coat. Significant reductions in body weight were reported beginning on post-
weaning day 50 at 3 mg/kg/day and for most of the study until termination in 10 and
30 mg/kg/day dose groups (6%, 11%, and 25% decrease from controls, respectively, at the end
of premating; p<0.05). Absolute food consumption was significantly reduced to approximately
91% of the control level during the study in the 30-mg/kg/day dose group but not for the lower
dose groups. Mean food consumption relative to body weight was increased in a dose-related
manner for all treated males with statistical significance at >3 mg/kg/day; overall relative food
consumption was 101, 105, 110, and 118% of the controls in the 1, 3, 10, and 30 mg/kg/day
groups, respectively. The body weight and food consumption effects were not observed in
female rats at any dose.
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Organ weight data for the FO male rats is shown in Table 3-19. The absolute and relative-to-
body and -brain weights of the liver were statistically significantly increased in all dose groups.
Absolute kidney weights were statistically significantly increased in the 1-, 3-, and 10-mg/kg/day
dose groups, but significantly decreased in the 30-mg/kg/day group. Organ weight-to-terminal
body weight ratios for the left and right kidney were statistically significantly increased in all
treated groups. Kidney weight-to-brain weight ratios were significantly increased at 1, 3, and
10 mg/kg/day, but decreased at 30 mg/kg/day, following the trends in absolute weights. In the
high-dose group, absolute and relative kidney weight changes occurred in a pattern typically
associated with decrements in body weight. However, in the lower dose groups, consistent
significant increases in absolute kidney weight and relative-to-body and -brain weights are a
response to the challenge of providing transporters for renal removal of the foreign molecule.
Increased kidney weight can be regarded as an adaptive response to the transport challenge. It is
beneficial for the individual but adverse in the sense that it signifies the need to upregulate
tubular transporters in the kidney to excrete the foreign material and a reflection of PFOA
bioaccumulation in serum and tissues.
Table 3-19. Organ Weight Data from FO Male Rats
Body weight (g)
Brain weight (g)
Liver weight (g)
Liver/body (%)
Liver/brain (%)
Rt. kidney (g)
Rt. kidney /body (%)
Rt. kidney /brain (%)
Lt. kidney (g)
Lt. kidney/body (%)
Lt. kidney/brain (%)
0 mg/kg/day
581 ±40
2.26 ±0.17
20.3 ±2.5
3.49 ±0.29
903 ±119
2.19±0.18
0.379 ±0.030
97.5 ±9.9
2.19 ±0.20
0.378 ±0.036
97.5 ± 10.7
1 mg/kg/day
575 ± 48
2.28 ±0.10
24.3 ±3.2"
4.22 ±0.50"
1066 ± 154"
2.54 ±0.30"
0.443 ±0.048"
111.6±13.5"
2.51 ±0.28"
0.437 ±0.047"
110.1 ± 12.6"
3 mg/kg/day
542 ± 47"
2.26 ±0.12
27.7 ±2.7"
5. 13 ±0.47"
1230 ± 120"
2.50 ±0.18"
0.463 ± 0.039"
111.0±9.5"
2.51 ±0.21"
0.465 ± 0.043"
111.7 ±10.5"
10 mg/kg/day
513 ±54"
2.24 ±0.12
28.7 ±3.9"
5.61 ±0.51"
1285 ± 183"
2.36 ±0.25"
0.462 ±0.034"
105.6 ±12.4"
2.34 ±0.24*
0.457 ± 0.040"
104.6 ± 11.7*
30 mg/kg/day
432 ± 64"
2.20 ±0.14
27.5 ±3.7"
6.42 ±0.73"
1248 ± 144**
2.06 ± 0.20*
0.481 ±0.051"
93.5 ±8.7
1.99 ±0.19"
0.466 ± 0.054"
90.4 ±8.7*
Source: Butenhoff et al. 2004a
Notes: Mean±SD; n = 29-30; significantly different from control: *p < 0.05, "p < 0.01.
The only histologic finding was increased thickness and prominence of the zona glomerulosa
and vacuolation in the cells of the adrenal cortex observed in 2/10 males in the 10-mg/kg/day
dose group and 7/10 males in the 30-mg/kg/day dose group.
No treatment-related effects were reported at any dose level for any of the male reproductive
parameters assessed. There was no evidence of altered testicular and sperm structure and
function in PFOA-treated FO rats with mean group serum PFOA concentrations of up to
approximately 45 |ig/mL (York et al. 2010). There was a significant dose-related increase in
seminal vesicle weight (p<0.05) with and without fluid in the Fl males, but fertility of the
exposed males in all generations was comparable to the controls.
No treatment-related effects were seen at necropsy or upon microscopic examination of the
reproductive organs.
Under the conditions of the study, the LOAEL for FO parental male rats is 1 mg/kg/day, the
lowest dose tested, based on significant increases in absolute and relative liver and kidney
weights. A NOAEL for the FO parental males could not be determined.
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FO Female Rats
There were no treatment-related effects on clinical signs, body weight, food consumption,
organ weights, or histology of the organs. There were no treatment-related effects on any of the
reproductive parameters assessed, and no treatment-related effects were seen at necropsy other
than slightly decreased liver weights (p<0.05) at doses of 3 and 10 mg/kg/day, but not
30 mg/kg/day. No abnormalities were seen with microscopic examination of the reproductive
organs. The NOAEL for FO parental females is 30 mg/kg/day, the highest dose tested.
Fl Generation
Pup body weight on a per-litter basis (genders combined) was significantly reduced (p<0.01)
by 8-10% throughout the first 2 weeks of lactation in the 30-mg/kg/day group; at weaning, the
mean body weight was reduced 4.5%, but the difference was not statistically significant.
Although there were no effects on the viability and lactation indices, the total number of dead
pups during lactation was increased in the 30-mg/kg/day groups; the difference was statistically
significant on LDs 6-8. No other effects were noted, and there were no treatment-related
findings for the pups necropsied at weaning. The offspring toxicity LOAEL is 30 mg/kg/day
based on decreased body weight and an increase in the number of dead pups; the NOAEL is
10 mg/kg/day.
Fl Male Rats
Significant increases in treatment-related deaths (5/60 animals) were reported in Fl males in
the high-dose group between days 2-4 postweaning. One rat was moribund sacrificed on day 39
postweaning and another was found dead on day 107 postweaning. Clinical signs included a
significant increase in emaciation at 10 and 30 mg/kg/day, and in urine-stained abdominal fur,
decreased motor activity, and abdominal distention at 30 mg/kg/day.
Mean body weight was significantly reduced in the 30-mg/kg/day group beginning on
postweaning day 8, in the 10-mg/kg/day group beginning on postweaning day 36, and towards
the end of the study in the 1- and 3-mg/kg/day groups. Terminal mean body weight was reduced
in all treated groups at the time of sacrifice. For all groups, there was a significant, dose-related
reduction in mean body weight gain for the entire dosing period (days 1-113). Absolute food
consumption values were significantly reduced at 10 and 30 mg/kg/day during the entire
precohabitation period (days 1-70 postweaning), while relative food consumption values were
significantly increased.
Statistically significant delays in the average day of preputial separation (p< 0.01) were
observed in high-dose animals versus concurrent controls (52.2 days of age versus 48.5 days of
age, respectively). There were no other effects on any of the reproductive parameters assessed,
and at necropsy no effects on reproductive organs or fertility were noted (York et al. 2010).
The absolute and relative weights of the liver were statistically significantly increased in all
treated groups (p < 0.01). Treatment-related microscopic changes were described as diffuse
hepatocellular hypertrophy in rats receiving doses of > 3 mg/kg/day. At the same dose levels,
there were scattered incidences of focal-to-multifocal necrosis and inflammation in the livers of
the Fl male rats. As in the FO males, the relative weight of the left and/or right kidneys was
statistically significantly increased compared to controls for all dose groups, except for the right
kidney at the high dose in which it was lower than for the controls (Table 3-20). Organ weight-
to-terminal body weight and brain weight ratios for the kidney were statistically significantly
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increased in all treated groups. All other organ weight changes observed (i.e., thymus, spleen,
left adrenal, brain, prostate, seminal vesicles, testes, and epididymis) were probably due to
decrements in body weight and not a reflection of target organ toxicity. Treatment-related
microscopic changes were observed in the adrenal glands of high-dose animals (i.e., cytoplasmic
hypertrophy and vacuolation of the cells of the adrenal cortex) and in the liver of rats treated
with 3, 10, and 30 mg/kg/day (hepatocellular hypertrophy).
Table 3-20. Organ Weight Data from Fl Male Rats
Body weight (g)
Brain weight (g)
Liver weight (g)
Liver/body (%)
Liver/brain (%)
Rt. kidney (g)
Rt. kidney/body (%)
Rt. kidney/brain (%)
Lt. kidney (g)
Lt. kidney/body (%)
Lt. kidney/brain (%)
0 mg/kg/day
560 ± 60
2.34 ±0.13
21.7 ±3.2
3.86 ±0.32
927 ± 136
2.24 ±0.21
0.402 ±0.034
95. 9 ±9.1
2.21 ±0.20
0.396 ±0.031
94.8 ±7.9
1 mg/kg/day
527 ±55*
2.28 ±0.16
24.6 ±4.0"
4.65 ±0.51"
1075 ± 150"
2.34 ±0.28
0.446 ±0.041"
102.6 ±7.7"
2.35 ±0.26*
0.446 ± 0.042**
102.8 ±7.6**
3 mg/kg/day
524 ± 48*
2.31±0.12
28. 2 ±4.2"
5.41 ±0.75"
1224 ± 179"
2.48 ±0.24"
0.474 ±0.041"
107.4 ± 10.2"
2.46 ± 0.20**
0.472 ± 0.045**
106.6 ±9.1"
10 mg/kg/day
499 ± 64"
2.28 ±0.10
29.3 ±4.1"
5. 90 ±0.70"
1285 ± 159"
2.33 ±0.25
0.469 ±0.050"
102.3 ±9.8*
2.30 ±0.22
0.464 ± 0.046**
101.0 ±7.9*
30 mg/kg/day
438 ±42"
2.18 ±0.14"
29.7 ±4.0"
6.79 ±0.55"
1364 ±166"
2.04 ±0.21"
0.467 ±0.036"
93.6 ±7.9
2.03 ±0.22"
0.465 ±0.038"
93. 3 ±10.0
Source: Butenhoff et al. 2004a
Notes: Mean±SD; n = 29-30; significantly different from control: *p<0.05, **p<0.01.
The LOAEL for adult systemic toxicity in the Fl males is 1 mg/kg/day based on significant,
dose-related decreases in body weights and body weight gains, and in terminal body weights;
and significant increases in absolute and relative kidney weights. ANOAEL for adult systemic
toxicity in the Fl males could not be determined. Liver weights were significantly increased at
all doses, but only accompanied by microscopic lesions at doses > 3 mg/kg/day.
Fl Female Rats
A statistically significant increase in treatment-related mortality (6/60 animals) was observed
in Fl females on postweaning days 2-8 at the highest dose of 30 mg/kg/day. No adverse clinical
signs of treatment-related toxicity were reported. Statistically significant decreases in body
weight were observed in high-dose rats on days 8, 15, 22, 29, 50, and 57 postweaning, during
precohabitation (recorded on the day cohabitation began, when Fl generation rats were 92-106
days of age), and during gestation and lactation. Body weight gain was significantly reduced
during days 1-8 and 8-15 postweaning. Statistically significant decreases in absolute food
consumption were observed during days 1-8, 8-15, and 15-22 postweaning, during
precohabitation, and during gestation and lactation in animals treated with 30 mg/kg/day.
Relative food consumption values were comparable across all treated groups.
Statistically significant delays (p<0.01) in sexual maturation (the average day of vaginal
patency) were observed in high-dose animals versus concurrent controls (36.6 days of age versus
34.9 days of age, respectively). Prior to the rats mating, the study authors noted a statistically
significant increase in the average numbers of estrous stages per 21 days in high-dose animals
(5.4 versus 4.7 in controls). For this calculation, the number of independent occurrences of estrus
in the 21 days of observation was determined. This calculation can be used as a screen for effects
on the estrous cycle, but should be followed by a more detailed analysis.
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Both 3M (2002, cited in USEPA [2005c]), and EPA (USEPA 2002b) conducted a more
detailed analysis of the estrous cycle data. The 3M analysis of the data concluded that there were
no differences in the number of females with >3 days of estrus or with >4 days of diestrus in the
control and high-dose groups. This conclusion is consistent with that of EPA (USEPA 2002b).
The cycles were evaluated as having either regular 4-5-day cycles, uneven cycling (defined as
brief periods with irregular pattern) or periods of prolonged diestrus (defined as 4-6-day diestrus
periods), extended estrus (defined as 3-4 days of cornified smears), possible pseudopregnancy
(defined as 6 or more days of leukocytes), or persistent estrus (defined as 5 or more days of
cornified smears). The two groups were not different in any of the parameters measured. Thus,
the increase in the number of estrus stages per 21 days noted by the study authors was an
outcome of the approach used for the calculations and is not biologically meaningful. There were
no effects on the other reproductive parameters assessed, and at necropsy no effects on
reproductive organs were noted.
No treatment-related effects were observed in the terminal body weights of the Fl female
rats. The absolute weight of the pituitary, the pituitary weight-to-terminal body weight ratio, and
the pituitary weight-to-brain weight ratio were statistically significantly decreased at 3
mg/kg/day and higher. Since there is not a clear dose-response relationship and histologic
examination reveal no lesions, the biological significance of the pituitary weight data is
problematic. No other differences were reported for the absolute weights or ratios for other
organs evaluated. No treatment-related effects were reported following macroscopic and
histopathologic examinations.
For Fl 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 Fl females are 10
and 30 mg/kg/day, respectively, based on statistically significant decreases in body weight and
body weight gains.
F2 Generation Rats
There was a statistically significant increase (p < 0.01) in the number of pups found dead on
LD 1 in the 3- and 10-mg/kg/day groups. An independent statistical analysis was conducted by
EPA (USEPA 2002c), and no significant differences were observed between dose groups and the
response did not have any trend in dose. There were no treatment-related effects on any of the
developmental parameters assessed, and at necropsy, no treatment-related effects were noted.
The NOAEL for developmental/reproductive toxicity in the F2 offspring was 30 mg/kg/day.
Mouse. In a follow-up to the 28-day component of the Lu et al. study (2015) of the impact on
PFOA on the blood testes barrier (section 3.2.2), the authors examined the impact exposure to
0 and 5 mg/kg/day PFOA had on the fertility of the treated male mice (6-8 weeks old; 15 per
dose group). Each treated male was mated with three virgin ICR females (8-10 weeks old).
Successful mating was determined by the presence of a vaginal plug in the morning. The
pregnant females were separated from the males and caged alone throughout the pregnancy; they
were not dosed with PFOA. At parturition, the pups were counted and the litter body weight
recorded. There was a statistically significant decrease in the number of mated females per male
mouse and pregnant females per male mouse (p<0.001) for the exposed males compared to the
controls. The average number of pups per litter was smaller for the exposed group (10.20 ± 0.72
versus 11.89 ± 0.54), but the difference was not significant. The average litter weight was
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significantly lower for the offspring of the paternally exposed mice than of the paternal controls
(16.17 ± 1.63 g versus 19.95 ± 0.80 g; p<0.05). The 5-mg/kg/day dose was a LOAEL for effects
on male fertility and the significantly lower body weight of their progeny.
Developmental Effects
Standard developmental studies in rats and mice found impacts on pup body weight and
developmental delays. Most studies used the oral route of exposure; one study used inhalation
exposure to PFOA particulate matter. Some examined the developmental impact of PFOA
associated with exposures that occurred only during gestation and lactation or during the
prepubertal period and the association of PPARa with the developmental effect spectrum. A
substantial number of studies in mice focused on PFOA's impact on mammary gland
development.
3.2.6 Prenatal Development
Rat. Pregnant Sprague-Dawley rats were gavage-dosed with 0, 3, 10, and 30 mg/kg/day PFOA
during days 4-10, 4-15, and 4-21 of gestation, or from GD 4 to LD 21 (Hinderliter et al. 2005;
Mylchreest 2003). Clinical observations and body weights were recorded daily. On GDs 10, 15,
and 21, five rats per group per time-point were sacrificed and the number, location, and type of
implantation sites were recorded. Embryos were collected on day 10, and placentas, amniotic
fluid, and embryos/fetuses were collected on days 15 and 21. Maternal blood samples were
collected at 2 hours ± 30 mins post-dose. The remaining five rats per group were allowed to
deliver. On LDs 0, 3, 7, 14, and 21, the pups were counted, weighed (genders separate), and
examined for abnormal appearance and behavior. Randomly selected pups were sacrificed and
blood samples were collected. On LDs 3, 7, 14, and 21, the dams were anesthetized and milk and
blood samples were collected; dams were removed from their litters 1-2 hours prior to
collection. Plasma, milk, amniotic fluid extract, and tissue homogenate (i.e., placenta, embryo,
and fetus) supernatants were analyzed for PFOA concentrations by UPLC/MS.
All dams survived and there were no clinical signs of toxicity. In the 30-mg/kg/day group,
mean body weight gain was approximately 10% lower than in the control group during gestation,
and mean body weight was approximately 4% lower than for controls throughout gestation and
lactation. The number of implantation sites, resorptions, and live fetuses were comparable among
groups on days 10, 15, and 21 of gestation. One dam in the 3-mg/kg/day group and two dams in
the 30-mg/kg/day group delivered small litters (3-6 pups per litter compared to 12-19 pups per
litter in the control group); however, given the small sample size, the biological significance of
this finding is unclear. There were no clinical signs of toxicity in the pups, and pup survival and
body weights were comparable among groups. Under the conditions of this study, the maternal
LOAEL was 30 mg/kg/day for decreased body weight gain during gestation, and the NOAEL
was 10 mg/kg/day. The developmental NOAEL was 30 mg/kg/day.
Mouse. Lau et al. (2006) conducted a developmental toxicity study of PFOA in mice to ascertain
whether there was a gender-related difference in the bioaccumulation of PFOA in the mouse and
to evaluate the effects of PFOA on prenatal and postnatal development in offspring exposed
during pregnancy. In that study, groups averaging 9-45 timed-pregnant CD-I mice were given 0,
1, 3, 5, 10, 20, and 40 mg/kg PFOA daily by oral gavage on GDs 1-17. Maternal weight was
monitored during gestation. Dams were divided into two groups. In the first group, dams were
sacrificed on GD 18 and underwent maternal and fetal examinations that included measure of
maternal liver weight and examination of the gravid uterus for numbers of live and dead fetuses
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and resorptions. Maternal blood was collected and analyzed for PFOA serum concentration.
PFOA levels in the fetuses were not examined. Live fetuses were weighed and subjected to
external gross necropsy and skeletal and visceral examinations. In the second group of dams, an
additional dose of PFOA was administered on GD 18. Dams were allowed to give birth on
GD 19.
The day following parturition was designated as PND 1. Time of parturition, condition of
newborns, and number of live offspring were recorded. The number of live pups in each litter
and pup body weight were noted for the first 4 days after birth and then at corresponding
intervals thereafter. Eye opening was recorded beginning at PND 12. Pups were weaned on PND
23 and separated by gender. The time to sexual maturity was determined by monitoring vaginal
opening and preputial separation beginning on PND 24. Two to four pups per gender per litter
were randomly selected for observation of postnatal survival, growth, and development. Estrous
cyclicity was determined daily by vaginal cytology. After weaning, dams were sacrificed and the
contents of the uteri examined for implantation sites. Postnatal survival was calculated based on
the number of implantations for each dam.
Signs of maternal toxicity were observed following exposure to PFOA during pregnancy.
Statistically significant dose-related increases (p < 0.05) in maternal liver weight also were
observed, beginning at 1 mg/kg/day. Dose-related decreases in maternal weight gain during
pregnancy were observed beginning at 5 mg/kg/day, with statistical significance (p < 0.05) seen
in the 20- and 40-mg/kg/day dose groups. Under the conditions of the study, a maternal LOAEL
of 1 mg/kg was indicated based on increased liver weight, and a NOAEL was not established.
Signs of developmental toxicity were observed following in utero exposure to PFOA. The
number of implantations in treated mice was comparable to control mice. Statistically significant
increases (p < 0.05) in full-litter resorptions were reported at doses of > 5 mg/kg/day, with
complete loss of pregnancies at the highest dose group of 40 mg/kg/day. A 20% reduction
(p < 0.05) in live fetal body weight at term was reported at 20 mg/kg/day. A statistically
significant increase in prenatal loss was observed in the 20-mg/kg/day dose group. Ossification
(number of sites) of the forelimb proximal phalanges was significantly reduced at all doses
except 5 mg/kg. Ossification of hindlimb proximal phalanges was significantly reduced at all
doses except 3 and 5 mg/kg. Reduced ossification (p < 0.05) of the calvaria and enlarged
fontanel was observed at 1, 3, and 20 mg/kg and at > 10 mg/kg in the supraoccipital bone.
Statistically significant increases (p < 0.05) in minor limb and tail defects were observed in the
fetuses at doses > 5 mg/kg/day. Under the conditions of the study, a prenatal developmental
LOAEL of 1 mg/kg was indicated based on increased skeletal defects, and the NOAEL was not
established.
Slight, but statistically significant, increases (p < 0.05) in the average time to parturition were
observed at 10 and 20 mg/kg/day. Increases (p < 0.05) in stillbirths and neonatal mortality (or
decreases in postnatal survival) were observed at doses > 5 mg/kg/day, with as much as a 30%
increase in these effects seen in the 10- and 20-mg/kg/day dose groups. Postnatal survival and
viability in the 1- and 3-mg/kg/day dose groups were comparable to controls. At doses
> 3 mg/kg/day, a trend in growth retardation (body weight reductions of 25-30%; p < 0.05), was
observed in the neonates at weaning. Body weights were at control levels by 6 weeks of age for
females and by 13 weeks of age for males. A trend for increasing body weight (-6-10% greater
than controls) was observed in animals dosed with 5 mg/kg at 13 weeks and in animals dosed
with 1 and 3 mg/kg at 48 weeks. Deficits in early postnatal growth and development also were
manifested by significant delays (p < 0.05) in eye opening at doses > 5 mg/kg/day. Slight delays
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(p < 0.05) in vaginal opening and in time to estrous were observed at 20 mg/kg/day in females;
in contrast, significant accelerations (p < 0.05) in sexual maturation were observed in males, with
preputial separation occurring 4 days earlier than controls at the 1-mg/kg/day dose and 2-3 days
earlier in the 3-10-mg/kg/day dose groups, but the 20-mg/kg/day dose group was only slightly
delayed compared to controls. Under the conditions of the study, a LOAEL for developmental
toxicity of 1 mg/kg/day for males was indicated based on accelerated pubertal development, and
a NOAEL was not established. For females, the developmental LOAEL was 3 mg/kg/day based
on growth retardation, and the NOAEL was 1 mg/kg/day.
Values for the benchmark dose (BMD) and the lower 95th percentile confidence bound on the
BMD (BMDL) for the maternal and developmental endpoints (BMDs and BMDLs) were
calculated by the study authors and reported in Lau et al. (2006). For maternal toxicity, BMDs
and BMDLs estimates for decreases in maternal weight gain during pregnancy were 6.76 and
3.58 mg/kg/day, respectively. For increases in maternal liver weight at term, BMDs and BMDLs
estimates were 0.20 mg/kg/day and 0.17 mg/kg/day, respectively. BMDs and BMDLs estimates
for the incidence of neonatal mortality (determined by survival to weaning) at 5 mg/kg/day were
2.84 and 1.09 mg/kg/day, respectively. Significant alterations in postnatal growth and
development were observed at 1 and 3 mg/kg/day, with BMDs and BMDLs estimates of 1.07 and
0.86 mg/kg/day, respectively, for decreased pup weight at weaning; and 2.64 and 2.10
mg/kg/day, respectively, for delays in eye opening. The BMDs and BMDLs estimates for reduced
phalangeal ossification were < 1 mg/kg/day. BMDs and BMDLs estimates for reduced fetal
weight at term were estimated to be 10.3 and 4.3 mg/kg/day, respectively.
Male and female 129Sl/SvImJ and PPARa-null mice were used in studies to determine if
PFOA-induced developmental toxicity was mediated by PPARa (Abbott et al. 2007). Pregnant
12981/SvlmJ wild-type and PPARa-null mice were orally dosed from GD 1-17 with 0, 0.1, 0.3,
0.6, 1, 3, 5, 10, and 20 mg PFOA/kg/day. Heterozygous (HET) litters also were produced by
mating wild-type and PPARa-null males with wild-type and PPARa-null dams to determine if
genetic background affected survival. The FtET litters were sacrificed on PND 15. Survival at
birth was recorded and live offspring counted and weighed by gender. Litters were counted and
offspring weighed on PND 1-10, 14, 17, and 22. Weaning occurred on PND 22, and all dams
and one pup per litter were sacrificed. Blood was collected and the uteri were stained for
implantation counts.
There was no effect of treatment on maternal weight or maternal weight gain (excluding
nonpregnant females and those with full-litter resorptions), number of implants, or pup weight at
birth. Wild-type dams exposed to >0.6 mg/kg/day and PPARa-null dams exposed to >5
mg/kg/day had a significantly greater percentage of litter loss compared to their respective
controls. At >5 mg/kg/day in wild-type dams and 20 mg/kg/day in PPARa-null dams, 100%
litter loss occurred. Relative liver weight was significantly increased in wild-type adult females
dosed with >1 mg/kg/day and in PPARa-null adult females dosed with >3 mg/kg/day.
Body weight in wild-type offspring born of dams dosed with 1.0 mg/kg/day was significantly
reduced (p<0.05) compared to control offspring body weight gain on PND 9, 10, and 22 (males)
and PND 7-10 and PND 22 (females). No differences were observed between PPARa-null
offspring body weight and control offspring body weight. Survival of pups from birth to weaning
was significantly reduced (p<0.05) in wild-type litters exposed to >0.6 mg/kg/day, but was not
affected in PPARa-null litters. Survival was significantly decreased (p<0.05) for wild-type and
FIET pups born to wild-type dams dosed with 1 mg/kg/day and for FtET pups born to PPARa-
null dams dosed with 3 mg/kg. Offspring born of wild-type dams showed a dose-related trend for
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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 PPARa-
null dams. At weaning, relative liver weight was significantly increased (p<0.05) in wild-type
offspring gestationally exposed to >0.1 mg/kg/day and in PPARa-null offspring gestationally
exposed to 3 mg/kg/day.
The authors concluded that survival of PPARa-null pups and deaths of HET pups born to
PPARa-null dams indicates that expression of PPARa is required for PFOA-induced postnatal
lethality; however, early prenatal lethality was independent of PPARa. Delayed eye opening and
reduced postnatal weight gain appeared to be mediated by PPARa, but other mechanisms might
also contribute. Under the conditions of the study, the maternal/reproductive LOAEL for wild-
type mice was 0.6 mg/kg/day based on increased percentage of litter loss, and the NOAEL was
0.3 mg/kg/day. The developmental LOAEL for wild-type offspring was 0.1 mg/kg/day based on
increased liver weight, and the NOAEL was not established. The maternal LOAEL for PPARa-
null mice was 3 mg/kg/day based on increased liver weight, and the NOAEL was 1 mg/kg/day.
The developmental LOAEL for PPARa-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 PPARa, groups of
female wild-type, PPARa-null, and PPARa-humanized mice were given 0 and 3 mg PFOA/kg
on GDs 1-17 by oral gavage (Albrecht et al. 2013). Controls received the water vehicle. Females
were either sacrificed on GD 18 (n = 5-8 per group) or allowed to give birth and then sacrificed,
along with their litters (n = 8-14), on PND 20. Livers from dams, fetuses, and pups were
weighed and collected for histopathological evaluation and RNA analysis. Gene expression
results are given in section 3.3.4. Mammary gland whole mounts were prepared from female
pups on PND 20 for quantification of ductal length and number of terminal end buds; these
results are described below with other studies evaluating the effects of PFOA on mammary gland
development.
Evaluation on GD 18 showed no effects of PFOA administration on maternal body weight,
body weight gain, gravid uterine weight, number of implantations per dam, or number of
resorptions per litter in dams of any genotype. For animals allowed to litter, the average day of
parturition was slightly later in PFOA-treated humanized mice than in the controls. Body weight
of dams during lactation, the number of pups born per litter, pup body weight during lactation,
and the onset of pup eye opening were similar between treated and control groups for all
genotypes. Offspring survival during PNDs 1-5 was significantly reduced in the wild-type
PFOA-treated group, but not in the other genotypes.
Maternal liver weight was significantly increased in the treated groups of all genotypes on
GD 18 and in wild-type animals on PND 20. Maternal liver weight was not affected on PND 20
in the PPARa-null or PPARa-humanized mice. Relative fetal liver weight on GD 18 was
significantly increased in fetuses from treated wild-type and humanized dams. On PND 20,
relative liver weight was increased only in pups from treated wild-type dams. Microscopic
evaluation of the maternal liver showed centrilobular hepatocellular hypertrophy in all PFOA-
treated groups on GD 18 and PND 20, with decreased incidence and severity by PND 20. On GD
18, the liver lesions were graded as mild in the wild-type mice, minimal-to-mild in the
humanized mice, and minimal in the null mice. The morphological features of the liver lesions
differed slightly between genotypes and are described in more detail in section 3.4.1. Only wild-
type fetuses and pups from treated dams showed similar liver lesions.
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Yahia et al. (2010) gavage-dosed pregnant ICR mice (n = 5 per group) with 0, 1, 5, and
10 mg PFOA/kg/day from GDs 0-17 or 0-18. The dams dosed from GDs 0-17 were sacrificed
on GD 18, and the fetal skeletal morphology was evaluated. Dams dosed from GDs 0-18 were
allowed to give birth and their offspring were either processed for pathological examination or
observed for 4 days for neonatal mortality. Maternal liver, kidney, brain, and lungs were
histologically examined after necropsy. Serum was collected for clinical chemistry and lipid
analysis. Body weight was significantly decreased in dams receiving 10 mg/kg/day. Maternal
absolute liver weight was significantly increased (p<0.05) at doses > 5 mg/kg/day and relative
liver and kidney weights were significantly increased at all doses. Hepatic hypertrophy was
localized to the centrilobular region at the two lower doses and was diffuse at the highest dose.
Renal cells in the outer medullar and proximal tubule were slightly hypertrophic at all doses.
Treatment with 10 mg/kg/day caused a significant increase in AST, ALT, GGT, and ALP and a
significant decrease in total serum protein, albumin, globulin, triglycerides, phospholipids, TC,
and free fatty acids. At 5 mg/kg/day, total serum protein and globulin were significantly
decreased, and phosphorus was increased. At 1 mg/kg/day, BUN and phosphorus were
significantly increased. The maternal LOAEL was 1 mg/kg/day based on significantly increased
relative liver and kidney weight, and no NOAEL was established.
Live fetal birth weight was significantly decreased at the two highest doses. There was no
difference in the percentage of live fetuses between treated and control groups. At 10 mg/kg/day,
increased incidence of cleft sternum, reduced phalanges ossification, and delayed eruption of
incisors was observed. Delayed parturition was observed in dams treated with 10 mg/kg/day, and
-58% of all pups born to those dams were stillborn. Death occurred within 6 hours of delivery in
the remaining pups, and whole body edema was observed in some of the pups. The body weight
of the live pups born to dams treated with 5 or 10 mg/kg/day was significantly reduced compared
to control pup body weight. By PND 4, 16% of offspring born to dams dosed with 5 mg/kg/day
had died. No pathological differences were observed in the lungs or brains of treated and control
offspring. The developmental LOAEL was 5 mg/kg/day based on decreased body weight and
decreased survival rate, and the NOAEL was 1 mg/kg/day.
Suh et al. (2011) examined placental prolactin-family hormone and fetal growth retardation
in mice treated with PFOA. Pregnant CD-I mice (n = 10 per group) were treated with 0, 2, 10,
and 25 mg/kg/day PFOA from GDs 11-16. Dams were sacrificed on GD 16 and uteri were
removed and examined. Three placentas per group were analyzed histochemically and the
numbers of glycogen trophoblast cell (GlyT) in the junctional zone plus sinusoidal trophoblast
giant cells (S-TGC) in the labyrinth zone were counted and compared. Trophoblast cells express
prolactin-family genes. mRNA from three placentas per group were analyzed using situ
hybridization, northern blot hybridization, and RT-PCR for mouse placental lactogen- (mPL-) II,
mouse prolactin-like protein- (mPLP-) E and F, Pit-la, and P isotype (transacting factors of mPL
and mPLP genes).
A significant difference in maternal body weight was observed from GD 13-16 in dams
treated with 25 mg/kg/day of PFOA compared to controls. At >2 mg/kg/day of PFOA, placental
weight was significantly decreased and the number of resorptions and dead fetuses was
significantly increased. At >10 mg/kg/day of PFOA, fetal weight and the number of live fetuses
were significantly decreased. There were no differences in the number of implantation sites
among the groups, and postimplantation loss was 3.87, 8.83, 30.98, and 55.41% for the 0, 2, 10,
and 25 mg/kg/day PFOA groups, respectively.
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The placentas of dams dosed with >10 mg/kg/day of PFOA displayed necrotic changes.
Parietal and S-TGC and GlyT cell frequency in the placental junctional and labyrinth zones was
significantly decreased (p<0.05) in a dose-dependent manner in treated dams. At 25 mg/kg/day
of PFOA, S-TGCs showed signs of atrophy with crushed cell nucleus. A significant dose-
dependent decrease in mPL-II, mPLP-E, mPLP-F, and Pit-la and P isotype mRNA and
expression was observed. Correlation coefficients between fetal weight and maternal mPL-II,
mPLP-E, and mPLP-F mRNA levels were positive (p<0.001). Based on the results, the authors
suggested that inhibited prolactin-family gene expression could be secondary to insufficient
trophoblast cell differentiation and increased cell necrosis. These effects reduced placental
efficiency and contributed in part to fetal growth retardation. The 2-mg/kg/day dose was a
LOAEL for increases in resorption and dead fetuses plus decreased placental weight. There was
no NOAEL.
A meta-analysis was conducted to determine whether developmental exposure to PFOA was
associated with fetal growth effects in animals (Koustas et al. 2014). Eight studies identified in
the published literature met the criteria of the Navigation Guide systematic review methodology
for inclusion in the analysis (Woodruff and Sutton 2014). The data sets included mouse gavage
studies with offspring body weight at birth. Maternal PFOA doses ranged from 0.01 to
20 mg/kg/day. The results from the meta-analysis showed that a 1-mg/kg/day increase in PFOA
dose was associated with a -0.023 g (95% CI -0.029, -0.016) difference in pup birth weight. All
of the studies included in the analysis are described either above with the developmental toxicity
studies or with the specialized developmental studies that follow.
3.2.7 Mammary Gland Development and Other Specialized Developmental Studies
The following studies used experimental study designs and/or examined endpoints not
typically included in standard developmental toxicity studies. The studies were conducted to
determine critical periods of exposure for outcomes that occurred later in life. A number of the
studies focused on mammary gland development in dams and female offspring. Researchers
focused on effects resulting from indirect exposure of offspring via treatment of pregnant
animals and/or direct exposure of peripubertal animals starting about the time of weaning.
Indirect gestational and/or lactational exposures
Many studies evaluating indirect gestational and/or lactational exposure to PFOA are
available and Table 3-21 provides an overview of experiments designed to assess the
developmental effects of PFOA following exposures during gestation. Most of the studies focus
on mammary gland effects as a consequence of gestational and lactational or prepubertal
exposures in CD-I mice. Some have included postnatal assessment of the liver. Additional
details of the studies are described in the section following the table.
Timed pregnant CD-I mice were given PFOA by gavage at doses of 0, 0.01, 0.1, 0.03, and
1 mg/kg/day from GD 1 through the end of lactation (PND 21) in a study by Quist et al. (2015).
The litters were equalized on PND 4 to four females and six males. Only the females continued
in the study after weaning. At the end of lactation, litters with less than four females were
removed from the study. On PND 21, seven to 10 female pups per dose group were sacrificed by
decapitation. The livers were removed for analysis.
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Table 3-21. Studies of Pregnant CD-I Mice Following Administration of PFOA
Dose
(mg PFOA
/kg/day)
0,0.01.0.1,
0.3, 1
0,5
0,5
20
0,3,5
0,5
0,0.3, 1.0,3
0,0.01,0.1, 1
0,1,5
0, 1
+ 5 ppb in
drinking water
to both groups
0,3
0,0.01,0.1,
0.3, 1
Timing
GD ltoPND21
GD 1-17, 8-17, or 12-17
GD7-17, 10-17, 13-17,
15-17
GD 15-17
GD 1-17
Cross-fostered at birth
GD 8-17
Cross-fostered at birth
GD 1-17
GD 10-17
GD 1-17
GD 1-17
Drinking water started on
GD 7 and continued to F2
generation
GD 1-17
GD 1-17
Study included both CD-I
and C57BL/6 mice
Endpoints
Liver histopathology; periportal inflammation;
clinical chemistry; impact of postweaning HFD
Body weight; mammary gland morphology GD
18 (dams) and PNDs 10 and 20 (dams, female
pups)
Body weight; developmental landmarks and
growth to PND 189; mammary gland
morphology of female pups up to 18 months
Body weight; developmental landmarks and
growth to PND 245; mammary gland
morphology of female pups up to 18 months
Mammary gland morphology of dams and female
pups on PNDs 1, 3, 5, and 10
Liver weight; mammary gland morphology of
female pups on PNDs 7, 14, 21, 28, 42, 63, and
84
Body weight; reproductive parameters;
mammary gland morphology of FO, Fl, and F2
females
Wild-type, PPARa-null, and hPPARa w/129
mice; pup body weight at PNDs 14 and 20 plus
mammary gland structure
Body weight; net body weight; absolute and
relative liver weight on PNDs 21, 35, and 56;
serum estradiol and progesterone; mammary
gland morphology
Reference
Quistetal. 2015
White et al. 2007
White et al. 2009;
Wolfetal. 2007
White et al. 2009;
Wolfetal. 2007;
White et al. 2009
Maconetal. 2011
White etal. 2011
Albrecht et al.
2013
Tucker et al.
2015
During the lactation period all pups received a Purina control diet with a normal fat content.
At sacrifice on PND 21, there was a dose-related increase in relative liver weight at doses
>0.3 mg/kg/day. When the animals from the same dose group were sacrificed on PND 91, there
was no observed impact on relative liver weight.
Starting on PND 35, one pup from each of 20 dams was placed on a HFD (with 60% of the
calories from fat) until sacrificed on PND 91. Half were fasted for 4 hours before blood
collection and sacrifice. Another seven to 10 pups per dose group received the Purina control diet
with lower fat content until their sacrifice on PND 91. Blood samples were collected at sacrifice
for determination of ALP, AST, ALT, SDH, LDL, HDL, cholesterol, triglycerides, total bile acid
(TEA), glucose, leptin and insulin. Liver sections were collected for histological analysis and
graded 1 to 4 for lesion severity (1 = minimal; 4 = severe). Selected samples of the liver tissues
(four from the HFD groups and four from the PFOA-exposed control diet group) were fixed in
osmium trioxide and prepared for evaluation using transmission electron microscopy.
At PND 91, the animals on the PFOA + HFD weighed more than the Purina controls. The
body weight for the group that received PFOA and the control diet did not differ from the
untreated controls on the same diet. Serum samples from the PFOA-treated Purina controls, and
the fasted high-fat and nonfasted high-fat groups were analyzed for LDL, HDL, TC, and leptin.
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At PFOA doses < 0.3 mg/kg/day, the LDL and TC levels in the fasted and nonfasted HFD
animals were greater than in the untreated Purina controls. There was no impact of PFOS on
either parameter in animals on the PFOA plus Purina control diet, but both LDL and TC were
statistically lower at the high PFOA dose than they were at the low doses for both parameters. A
similar pattern was seen for the HDL levels.
The impact of PFOA dose on leptin was variable and not significant for the PFOA plus
Purina control animals and the high-fat, fasted animals. For the high-fat, nonfasted animals, there
was a trend towards decreasing leptin as the PFOA dose increased, which was significant at the
high dose of 1 mg/kg/day (P< 0.01). In those animals, the liver showed chronic periportal
inflammation and microvescicular intracytoplasmic lipid droplets. The transmission electron
microscopy slides showed that the hypertrophic liver cells presented evidence of cellular damage
and changes in both mitochondrial morphology and numbers. The observed mitochondrial
abnormalities were not those generally associated with PPARa activation. The 0.01 mg/kg/day
dose was a NOAEL. The LOAEL was 0.3 mg/kg/day for the effects on TC for animals receiving
a FIFD, but not for those receiving the PFOA plus Purina control diet.
Effects of PFOA exposure on mammary gland morphology of CD-I mice were evaluated in a
series of studies by the same researchers (Macon et al. 2011; Tucker et al. 2015; White et al.
2007, 2009, 2011). The focus was on mammary gland development of female pups, although
limited evaluations were conducted on the dams. Mammary gland whole mounts were scored on
a 1 to 4 subjective, age-adjusted, developmental scale (1 = poor development/structure;
4 = normal development/structure). Tissue was assessed qualitatively for the gross presence of
several histological criteria by two independent scorers and a mean score calculated. Neither
standardization of these subjective measures nor training of the scorers was described in the
publications. Quantitative measures of longitudinal growth, lateral growth, and number of
terminal end buds also were made in the Macon et al. (2011) and Albrecht et al. (2013) studies.
White et al. (2007) orally dosed pregnant CD-I mice with 0 and 5 mg PFOA/kg/day on GD
1-17 (n = 14), 8-17 (n = 16), and 12-17 (n = 16) to determine if decreased neonatal body
weights and survival were linked to gestational exposure or lactational changes in milk quantity
or quality. The control mice (n = 14) were dosed with vehicle on GD1-17. A subset of dams was
sacrificed on GD 18. The remaining dams were allowed to give birth, and pups were pooled and
randomly redistributed among the dams of the respective treatment groups. Litters were
equalized to 10 pups per litter. Half of the dams and litters were sacrificed on PND 10, and the
remaining dams and litters were sacrificed on PND 20. Mammary glands were collected from
dams and female pups at time of sacrifice.
Treatment with PFOA did not affect maternal weight gain, number of implants, or the
number of live fetuses. There was a significant increase (p<0.05) in prenatal loss in dams
exposed during GD1-17. Body weight of pups exposed gestationally to PFOA was significantly
decreased (p<0.05) at all time points measured and for all dosing regimens. On GD 18, stunted
alveolar development was observed in the mammary gland of dams treated with PFOA on GD
1-17 compared to the mammary glands of the control dams, which were saturated with milk-
filled alveoli. Dams treated with PFOA on GD 1-17 or 8-17 exhibited significant mammary
gland epithelial differentiation delays on PND 10 as evidenced by an excessive amount of
adipose tissue. In comparison, mammary glands from control dams on PND 10 were well
differentiated, full of alveoli filled with milk, and contained few apoptotic bodies and little
adipose tissue. The mammary gland developmental score in dams treated on GD 12-17 was not
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statistically different from control dams on PND 10. At PND 20, the mammary gland scores
from all PFOA-treated dams were not significantly different from those of the control group.
The pups were impacted by their in utero PFOA exposure over all dosing intervals. Their
mammary glands exhibited significantly stunted epithelial branching and longitudinal growth at
PNDs 10 and 20; the resulting developmental scores were significantly less than those of
controls. Very little mammary gland development occurred between PND 10-20 in the offspring
of dams exposed to PFOA, even though postnatal growth and body weight gain increased in
parallel to that of the controls. Thus, at the only dose tested, 5 mg/kg/day, effects were observed
on the dam and pup mammary gland, accompanied by decreased pup body weight and decreased
survival for the pups exposed during GD 1-17.
In the study by Wolf et al. (2007), CD-I mice were orally dosed with 0 and 5 mg
PFOA/kg/day on GD 7-17 (n = 14), 10-17 (n = 14), 13-17 (n = 12), and 15-17 (n = 12) or with
20 mg/kg on GD 15-17 (n = 6) to determine if there was a specific window during which PFOA
exposure produced developmental effects. The developmental results from this study were
published by Wolf et al. (2007) and the mammary gland effects were published by White et al.
(2009). On PND 22, all dams and one male and female pup from each litter were necropsied.
Blood samples were collected and livers from dams and offspring were removed and weighed.
Uterine implantation sites were counted. The fourth and fifth inguinal mammary glands were
removed from female offspring and analyzed at various intervals up to 18 months of age (White
et al. 2009). Mammary gland whole mounts from female offspring between PNDs 22 and 32
were scored as described above; whole mounts from lactating dams and female offspring at 18
months were qualitatively examined with respect to concurrent controls.
Maternal weight gain was increased in dams exposed to PFOA beginning on GDs 7 and 10,
but there was no effect on number of uterine implantation sites, litter loss, or number of pups per
litter at birth. Male pup weight at birth was significantly decreased (p<0.05) in dams dosed with
5 mg/kg/day on GD 7-17 or 10-17 or with 20 mg/kg/day on GD 15-17. By PND 78, all male
offspring had recovered to control body weight levels. On PND 161, the offspring of dams dosed
during GD 13-17 weighed significantly more than control. Litters exposed to 20 mg/kg/day on
GD 15-17 experienced decreased survival (p<0.01) during PND 1-22. Maternal relative liver
weight was significantly increased in all PFOA-treated dams except those treated during GD
15-17. Relative liver weight in all male and female pups was significantly increased (p<0.01).
Eye opening and growth of body hair were delayed in pups exposed GD 7-17 and 10-17 (Wolf
et al. 2007).
Mammary gland developmental scores for all offspring of dams exposed to PFOA were
significantly lower compared to scores from offspring of control dams at PND 29 and 32.
Delayed ductal elongation and branching and delayed appearance of terminal end buds were
characteristic of delayed mammary gland development at PND 32. At 18 months of age,
mammary tissues were not scored due to a lack of a protocol applicable to mature animals.
However, there were dark foci (composition unknown) in the mammary tissue that occurred at a
higher frequency in the exposed animals compared to controls, but did not display a consistent
response with dosing interval. Qualitatively, mammary glands from treated dams on LD 1
appeared immature compared with control dams (White et al. 2009). The 5-mg/kg/day dose was
associated with increased maternal and pup liver weight, altered pup mammary gland
development, and delayed pup eye opening and growth of body hair. The 20-mg/kg/day dose
was associated with decreased postnatal pup survival.
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The objective of a second component of the study by Wolf et al. (2007) and White et al.
(2009) was to determine if postnatal body weight deficits, neonatal lethality, and developmental
delays caused by PFOA exposure were the result of gestational exposure, lactational exposure, or
a combination of gestational and lactational exposure. Pregnant CD-I mice were orally dosed
with 0 (n = 48), 3 (n = 28), and 5 (n = 36) mg PFOA/kg/day on GD 1-17 and their offspring
cross-fostered at birth to create seven treatment groups: control, in utero exposure only (3U and
5U), lactation exposure only (PFOA stored in milk during gestation and released during
lactation; 3L and 5L), and in utero and lactation exposure (3U+L and 5U+L). On PND 22, all
dams and one male and female pup from each litter were necropsied. Blood samples were
collected and the liver was removed from dams and offspring and weighed. Implantation sites
were counted from the uteri of dams. The fourth and fifth inguinal mammary glands were
removed from female offspring and analyzed at various intervals up to 18 months of age (White
et al. 2009). Mammary gland whole mounts from female offspring between PND 22 and 63 were
scored as described above; whole mounts from female offspring at 18 months were qualitatively
examined with respect to concurrent controls due to lack of an applicable protocol for mature
animals.
Maternal weight and weight gain were higher in PFOA-treated dams compared to control
dams. Whole litter loss was significantly increased (p<0.05) at 5 mg/kg/day, but no differences
in the number of implantation sites were observed between the treated and control mice.
Absolute and relative liver weights of PFOA-treated dams from both dose groups were
significantly increased (p<0.001) compared to absolute and relative liver weights of control dams
23 days after the last dose (PND 22). No difference in the number of live pups born per litter was
found between treated and control mice, but male and female pup birth weight was reduced
(p<0.01) in dams receiving 5 mg/kg/day (Wolf et al. 2007). The 3-mg/kg/day dose was a
LOAEL for increases in liver weight in the dams while 5 mg/kg/day was a LOAEL for the pups,
based on whole litter loss and significantly reduced male and female birth weight.
A dose-dependent increase of PFOA was observed in the serum of dams treated with PFOA,
providing a reservoir for lactational transfer. The control dams that nursed offspring exposed in
utero (3U and 5U) had low concentrations of PFOA in their serum that originated from maternal
grooming behavior of the pups and allowed for low-level lactational transfer.
Body weight of male and female pups (3U+L, 5U, and 5U+L) was significantly reduced as
early as PND 2 and 1, respectively, and remained reduced throughout the lactation period. Body
weight recovery to control levels was reached by male offspring within 2 weeks of weaning, but
recovery in female offspring in the 5U and 5U+L groups did not occur until after PND 85.
Postnatal survival in 5U+L pups was significantly decreased compared to control survival
beginning at PND 4 and continuing throughout lactation. Survival in the other groups was no
different than control survival. Eye opening and body hair growth were significantly delayed in
the 3U+L, 5U, and 5U+L offspring. The relative liver weight was significantly increased in all
offspring regardless of exposure scenario (Wolf et al. 2007).
All female offspring of PFOA-exposed dams had reduced mammary gland developmental
scores at PND 22, except for females in the 3L group. At PND 42, mammary gland scores from
females in the 3U+L group were the only ones not statistically different from control scores. This
might have been due to interindividual variance and multiple criteria used to calculate mammary
gland development scores. All offspring of dams exposed to PFOA exhibited delayed mammary
gland development at PND 63, including those exposed only through lactation (3L and 5L). A
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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-I mice orally dosed with 0
(n = 56) and 5 (n = 56) mg PFOA/kg/day from GD 8-17 to determine the timing of the
mammary gland development deficits observed following gestational or lactational exposure to
PFOA. The groups were cross-fostered at birth to create four treatment groups: control, in utero
exposure only (5U), lactation exposure only (5L), and in utero and lactation exposure (5U+L).
Dams and litters were sacrificed on PNDs 1, 3, 5, and 10. Blood and liver samples were collected
for PFOA analysis. The fourth and fifth inguinal mammary glands were collected from dams and
female offspring and analyzed as described above; whole mounts from lactating dams were
qualitatively examined with respect to concurrent controls.
Maternal weight gain in treated dams was significantly higher than control weight gain, but
there were no effects of treatment on litter size or pup birth weight at PND 1. Significantly
decreased body weight occurred in the pups of the 5U+L group on PND 3 and in all PFOA-
exposed pups on PNDs 5 and 10. Relative liver weight of the treated dams was significantly
increased (p<0.05) compared to relative liver weight of control dams. On PND 1, liver-to-body
weight ratios were significantly increased (p<0.05) in pups exposed in utero (5U, 5U+L); serum
PFOA levels were 65,000-70,000 ng/mL. The liver-to-body weight ratio was increased in pups
exposed lactationally by PND 5; serum PFOA levels were approximately 15,000 ng/mL (White
et al. 2009).
On PND 1, the mammary glands of PFOA-exposed dams were qualitatively similar to glands
seen in late pregnancy, prior to parturition. In control dams nursing offspring from PFOA-
exposed dams, reduced alveolar filling was noted as early as PND 3, presumably a result of
exposure of the dam from maternal grooming behavior. The delayed lactational morphology in
dams treated with PFOA and control dams nursing offspring from PFOA-treated dams was
persistent up to PND 10 (terminal sacrifice). Reduced mammary gland developmental scores
were observed as early as PND 1 in all female offspring from PFOA-exposed dams, including
those exposed through lactation only (5L). Delayed mammary gland development persisted
throughout the study duration (White et al. 2009).
Macon et al. (2011) gavage-dosed CD-I mice with 0, 0.3, 1.0, and 3.0 mg PFOA/kg/day
from GD 1-17 (n = 13 dams per group). Six offspring per group were sacrificed on PNDs 7, 14,
21, 28, 42, 63, and 84, and blood, liver, brain, and the fourth and fifth mammary glands were
collected from female pups. Mammary gland developmental scores were not included in the
published article, but were available in supplemental materials.
Body weight in male and female offspring was not affected through PND 84. Absolute liver
weight was significantly increased at >0.3 mg/kg/day in females and at >1.0 mg/kg/day in males
on PND 7, and at 3.0 mg/kg in females at PND 14. Relative liver weight was significantly
increased at >0.3 mg/kg/day in males and females on PND 7, at >1.0 mg/kg/day in females on
PND 14, and at 3.0 mg/kg/day in males and females on PNDs 14, 21, and 28. No dose-related
differences were observed in absolute and relative brain weights.
Delayed mammary gland development of female pups was observed as early as PND 7 at
>1.0 mg/kg/day and PND 14 at >0.3 mg/kg/day and persisted until the end of the study.
However, the developmental scores did not show dose-related trends at each interval. The
delayed development was characterized by reduced epithelial growth and the presence of
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numerous terminal end buds. Photographs of the mammary gland whole mounts at PNDs 21 and
84 show differences in the duct development and branching pattern of offspring from dams given
0.3 and 1 mg/kg/day (offspring from high-dose dams not pictured). The LOAEL was
0.30 mg/kg/day based on significantly increased liver weight and delayed mammary gland
development. No NOAEL was established. The lowest dose tested was a NOAEL at PND day 7
and is a LOAEL at day PND 14.
Maconetal. (2011) also gavage-dosed CD-I mice with 0, 0.01, 0.1, and 1.0 mg
PFOA/kg/day from GD 10-17 (n = 5-8 dams per group) to examine the effects of low doses of
PFOA on mammary gland development. Female offspring (one from at least three litters per
group) were sacrificed on PNDs 1, 4, 7, 14, and 21, and blood, liver, and the fourth and fifth
mammary glands were collected. In addition to the qualitative mammary gland developmental
scores, quantitative measurements of longitudinal growth, lateral growth, and numbers of
terminal end buds and terminal ends were recorded. These data were presented only for animals
sacrificed on PND 21.
No differences in body weight or brain weight were observed for male or female offspring.
At 1 mg/kg, absolute and relative liver weights were significantly increased at PNDs 4 and 7.
Relative liver weight also was significantly increased at PND 14. Mammary gland development
was delayed by exposure to PFOA, especially longitudinal epithelial growth. At PND 21, all
treatment groups had significantly lower developmental scores. At the highest dose, poor
longitudinal epithelial growth and decreased number of terminal end buds were observed. As
seen in Table 3-22, the quantitative measures were statistically significant only for the high dose
compared to the controls, while the qualitative scores were significantly different from controls
at all doses. The LOAEL was 0.01 mg PFOA/kg/day based on the qualitative / quantitative
developmental score for mammary gland development and 1 mg/kg/day based on the
quantitative score in the absence of the qualitative component. No NOAEL was established.
Table 3-22. Mammary Gland Measurements at PND 21 from Female Offspring of
Dams Treated GD 10-17
Dose
mg/kg/d
On=5
0.01n=4
O.ln = 3
1.0 n = 5
Score
3.3 ±0.3
2.2 ± 0,2*
1.8 ±0.3"
1.6 ±0.1"*
Longitude
jim
4321 ±306
3803 ± 386
3615 ±320
2775 ± 285**
Lateral
jim
5941 ±280
5420 ± 326
4822 ± 672
4822 ±3 13"
Longitude = longitudinal epithelial growth
Lateral = Lateral epithelial growth
A =change in
*= p<0.05, " p<0.01, "*p<0.001
A Longitude
jim
3394 ±306
3087 ± 386
2370 ± 320
1553 ±301
A Lateral
jim
4358 ±280
3899 ± 326
3035 ± 672
3380 ±313
TEBs
#/gland
40 ±4
33 ±4
24 ±4
15 ±2"*
TEs
#/gland
81 ±12
61±8
58 ±4
47 ±11
# = number
TEBs = terminal end buds
TE = differentiating duct ends
Source: Macon et al. 2011
White et al. (2011) examined the extended consequences of PFOA-induced altered mammary
gland development in a multigenerational study in CD-I mice. Pregnant mice (FO, n = 10-12
dams per group) were gavage-dosed with 0, 1, and 5 mg PFOA/kg/day from GD 1-17. A
separate group of pregnant mice (n = 7-10 dams per group) was gavage-dosed with either 0 or
1 mg PFOA/kg/day from GD 1-17 and received drinking water containing 5 ppb PFOA
beginning on GD 7. Fl 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 Fl breeding
and early gestation, to simulate a chronic low-dose exposure. Only the FO dams were given
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PFOA by gavage. Total doses were not calculated for the groups receiving drinking water with
5 ppb PFOA. Table 3-23 shows the array of dosing regimens used in the study and the estimated
average daily PFOA intake by FO dams. The average daily intake from the chronic water
exposures were calculated from total weekly water consumption, divided by the number of days
per week (values given in supplemental materials; intake by the Fl animals was not calculated).
Table 3-23. Dosing Regimens Used in the Multigeneration Study of CD-I Mice
Treatment
Dose
Duration
Gavage
Drinking
water
Total Daily PFOA
intake to dams
from gavage and
drinking water
FO Dams
Gavage
0, I,or5
mg/kg/day
GD 1-17
None
Not relevant (0
mg/kg/day)
36 ug/day (1
mg/kg/day)
187 ug/day (5
mg/kg/day
FO Dams -> Fl Offspring
Drinking water
0+5 ppb
GD 1-17
GD 7-LD 22
0.054 ug/day
(gestation)
0.105 ug/day
(lactation)
Gavage +
drinking water
1+5 ppb
GD 1-17
GD 7-LD 22
37 + 0.051
ug/day
(gestation)
0+0.130 ug/day
(lactation)
Fl Dams -> F2
Offspring
Drinking water
5 ppb
None
Through LD 22
Not calculated
F2 Offspring
Drinking
water
5 ppb
None
Through PND
63
Not calculated
Source: White etal. 2011
FO females were sacrificed on PND 22. Fl offspring were weaned on PND 22 and bred at 7-
8 weeks of age. F2 litters were maintained through PND 63. Groups of Fl and F2 offspring
(n = 1-2 offspring per litter from 5-7 litters per group) were sacrificed on PND 22, 42, and 63. A
group of F2 offspring (n = 6-10 per group) also was sacrificed on PND 10. A lactational
challenge experiment was performed on PND 10 with Fl dams and F2 offspring. Mammary
glands were evaluated from FO dams on PND 22, from Fl dams on PNDs 10 and 22, and from
Fl and F2 female offspring on PNDs 10 (F2 only), 22, 42, and 63. Mammary gland whole
mounts were scored qualitatively as described above.
Exposure to 5 mg PFOA/kg/day significantly increased prenatal loss in FO mice and
significantly decreased the number of live offspring and the postnatal survival of the viable pups.
Maternal weight gain and number of implants did not differ among FO the groups. There was no
indication of toxicity in Fl adult females. Exposure to PFOA did not affect prenatal loss or
postnatal survival, although Fl females that had been exposed in utero to 5 mg/kg/day had
significantly fewer implants.
On PND 22, Fl pup body weight was similar across all treated and control groups. Fl
offspring body weight at PND 42 was significantly reduced for those whose dams received
5 mg/kg/day; at PND 63, body weight was significantly reduced for offspring from dams given
1 mg/kg/day plus 5 ppb in the drinking water compared to offspring from dams given
1 mg/kg/day. Liver-to-body weight ratios were significantly increased at 1 mg/kg/day on PND
22 and at 5 mg/kg on PNDs 22 and 42. For the F2 pups, a significant reduction in body weight
was observed in control plus 5 ppb drinking water PFOA offspring on PND 1, but there was no
difference by PND 3. F2 offspring from the 1 mg/kg/day and 1 mg/kg/day plus 5-ppb drinking
water PFOA groups had increased body weight compared to controls on PNDs 14, 17, and 22.
Liver-to-body weight ratios were no different across the groups.
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Mammary gland developmental scores for the three generations of females are summarized
in Table 3-24. At PND 22, control FO dams displayed weaning-induced mammary involution. At
PND 22, the mammary glands of all PFOA-exposed FO dams, including the control dams
receiving 5 ppb PFOA in drinking water, resembled glands of mice at or near the peak of
lactation (~PND 10). The Fl dams examined on PNDs 10 and 22 had significantly lower
developmental scores on PND 10, but that was no longer evident at PND 22, except for those
exposed in utero to 5 mg/kg/day.
Table 3-24. Mammary Gland Scores from Three Generations of CD-I Female Mice
Group
FO dams on PND 22
Fl as pups PND 63
Fl as dams on PND 10
Fl as dams on PND 22
F2 PND 10
F2 PND 22
F2 PND 42
F2 PND 63
Control
2.4 ±0.2
3.8 ±0.2
4.0 ±0.0
2.1 ±0.3
2.8 ±0.3
3.1±0.4
3.5 ±0.2
3.4 ±0.2
Control
+5 ppb
3.4±0.1*
2.6 ±0.4*
2.8 ±0.5*
2.2 ±0.2
3.0 ±0.2
1.9 ±0.3
2.5 ±0.4*
3. 5 ±0.2
1 mg/kg/day
3.0 ±0.2*
2.9 ±0.2*
2.5 ±0.2*
1.9 ±0.4
1.9 ±0.3
2.3 ±0.1
3.4 ±0.2
2.4 ±0.2*
1 mg/kg/day +
5 ppb
3.2 ±0.2*
2.0 ±0.3**
2.0 ±0.2*
1.5 ±0.2*
2.6 ±0.2
2.3 ±0.2
2.4±0.2*#
2.6 ±0.5
5 mg/kg/day
3.9±0.1*
2.2 ±0.2*
2.5 ±0.2*
3.2 ±0.3*
2.0 ±0.2
2.0 ±0.2
3.3 ±0.4
2.6 ±0.4
Notes: n = 4-11.
* p<0.05 compared with control.
* p<0.05 compared with 1 mg/kg/day.
Fl and F2 animals represented in each data set are different. They represent members of litters within each group at different
stages of development.
In the Fl female offspring not used for breeding, the mammary glands of all mice exposed to
PFOA were significantly delayed in development on PNDs 22, 42, and 63. For the F2 female
offspring, some differences in mammary gland scores were observed between the groups, but
most were not significantly different from controls.
In the lactational challenge experiment, dams were removed from their litters for 3 hours,
then returned to their litters and allowed to nurse for 30 mins. The time from the dam's return to
the litter and nursing initiation was recorded. The litters were weighed before and after nursing to
estimate volume of milk produced. The results from the lactational challenge on PND 10 for the
Fl dams showed a slight dose-related trend for decreased milk production (measured in grams)
over a 30-min period (differences from controls not identified as significant), but no clear
differences in time to initiate nursing (measured in seconds). As discussed above, morphological
differences were seen in developmental scores for the treated Fl dams on PND 10 and were
generally no longer evident at PND 22.
White et al. (2011) demonstrated that no significant dose-related differences were found in
the ability of the CD-I mice given 1 mg/kg/day to provide nourishment to their young as
reflected in measurements of body weight in Fl and F2 pups across a 63-day postnatal period.
There were body weight effects in the pups from dams given 5 mg/kg/day and in pups from
dams that received 1 mg/kg/day by gavage with 5 ppb in the drinking water.
In the study by Albrecht et al. (2013) discussed earlier, groups of female wild-type, PPARa-
null, and PPARa-humanized mice on a SV/129 genetic background were given 0 and 3 mg
PFOA/kg on GD 1-17 by oral gavage. Controls received the water vehicle. The study was
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designed with the goal of identifying the contribution of PPARa activation to the responses
evaluated. Mammary gland structure was one of the endpoints evaluated. Females were either
sacrificed on GD 18 (n = 5-8 per group) or allowed to give birth and then sacrificed, along with
their litters (n = 8-14), on PND 20. The left and right fourth and fifth mammary glands were
removed, spread on a glass slide, and stained. Ductal length and terminal end buds were
quantified in the offspring of from three to nine dams. There was no significant difference in the
measurements for either parameter at either timepoint for the offspring of PFOA-treated animals
compared to the controls. In the case of the wild-type mice, the terminal end bud measurements
were 2.1 ± 0.01 terminal end buds/gland for the control and 2.2 ± 0.2 terminal end buds/gland
based on the mean for three control litters and four PFOA-exposed litters. For the ductal lengths,
the values were 2.4 ± 0.3 millimeter (mm) for the control and 2.4 ± 0.4 mm for the PFOA-
exposed animals. There was no qualitative component of the scoring approach used by Albrecht
et al. (2013). The fewest number of terminal end buds and the longest ductal length measurement
were those for the animals with the hPPARa.
To examine the impact of differences in mouse strains, Tucker et al. (2015) conducted a
study of the effects of gestational exposure on mammary gland development as measured at
prepubertal time points. Doses of 0, 0.01, 0.1, 0.3, and 1 mg/kg/day were administered to timed
pregnant CD-I and C57B1/6 mice by gavage on GD 1-17. After parturition, the number of pups
was reduced so that there were ultimately four to eight CD-I litters per treatment block and three
to seven B57BL/6 litters per treatment. Endpoints monitored included body weight; net body
weight; absolute and relative liver weight on PNDs 21,35, and 56; neonatal developmental
endpoints (e.g., vaginal opening, first estrus); and serum estradiol and progesterone (P)
measurement; and as well as mammary gland development scores. Qualitative assessment of
mammary gland scores was as described above. Different treatment blocks monitored different
endpoints at different times. Serum POA levels were measured at PNDs 21, 35, and 56 for the
CD-I mice (n = 4-12) and at PND 21 and 61 for the C57BL/6 mice (n = 2-6). At each time
point, the serum concentration increased with dose and decreased with duration.
There were no measures that were significantly (p<0.05) different from controls for the CD-I
anthropometric parameters, except relative liver weight on PND 56 at 0.3 mg/g/day and on PND
21 at 1 mg/kg/day. Net body weight was significantly increased (p<0.05) at PNDs 21 and 35 in
the 1-mg/kg/day group. No significant differences were observed in the C57B1/6 mice at any
dose or duration. There were no significant differences for postnatal developmental endpoints,
estradiol, or P in either mouse strain. There was a trend towards decreasing mammary gland
developmental scores with dose for both strains of mice. In the CD-I mice, mammary gland
developmental scores were significantly reduced at >0.01 mg/kg/day on PND 35 and at
>0.1 mg/kg/day on PND 21 compared to scores in the controls. However, in the C57B1/6 mice,
mammary gland developmental scores were significantly reduced only at 0.3 and 1.0 mg/kg/day
on PND 21 compared to scores in the controls.
Serum P was higher in the control and treated CD-I mice on PND 56 than at the other two
time points but lacked dose response; estradiol was relatively consistent across time points. For
the C57BL/6 mice, the estradiol levels at PND 61 were higher in all treated groups but lacked
dose-response; P changed little with time and was similar between treated and control groups.
The LOAEL was 0.01 mg/kg/day for aberrant mammary gland development in the CD-I mice
and 0.3 mg/kg/day for the C57BL/6 mice. The CD-I mice lacked a NOAEL. The NOAEL for
the C57/BL/6 mice was 0.1 mg/kg/day. Although both strains experienced delayed prepubertal
mammary gland development, there were no significant changes in other postnatal
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developmental events. The relevance of the mammary gland changes at maturity in the absence
of any postlactational PFOA exposure is uncertain, especially as it relates to humans.
Direct peripubertal exposures
C. Yang et al. (2009) gavage-dosed 21-day-old female BALB/c mice (5 per group) with 0, 1,
5, and 10 mg PFOA/kg/day for 5 days per week for 4 weeks to determine the effects of
peripubertal PFOA exposure on puberty and mammary gland development. At necropsy, uteri
and livers were weighed and processed for histological examination. Mammary glands were
collected and processed for histological and whole-mount examination. A significant decrease in
body weight was observed following exposure to 10 mg/kg/day. The mammary glands of female
BALB/c mice treated with 5 or 10 mg/kg/day had reduced ductal length, decreased number of
terminal end buds, and decreased stimulated terminal ducts compared to the mammary glands of
control mice. BrdU incorporation into the mammary gland revealed a significantly lower number
of proliferating cells in the ducts and terminal end buds/terminal ducts at 5 mg/kg/day (not tested
at 10 mg/kg/day). Absolute and relative liver weight was significantly increased in all treated
BALB/c mice. The absolute and relative uterine weight was significantly decreased in all treated
mice compared to uterine weight in control mice. Vaginal opening was significantly delayed in
mice dosed with 1 mg/kg/day and did not occur at 5 or 10 mg/kg/day. The LOAEL was 1
mg/kg/day based on delayed vaginal opening, increased liver weight, and decreased uterine
weight; and no NOAEL was established.
C. Yang et al. (2009) also dosed 21-day-old female C57BL/6 mice in the same manner as the
BALB/c mice and examined the effects of PFOA on mammary gland development and vaginal
opening. The body weight effects were similar in both strains with 10 mg/kg/day causing
significantly reduced body weight. At 5 mg/kg/day, PFOA had a stimulatory effect on the
mammary glands. There was a significant increase in the number of terminal end buds and
stimulated terminal ducts. Ductal length was not affected. Mammary gland development was
inhibited in mice dosed with 10 mg/kg/day, with no terminal end buds or stimulated terminal
ducts present and very little ductal growth. Absolute and relative liver weight was significantly
increased in all treated mice. The absolute and relative uterine weight was significantly increased
in C57BL/6 mice dosed with 1 mg/kg/day and significantly decreased in C57BL/6 mice dosed
with 10 mg/kg/day. There was no difference in uterine weights between mice treated with
5 mg/kg/day and control mice. Vaginal opening was delayed in C57BL/6 mice dosed with
5 mg/kg/day and did not occur in mice dosed with 10 mg/kg/day. The LOAEL was 1 mg/kg/day
based on increased liver and uterine weights, and no NOAEL was established.
Y. Zhao et al. (2010) conducted several experiments in C57BL/6 mice to determine the
potential mechanism by which peripubertal PFOA exposure resulted in the stimulation of
mammary gland development observed by C. Yang et al. (2009). In experiments to determine if
PFOA has a hormonal effect on mammary gland development, C57BL/6 mice (n = 10 per group)
were OVX at 3 weeks of age, allowed 1 week to recover, and treated with 0 and 5 mg PFOA/kg
bw/day for 4 weeks. Abdominal and inguinal mammary glands were collected at sacrifice,
prepared as whole mounts, and scored for growth and development. The mammary glands of the
OVX control and PFOA-treated OVX mice were similarly stunted in growth as evidenced by no
outgrowth of ducts or presence of terminal end buds. This was in contrast to the stimulatory
effect of PFOA observed by C. Yang et al. (2009) in intact mice.
In experiments to determine if PFOA-affected mammary glands respond to hormone
treatment, intact C57BL/6 mice were dosed with 0 or 5 mg/kg bw/day of PFOA for 4 weeks
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starting at 21 days of age. After the last dose, the mice were OVX, allowed to recover for 1
week, and injected subcutaneous for 5 days with E2 (1 ng/0.2 ml per mouse), P (1 mg/0.2 ml per
mouse), or both (E+P, 1 |ig+l mg/0.2 ml per mouse). The mice were sacrificed 24 hours after the
last hormone injection. Abdominal and inguinal mammary glands were collected at sacrifice,
prepared as whole mounts, and scored for growth and development. In the mammary glands of
mice treated with PFOA and estradiol, stimulated terminal ducts were observed, and in 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 P.
In experiments to determine if PFOA-induced mammary gland development stimulation is
related to PPARa expression and the impact of PFOA on steroid hormones and growth factors,
female C57BL/6 and PPARa-null C57BL/6 mice (n = 5-10 mice per group) were gavage-dosed
with 0 or 5 mg/kg bw/day of PFOA 5 days per weeks for 4 weeks starting at 21 days of age (Y.
Zhao et al. 2010). Vaginal opening was monitored daily and estrous cycle state was determined
at sacrifice after 4 weeks of treatment. At necropsy, blood was collected for measurement of
serum steroid hormones and binding proteins. Portions of the mammary glands, ovaries, and
livers were collected and processed for histological examination. RNA was extracted from the
livers for quantitative RT-PCR and PCR array for selected genes related to metabolism of drugs,
toxic chemicals, hormones, and micronutrients. Portions of mammary glands were used in
western blot analysis of several enzymes, local growth factors, and receptors, including
aromatasewhich aids in converting testosterone to estradiol and androstenedione to estrone,
hydroxysteroid 1?P dehydrogenase 1 (HSD17P1)which aids in converting estrone to estradiol,
and hydroxysteroid 3p dehydrogenase 1 (HSD3P1)which aids in converting pregnenolone to P
and androstenedione to testosterone. Growth factors critically involved in mammary gland
development, including amphiregulin (Areg), insulin like growth factor I (IGF-I), and hepatocyte
growth factor (HGFa), and markers of cell proliferation (e.g., cyclin Dl and PCNA) were
analyzed by western blot. Areg mediates estrogen receptor a (ERa) function and is a ligand for
the epidermal growth factor receptor (EGFR). These receptors also were analyzed by western
blot.
The mammary glands of PPARa-null mice treated with PFOA had an increased number of
terminal end buds and stimulated terminal ducts compared to control PPARa-null mice. Protein
levels of Areg, IGF-I, HGFa, ERa, and EGFR were significantly increased (p<0.05) in PFOA-
treated C57BL/6 mice; and Areg, HGFa, ERa, and EGFR were significantly increased (p<0.05)
in PFOA-treated PPARa-null mice. Cyclin Dl and PCNA were significantly increased (p<0.05)
in C57BL/6 and PPARa-null mice treated with PFOA compared to levels in control mice.
Immunofluorescent staining of the mammary glands for ERa and Areg showed a significant
increase (p<0.05) in Areg positive luminal epithelial cells and Areg and ERa double positive
staining cells in C57BL/6 and PPARa-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 PPARa expression and suggest that PFOA increases the levels of steroid
hormones, growth factors, and receptors, which promote mammary gland cell proliferation.
Estradiol levels were similar between intact control and treated wild-type mice, but P levels
were significantly increased (p<0.05) in PFOA-treated mice in proestrus and estrus compared to
control mice in the same stages of the estrous cycle. Serum SHBG and albumin levels were not
significantly changed by treatment with PFOA.
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The effect of PFOA on aromatase, HSD17P1, and HSD3P1 activity in the ovaries of
C57BL/6 PPARa-null mice was examined. In C57BL/6 mice, HSD17P1 and HSD3P1 proteins
were significantly increased (p<0.05), and in PPARa-null mice, HSD17P1 protein was
significantly increased. Aromatase levels were not affected by PFOA. The results suggest that
PFOA might increase serum steroid hormone levels in the ovaries.
Due to the increased P levels observed in PFOA-treated mice, the expression of liver
metabolic enzymes was analyzed. Liver metabolic function might affect steroid hormone serum
levels, which play a role in mammary gland development. In PPARa-null and C57BL/6 mice
treated with PFOA, detoxification enzymes in the liver, including glutathione s-transferase al,
u3, and u4, were upregulated (p<0.05). Expression of liver drug metabolic enzymes, including
CYPlal, CYPla2, and HSD17P2, was significantly downregulated (p<0.05) in C57BL/6 mice
treated with PFOA, but expression in PFOA-treated PPARa-null mice was comparable to that in
control mice. Hydroxysteroid 1?P dehydrogenase 4, an enzyme that converts estradiol to estrone,
was significantly upregulated (p<0.05) in C57BL/6 mice treated with PFOA. The results
suggested that PFOA-induced expression changes in liver enzymes might not contribute to
PFOA-induced mammary gland development stimulation.
Inhalation Exposure
Staples et al. (1984) exposed Sprague-Dawley rats to PFOA using whole-body dust
inhalation for 6 hours per day on GD 6-15. The MMAD of the particles ranged from 1.4 to
3.4 |im and the GSD ranged from 4.3 to 6.0. The study was carried out in two trials with each
trial including two experiments. In experiment 1, the dams were sacrificed on GD 21 prior to
parturition, and in experiment 2, the dams were allowed to litter and were sacrificed on PND 23;
offspring were sacrificed on PND 35. In the trial 1 (both experiments), dams (n = 12) were
exposed to 0, 0.1, 1, and 25 mg/m3. In trial 2, the high dose was reduced to 10 mg/m3. In
experiment 1 of trial 2, dams numbered 12-15 per group and two additional groups (6 dams per
group) were added and were pair-fed at 10 and 25 mg/m3. In experiment 2 of trial 2, only six
control and six dams dosed at 10 mg/m3 were allowed to litter.
In experiment 1, the dams were weighed on GDs 1, 6, 9, 13, 16, and 21 and observed daily
for abnormal clinical signs. On GD 21, the dams were sacrificed by cervical dislocation and
examined for any gross abnormalities, liver weights were recorded, and the reproductive status
of each animal was evaluated. The ovaries, uterus, and contents were examined for the number
of corpora lutea, live and dead fetuses, resorptions, and implantation sites. Pups (live and dead)
were counted, weighed and sexed, and examined for external, visceral, and skeletal alterations.
The heads of all control and high-dosed group fetuses were examined for visceral alterations and
macroscopic and microscopic evaluations were conducted of the eyes.
Treatment-related clinical signs of maternal toxicity occurred at 10 and 25 mg/m3 and
consisted of wet abdomens, chromodacryorrhea, chromorhinorrhea, a general unkempt
appearance, and lethargy in four dams at the end of the exposure period (high-concentration
group only). Three out of 12 dams died during treatment at 25 mg/m3 (on GDs 12, 13, and 17).
Food consumption was significantly reduced at 10 and 25 mg/m3; however, no significant
differences were noted between treated and pair-fed groups. Significant reductions in body
weight also were observed at these concentrations, with statistical significance at the high
concentration only. Likewise, statistically significant increases in mean liver weights (p<0.05)
were seen in the high-concentration group. The NOAEL and LOAEL for maternal toxicity were
1 and 10 mg/m3, respectively.
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No effects were observed on the maintenance of pregnancy or the incidence of resorptions.
Mean fetal body weights were significantly decreased in the 25 mg/m3 PFOA group (p = 0.002)
and in the pair-fed control group (p = 0.001). Interpretation of the decreased fetal body weight is
difficult given the high incidence of mortality in the dams. The NOAEL and LOAEL for
developmental toxicity were 10 and 25 mg/m3, respectively.
In experiment 2, in which the dams were allowed to litter, the procedure was the same as for
experiment 1 until GD 21. Two days before the expected day of parturition, each dam was
housed in an individual cage. The date of parturition was noted and designated PND 1. Dams
were weighed and examined for clinical signs on PNDs 1, 7, 14, and 22. On PND 23 all dams
were sacrificed. Pups were counted, weighed, and examined for external alterations. At birth,
each pup was subsequently weighed and then inspected for adverse clinical signs on PNDs 4, 7,
14, and 22. The eyes of the pups were also examined on PNDs 15 and 17. Pups were sacrificed
on PND 35 and examined for visceral and skeletal alterations.
Clinical signs of maternal toxicity seen at 10 and 25 mg/m3 were similar in type and
incidence to those described for trial 1. Maternal body weight gain during treatment at 25 mg/m3
was less than controls, although the difference was not statistically significant. In addition, two
out of 12 dams died during treatment at 25 mg/m3. No other treatment-related effects were
reported, nor were any adverse effects noted for any of the measurements of reproductive
performance. The NOAEL and LOAEL for maternal toxicity were 1 and 10 mg/m3, respectively.
Signs of developmental toxicity in this group consisted of statistically significant reductions
in pup body weight on PND 1 (6.1 g at 25 mg/m3 versus 6.8 g in controls, p = 0.02). On PNDs 4
and 22, pup body weight continued to remain lower than controls, although the difference was
not statistically significant. No significant effects were reported following external examination
of the pups or with ophthalmoscopic examination of the eyes. The NOAEL and LOAEL for
developmental toxicity were 10 and 25 mg/m3, respectively.
Dermal Exposure
No data on the developmental effects of dermal exposures to PFOA were identified in the
literature.
3.2.8 Chronic Toxicity
Oral Exposure
Monkey. Male cynomolgus monkeys (n = 4 or 6 per dose) were administered PFOA by oral
capsule containing 0, 3, 10, or 30/20 mg/kg/day for 26 weeks (Butenhoff et al. 2002). Dosing of
animals in the 30-mg/kg/day dose group ceased after 12 days and decreased to 20 mg/kg/day
when reinstated on day 22 because of low food consumption, decreased body weight, and
decreased feces. Sacrifice of the surviving monkeys, except for two control monkeys and two
monkeys from the mid-dose group (recovery animals) occurred at 26 weeks. The animals in the
recovery groups were sacrificed 13 weeks later.
Animals were observed twice daily for mortality and moribundity and were examined at least
once daily for signs of poor health or abnormal behavior. Ophthalmic examinations were
performed before treatment began and at weeks 26 and 40. Body weight, food consumption,
clinical hematology, clinical chemistry, urinalysis, serum hormone levels, and PFOA levels in
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blood and tissue were assessed throughout the study. One animal from the 30/20-mg/kg/day dose
group was sacrificed in moribund condition on day 29 with signs of dosing injury and liver
lesions. One animal from the 3-mg/kg/day dose was sacrificed (day 137) with signs of hind limb
paralysis, ataxia and hypoactive behavior, few feces, and no food consumption. Treatment of the
remaining three animals given 30/20 mg/kg/day was halted on days 43, 66, and 81, respectively,
because of thin appearance, few or no feces, low or no food consumption, and weight loss, but
the animals appeared to recover from compound-related effects within 3 weeks after cessation
of treatment. No significant changes in mean body weight were observed at doses of 3 or
10 mg/kg/day.
Serum hormone levels (i.e., estrone, estradiol, estriol, testosterone, TSH, FT4, total T4, and
CCK) were not significantly altered throughout the study. However, FT3 and total T3 levels
were significantly decreased (p<0.05) from weeks 5 to 10 and at week 27 in the 30/20-mg/kg/day
dose group compared to controls.
At terminal sacrifice (26 weeks), mean absolute liver weight was significantly increased in
all dose groups and the relative liver-to-body weight ratio was significantly increased for the
High-Dose Group. Final Body Weight And Liver Weight Data Are Presented In Table 3-25.
Table 3-25. Liver Weight Data in Monkeys Administered PFOA for 6 Months
Dose
0 mg/kg (n = 4)
3 mg/kg (n = 3)
10 mg/kg (n = 4)
30/20 mg/kg (n = 2)
Body Weight
3947 ±591
4486 ± 30
4447 ± 498
3925 ±583
Absolute Liver Wt (g)
60.2 ±6.9
81.8 ±2.8*
83.2 ±9.7*
90.4 ±4.2*
Relative Liver Wt (%)
1.5±0.1
1.8±0.1
1.9±0.1
2.4 ±0.5*
Source: Butenhoff et al. 2002
Note: * Significantly different from control, p<0.01.
The cause of the increase in liver weight was suggested to be hepatocellular hypertrophy
(indicated by decreased hepatic DNA content), which was hypothesized to result from
mitochondrial proliferation based on an increase in hepatic succinate dehydrogenase activity.
The two animals given 20 mg/kg/day had significantly decreased hepatic DNA content, and
increased succinate dehydrogenase and palmitoyl-CoA oxidase activities; glucose-6-phophatase
activity was slightly decreased in all treated groups, but a dose-response was not shown. These
data are shown in Table 3-26. Succinate dehydrogenase activity was highly variable in animals
given 3 mg/kg/day despite this group having the most consistent liver PFOA concentrations.
Although serum steady-state had been attained by 4-6 weeks of dosing (Table 3-26 (see section
2.2, Distribution), liver PFOA levels ranged from 11.3-18.5, 6.29-21.9, and 16-83.3 |ig/g tissue
in the 3, 10, and 20 mg/kg/day groups, respectively.
Because administration of PFOA to rats has been shown to result in liver, Ley dig cell tumors
(LCTs), and pancreatic acinar cell tumors (PACTs), Butenhoff et al. (2002) analyzed markers of
tumor formation in the monkey study just described. In the liver, a twofold increase in hepatic
palmitoyl-CoA oxidase activity was observed in the 30/20-mg/kg/day group, consistent with
reports for species that are not particularly responsive to PPARa agonists. Replicative DNA
synthesis in the liver, an indication of cell proliferation, was not altered in the treated animals. It
also has been proposed that changes associated with the PACTs in rats include increased serum
CCK concentrations and indications of cholestasis, including increases in ALP, bilirubin, and
bile acids. None of these changes were observed in the cynomolgus monkeys. There were also
no significant changes in estradiol, estriol, or testosterone in the monkeys. Each of these factors
Perfluorooctanoic acid (PFOA) - May 2016 3-103
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is associated with LCTs in rats. There were no changes in replicative DNA synthesis in the
pancreas or testes.
Table 3-26. Subcellular Liver Enzyme Activities and Liver PFOA Concentrations
Endpoint
DNA (mg/g liver)
Succinate dehydrogenase (umol
cytochrome c reduced/min/g liver)
Palmitoyl-CoA oxidation
(umol/min/g liver)
Acid phosphatase (umol/min/g
liver)
Glucose-6-phophatase (umol/min/g
liver)
PFOA liver level (ug/g tissue)
(individual animal)
0 mg/kg/day
1.44 ±0.28
0.21 ±0.15
0.53 ±0.12
0.78 ±0.10
12.32±3.11
0.09
-------
Table 3-27. Clinical Chemistry Values from Male Rats Given PFOA for 2 Years
Endpoint
ALT (IU/L)
AST (IU/L)
ALP (mg/dL)
Diet Level
(ppm)
0
30
300
0
30
300
0
30
300
3 Months
21.4 ±2.67
34.5 ±15.33*
31. 9 ±21.94*
45.3 ± 7.26
59.7 ±22.47
58.2 ±27.23
91.1 ±26.22
138.7 ±33. 14*
153.5 ±31. 84*
6 Months
24.1 ±3.75
53. 3 ±29.34*
54.8 ±29.26*
49.7 ±14.98
92.1 ±45.6*
87.8 ±34.83*
97.1 ±40.41
146.9 ±37.13*
147.3 ± 34.85*
12 Months
33.5 ±19.45
77.6 ±56.59*
106.1 ±70*
79.1 ±44.61
124.4 ± 94.04*
132.7 ±76.84*
150.8 ±43.94
128.3 ±41.75
166.5 ±59.28*
18 Months
34.1 ± 10.68
59.7 ±33.41*
84.3 ± 55.95*
99.1 ±68.14
116.4 ±57.99
123.3 ±62.98
85.2 ±33.76
112.5 ±32.61
184.4 ±73.37*
24 Months
33.4±8.1
42.5 ±10
61. 8 ±20.13*
64.9 ±25.76
68.0 ±17.64
95.7 ±29.76*
70.1 ±25.53
81.2 ±26.2
113.5 ±22.84*
Source: Butenhoff et al. 2012
Note:
* Significantly different from control, p < 0.05.
Incidence of selected microscopic lesions is detailed in Table 3-28; severity scores were not
given for any type of lesion. Significantly increased incidence of lesions in the liver was
observed in the high-dose male group. At 1 year, diffuse hepatocellular hypertrophy, portal
mononuclear cell infiltration, and hepatocellular necrosis were seen. At 2 years, significant
increases in hepatocellular hypertrophy were observed in the males and females in the high-dose
group. Hepatic cystoid degeneration, a condition characterized by areas of multilocular
microcysts in the liver parenchyma, also was significantly increased in high-dose males. The
incidence of hepatocellular necrosis did not increase for the high-dose males at the end of the
study compared with the interim rate.
Among the high-dose males, histological changes were noted in tissues other than the liver.
Small but statistically significant increases in vascular mineralization of the testes and of
pulmonary hemorrhage probably were not caused by treatment with PFOA. In the lung, while the
incidence of alveolar macrophages was increased, that of perivascular mononuclear infiltrate and
of pneumonia were decreased and vascular mineralization was a common finding in treated and
control animals.
The LOAEL for male rats is 300 ppm (14.2 mg/kg/day) based on a decrease in body weight
gain and histological changes in the liver. The LOAEL for female rats is 300 ppm
(16.1 mg/kg/day) based on decreased body weight gain. The NOAEL for both genders is 30 ppm
(1.3 mg/kg/day for males and 1.6 mg/kg/kg for females).
Biegel et al. (2001) conducted a 2-year mechanistic study in which male Crl:CD BR (CD)
rats (n = 156 per group) were fed a diet containing 0 or 300 ppm PFOA (0 or 13.6 mg/kg/day).
Interim sacrifices were conducted every 3 months up to 21 months for measurements of liver and
testes weights, peroxisome proliferation, and cell replication. Serum samples were collected and
reproductive hormones measured.
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Table 3-28. Incidence of Nonneoplastic Lesions in Rats Given PFOA for 2 Years
Lesion
0 ppm
30 ppm
300 ppm
Males
Liver
Cystoid degeneration
Hepatocellular hypertrophy
[incidence at 1 year]
Mononuclear cell infiltrate
[incidence at 1 year]
Necrosis
[incidence at 1 year]
Lung
Alveolar macrophages
Hemorrhage
Mononuclear infiltrate
Testes
Vascular mineralization
4/50
0/50
[0/15]
37/50
[7/15]
3/50
[0/15]
10/49
10/49
21/49
0/49
7/50
6/50
32/50
5/50
[-1
16/50
14/49
3/49*
3/50
28/50*
40/50*
[12/15]
48/50*
[13/15]
5/50
[6/15]
31/49*
22/50*
7/50*
9/50*
Females
Liver
Cystoid degeneration
Hepatocellular hypertrophy
Mononuclear cell infiltrate
Necrosis
0/50
0/50
19/50
5/50
0/50
1/50
11/50
6/50
0/50
8/50*
19/50
2/50
Source: Butenhoff et al. 2012
Notes:
* Significantly different from control, p<0.05.
- Not examined; interim sacrifice not done on animals at 30 ppm.
Body weight was significantly decreased from days 8 through 630 in PFOA-exposed rats. In
the treated group, relative liver weights and hepatic p-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
Ley dig cell proliferation was observed at any sampling times. The incidence of Ley dig cell
hyperplasia was significantly increased in PFOA-exposed rats (46% versus 14% in the control
group). Pancreatic acinar cell proliferation was significantly increased at 15, 18, and 21 months.
The incidence of acinar cell hyperplasia was 30/76 (39%) compared to the incidence in the
control group of 14/80 (18%). There were no significant differences in serum testosterone or
prolactin in the PFOA-treated rats when compared to the controls. Serum FSH was significantly
increased at 6 months, and LH was significantly increased at 6 and 18 months. There were
significant increases in serum estradiol concentrations in the treated rats at 1,3, 6, 9, and 12
months.
3.2.9 Carcinogenicity
Oral Exposure
Rat. Tissues from the animals in the Butenhoff et al. study (2012) were evaluated for neoplastic
and preneoplastic formations; this study was conducted from April 1981 through May 1983.
Hepatocellular carcinomas were observed at 6% (3/49), 2% (1/50), and 10% (5/50) in the
control, low-, and high-dose male rats, respectively. None were observed in females in the
control and low-dose groups, but a 2% (1/50) incidence was observed for female rats in the high-
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dose group. The differences between the treated and control groups were not significantly
different. No liver adenomas were observed.
At the 1-year sacrifice, testicular masses were found in 7/50 (14%) high-dose and 2/50 (4%)
low-dose rats, but not in any of the controls. A significant increase (p<0.05) in the incidence of
testicular (Ley dig) cell adenomas was observed in the high-dose male rats at the end of the study.
The LCT incidence in the control, low-, and high-dose groups was 0/50 (0%), 2/50 (4%), and
7/50 (14%), respectively. The increase also was statistically significant when compared to the
historical control incidence of 0.82% observed in 1,340 Sprague-Dawley control male rats used
in 17 carcinogenicity studies (Chandra et al. 1992). In a published workshop report on LCTs,
Clegg et al. (1997) identified the spontaneous incidence of LCTs in 2-year-old Sprague-Dawley
rats as approximately 5%.
A statistically significant, dose-related increase in the incidence of ovarian tubular
hyperplasia was found in female rats at the 2-year sacrifice. The incidence of this lesion in the
control, low-, and high-dose groups was 0%, 14%, and 32%, respectively. The biological
significance of this effect at the time of the initial evaluation was unknown, as there was no
evidence of progression to tumors.
Slides of the ovaries from the Butenhoff et al. study (2012)originally conducted from April
1981 through May 1983were reevaluated by Mann and Frame (2004) with emphasis placed on
the proliferative lesions of the ovary. Using more recently published nomenclature, the ovarian
lesions were diagnosed and graded as gonadal stromal hyperplasia and/or adenomas, which
corresponded to the diagnoses of tubular hyperplasia or tubular adenoma by the original study
pathologist. The data are summarized in Table 3-29. No statistically significant increases in
hyperplasia (total number), adenomas, or hyperplasia/adenoma combined were seen in treated
groups compared to controls. There was some evidence of an increase in size of stromal lesions
observed at the 300-ppm group; however, adenomas occurred in greater incidence in the control
group than in either of the treated groups. Results of this follow-up evaluation indicated that rats
sacrificed at the 1-year interim sacrifice, as well as rats that died prior to the interim sacrifice,
were not considered at risk for tumor development.
Table 3-29. Incidence of Ovarian Stromal Hyperplasia and Adenoma in Rats
Group
No. examined
Hyperplasia (Total)
Grade 1
Grade 2
Grade 3
Grade 4
Adenoma
Adenoma and/or Hyperplasia
0 ppm
45
8
6
2
0
0
4
12
30 ppm
47
16
7
3
5
1
0
16
300 ppm
46
15
5
1
6
3
2
17
Source: Mann and Frame 2004
Mammary gland tumors also were observed in the Butenhoff et al. (2012) bioassay. In the
original analysis of mammary tissues from female rats, the incidence of fibroadenoma of the
mammary gland in the female 300-ppm group (48%) was greater than that in either of the
concurrent control groups (22%). It also was similar to the incidence in the 30-ppm group (42%),
but considered to be within the norm for background variation of this lesion in Sprague-Dawley
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rats based on the published literature. As a result of questions raised about this conclusion, a
pathology working group (PWG) was commissioned to review the female mammary tumor
findings, blinded to treatment status, using current diagnostic criteria (Hardisty et al. 2010).
Table 3-30 compares the original mammary gland tumor findings to those of the PWG.
Table 3-30. Mammary Gland Tumor Incidence Comparison
Number necropsied
Lobular hyperplasia (%)
Adenocarcinoma (%)
Fibroadenoma3 (%)
Adenoma (%)
0 ppm
Butenhoff
50
6
(12%)
8
(16%)
10
(20%)
3
(6%)
Hardisty
50
0
(0%)
9
(18%)
18
(36%)
1
(2%)
30 ppm
Butenhoff
50
3
(6%)
14
(28%)
19
(38%)
0
(0%)
Hardisty
50
2
(4%)
16
(32%)
22
(44%)
0
(0%)
300 ppm
Butenhoff
50
2
(4%)
5
(10%)
21
(42%)
0
(0%)
Hardisty
50
0
(0%)
5
(10%)
23
(46%)
0
(0%)
Source: Hardisty 2005; Hardisty et al. 2010
Notes:
a Includes fibroadenoma, multiple counts.
The principal differences between the original reported findings and the PWG results relate
to changes in the mammary gland that were initially reported as lobular hyperplasia, which the
PWG felt had features more characteristic of mammary gland fibroadenoma (Table 3-30). As a
result, the numbers of rats with benign tumors (adenoma and fibroadenoma) were reclassified
from 13 to 19 in the control group, from 19 to 22 in the 30-ppm group, and from 21 to 23 in the
300-ppm group. Although the incidence of neoplasms varied among the control and treated
groups, there were no statistically significant differences when evaluated using the Fisher's exact
test for pairwise comparison for fibroadenoma, adenocarcinoma, total benign neoplasms, and
total malignant neoplasms. The morphologic appearance, overall incidence, and distribution of
the neoplasms observed in treated and control groups were similar, resulting in a conclusion that
they are not related to compound administration.
A 2-year mechanistic study in male Crl:CD BR (CD) rats (Biegel et al. 2001; Cook et al.
1992) resulted in liver tumors, LCTs, and PACTs. The rats (n = 156 per group) were fed diets
containing 0 ppm (ad libitum control and control pair-fed to the PFOA-exposed rats) or 300-ppm
PFOA (13.6 mg/kg intake). Rats were euthanized at interim time points of 1, 3, 6, 9, 12, 15, 18,
and 21 months. All rats surviving the 24-month test period were necropsied for microscopic
examination of various organs (e.g., kidneys, liver, testes, brain, heart, spleen). The incidence of
liver adenomas in the ad libitum control, pair-fed control, and treated groups was 3% (2/80), 1%
(1/79), and 13% (10/76), respectively. In the Butenhoff et al. study (2012), no hepatic adenomas
were observed. The incidence for liver carcinomas was 0% (0/80), 3% (2/79), and 0% (0/76) in
the ad libitum control, pair-fed control, and treated groups, respectively.
There was a significant increase in the incidence of Ley dig cell adenomas in the treated
rats11% (8/76) when compared to the pair-fed control rats (3%, 2/78)supporting the
observations from the Butenhoff et al. study (2012). The incidence in ad libitum control rats was
0% (0/80). In addition, the treated group had a significant increase in the incidence of liver
adenomas and pancreatic acinar cell adenomas when compared to the pair-fed and ad libitum
control groups. The incidence for the pancreatic acinar cell adenomas was 0% (0/80) in the
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treated rats, 1% (1/79) in the pair-fed control rats, and 9% (7/76) in the control rats. The
incidence of pancreatic acinar cell carcinoma was 1% (1/76) in the treated rats, 0% (0/79) in the
pair-fed control rats, and 0% (0/80) in the control rats.
In Butenhoff et al. (2012), there was no reported increase in the incidence of PACTs.
However, the incidence of pancreatic acinar hyperplasia in the male rats was 0/33, 2/34, and 1/43
in the control, 30-, and 300-ppm groups, respectively. To resolve this discrepancy, the
histological slides from both studies were reviewed by independent pathologists. This review of
the microscopic lesions of the pancreas in the two studies indicated that PFOA produced
increased incidence of proliferative acinar cell lesions of the pancreas in the rats of both studies
at the dietary concentration of 300 ppm. The differences observed were quantitative rather than
qualitative; more and larger focal proliferative acinar cell lesions and greater tendency for
progression of lesions to adenoma of the pancreas were observed in the Biegel et al. study (2001)
than in the Butenhoff et al. study (2012). The difference between pancreatic acinar hyperplasia
(Butenhoff et al. 2012) and adenomas (Biegel et al. 2001) in the rat was a reflection of arbitrary
diagnostic criteria and nomenclature by different pathologists. The basis for the quantitative
difference in the lesions observed is not known, but was believed most likely to have been
caused by the difference in the diets used in the two laboratories (Frame and McConnell 2003).
Mouse. Filgo et al. (2015) reported on tumor development in females from three strains of mice
(CD-I, SV-139, and SV-129 PPARa knock-out [KO]) at 18 months with exposures that occurred
only during development (gestation and lactation). The animals were from separate experiments
initially carried out by EPA and published as Hines et al. (2009) and Abbott et al. (2007). The
Filgo et al. (2015) analysis focused on the mature offspring from the earlier publications and was
carried out at the National Institute for Environmental Health Sciences (NIEHS). Dosing
regimens differed for the individual strains as did the doses and the number of animals per dose
group. Some of the animals in the original studies had died before the 18-month sacrifice at
NIEHS. After sacrifice, the livers were recovered for analysis. The tissue sections were reviewed
by a team of board-certified veterinary pathologists. Table 3-31 summarizes the tumor results.
Table 3-31. Liver Tumors in Three Strains of Mice at 18 Months with Exposure to PFOA
Only during Gestation and Lactation
Strain
CD-I
SV-129
SV-129-
PPARoKO
0
mg/kg/day
0.01
mg/kg/day
0.1
mg/kg/day
0.3
mg/kg/day
0.6
mg/kg/day
1
mg/kg/day
3
mg/kg/day
5
mg/kg/day
Number of Tumors / Total Number Tested
Tumor Type
1/29 L
0/10
0/6
1/29 HCA
NT
NT
1/37 HCA
1/10
HcyS
1/10
HCA
6/26
HCA(4),
HCC,L
0/8
2/10
HCA. ICT
NT
0/6
NT
2/31
HcyS, L
0/8
1/9
ICT
NT
NT
2/9
HCA
6/21
HmS(2),
HCA,
HCC,
HcyS, L
NT
NT
Notes:
HCA = hepatocellular adenoma, HCC = hepatocellular carcinoma, HcyS = histocytic sarcoma, HmS = hemangeosarcomas, ICT
= Ito cell tumor, L = lymphoma, NT = not tested.
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It is difficult to draw conclusions regarding the carcinogenicity of PFOA in mice based on
the data collected because of the small number of animals evaluated in both studies of SV-129
mice and the lack of PFOA exposure between PND 21 and 18 months for all dose groups. As
was the case for liver tumors in the Butenhoff et al. study (2012), there is a lack of dose-response
for total liver tumors, although the four hepatocellular adenomas seen at 0.3 mg/kg/day in CD-I
mice were significantly greater (p<0.05) than the control. Tumor types varied across the dose
groups. The authors also reported on preneoplastic basophilic, and eosinophilic foci were
observed in the CD-I mice but did not show a response to dose.
An interesting histological finding in both the CD-I and SV-129 mice was a trend for
increased Ito cell atrophy and lesion severity across the doses (Filgo et al. 2015). Since Ito cells
accumulate fat in the liver sinusoids, this observation provides additional support for hepatic
steatosis as a condition of concern following developmental PFOA exposure. There was an
increase in severity with dose for the Ito cell fat deposits for all but the high-dose group. The Ito
cell lesion was present in the SV-129 mice, but was not associated with tumors. CD-I mice had a
significant increase in Ito cell hypertrophy at 5 mg/kg/day compared to controls, but there was a
lack of dose-response. The authors concluded that liver damage from PFOA exposure occurring
early in development is not totally linked to PPAR-a and could progress as animals aged without
continued dosing, thus compromising liver function and possibly leading to tumor development.
Inhalation and Dermal Exposures
No data on the tumorigenic effects of chronic inhalation or dermal exposures to PFOA were
identified in the literature.
3.3 Other Key Data
3.3.1 Mutagenicity and Genotoxicity
PFOA has been tested for genotoxicity in a variety of in vivo and in vitro assays. The data
from the in vitro studies are summarized in Table 3-32.
PFOA was tested in a cell transformation and cytotoxicity assay conducted in
mouse embryo fibroblasts. The cell transformation was determined as both colony
transformation and foci transformation. There was no evidence of transformation at any of the
dose levels tested in either the colony or foci assay methods (Garry and Nelson 1981).
PFOA was tested twice (Lawlor 1995, 1996) for its ability to induce mutation in the
Salmonella - E. co//'/mammalian-microsome reverse mutation assay. The tests were performed
both with and without metabolic activation. A single positive response seen in S. typhimurium
TA1537 when tested without metabolic activation was not reproducible. PFOA did not induce
mutation in either S. typhimurium or E. coli when tested either with or without metabolic
activation. PFOA did not induce chromosomal aberrations in human lymphocytes when tested
with and without metabolic activation up to cytotoxic concentrations (Murli 1996a; NOTOX
2000). Sadhu (2002) reported that PFOA did not induce gene mutation when tested with or
without metabolic activation in the K-l line of CHO cells in culture.
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Table 3-32. Genotoxicity of PFOA In Vitro
Test System
C3H10T./2 mouse
embryo fibroblasts
C3H 10T./2 mouse
embryo fibroblasts
S. typhimurium
TA1537
E. coli
CHO cells
CHO cells
Human lymphocytes
K-l CHO cells
S. typhimurium
TA98, TA100,
TA102, TA104
End-point
Cell Transformation
Cytotoxicity
Gene Mutation
Gene Mutation
Chromosomal
Aberrations
Polyploidy
Chromosomal
Aberrations
Gene Mutation
Gene Mutation
With Activation
NA
NA
-
-
+, +
+, +
-
-
-
Without Activation
-
-
+
(not reproducible)
-
+, -
+, -
-
-
-
Reference
Garry and Nelson
1981
Garry and Nelson
1981
Lawlor 1995, 1996
Lawlor 1995, 1996
Murli 1996b, 1996c
Murli 1996b, 1996c
Murli 1996c;
NOTOX 2000
Sadhu 2002
Freire et al. 2008
Note: NA= not applicable.
Murli (1996b, 1996c) tested PFOA twice for its ability to induce chromosomal aberrations in
CHO cells. In the first assay, PFOA induced both chromosomal aberrations and polyploidy in
both the presence and absence of metabolic activation. In the second assay, no significant
increases in chromosomal aberrations were observed without activation. However, when tested
with metabolic activation, PFOA induced significant increases in chromosomal aberrations and
in polyploidy (Murli 1996b). The effects were observed only at toxic concentrations (EFSA
2008).
PFOA did not display mutagenic activity with or without metabolic activation in
S. typhimurium strains TA98, TA100, TA102, or TA104 (Freire et al. 2008).
The data summarized in Table 3-32 suggest that PFOA is not a mutagen. A single positive
result in S. typhimurium was not reproducible by the same authors and was not replicated in
other studies. Potential chromosomal effects were found in CHO cells at toxic concentrations,
but not in human lymphocytes.
Governini et al. (2015) collected semen samples from 59 healthy nonsmoking patients
attending a Center for Couple Sterility conference at the University in Siena, Italy. The subjects
were divided into those that were normozoospermic (13) and those that were oligoasthenoterato-
zoospermic (46). PFOA was present in 75% of the seminal plasma samples and only 16% of the
blood samples. Conversely, PFOS was present in 25% of the seminal plasma samples and 84%
of the serum samples. Sperm were evaluated for the presence of aneuploidy and diploidy, and
sperm DNA was evaluated for fragmentation using the terminal deoxynucleotidyl transferase
dUTP nick end labeling (TUNEL) assay. The frequencies of aneuploidy and diploidy were
significantly greater in the PFAS-positive samples than in the PFAS-negative samples (P<0.001
and P<0.05, respectively), suggesting the possibility for errors in cell division. The levels of
fragmented chromatin were significantly increased (P<0.001) for the PFAS-positive group
compared with the PFAS-negative group.
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PFOA was tested twice in the in vivo mouse micronucleus assay. PFOA did not induce any
significant increases in micronuclei and was considered negative under the conditions of this
assay (Murli 1995, 1996d).
G. Zhao et al. (2010) used AL cells to determine the mutagenicity of PFOA to mammalian
cells. AL cells are a human-hamster hybrid containing CHO-K1 chromosomes and a single copy
of human chromosome 11. The significance of human chromosome 11 is that it encodes for
expression of the human cell surface protein CD59. AL and mitochondria-deficient AL cells were
incubated with 0, 1, 10, 100, and 200 umol PFOA for up to 16 days and used in the 3-(4,5-
dimethylthiazol-2-yl)-2,5-diphenyltetrazolium bromide (MTT) viability, mutation, or caspace
assays. Reactive oxygen species (ROS), nitric oxide, and superoxide anion production were
measured in the cells, and the effects of ROS/reactive nitrogen species quenchers [0.5% dimethyl
sulfoxide (DMSO) and 0.2 mM NG-methyl-L-arginine, respectively] on mutagenicity and
caspace activities were determined. At 100 and 200 umol 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 umol PFOA. There was no increase in
mutations in mitochondria-deficient AL cells after incubation with 100 or 200 umol PFOA.
Production of ROS, nitric oxide, and superoxide anion was significantly increased at 100 and
200 umol PFOA after incubation of AL cells for 1, 4, and 16 days. Incubation with DMSO to
inhibit ROS production significantly decreased the CD59 mutation frequency caused by
200 umol PFOA after the 16-day incubation. In contrast, mitochondria-deficient AL cells had no
increase in ROS or superoxide production after incubation with up to 200 umol PFOA for
16 days.
To assess whether PFOA could induce the apoptotic pathway, caspase-3/7 and caspase-9
were examined in intact AL cells (mitochondria-deficient cells were not examined). The highest
concentration significantly increased caspase 3/7 and 9 activities after 1- and 4-day incubations.
Incubation with 0.5% DMSO and 0.2 mM NG-methyl-L-arginine significantly decreased the
increased caspace activity induced by 200 umol PFOA. The results led the authors to suggest
that mitochondrial-dependent ROS might play an important role in PFOA-induced mutagenicity
and that induction of caspase activities might be mediated by reactive oxygen and nitrogen
species.
3.3.2 Immunotoxicity
The impact of PFOA on the immune system has been the subject of considerable research,
primarily in mice. A number of the early studies by Yang et al. (2000, 2001, 2002a) used high-
dose exposures of 0.02% to 0.05% PFOA. Later studies by DeWitt et al. (2008, 2009, 2015) and
Loveless et al. (2008) used a range of doses from < 1 mg/kg/day to 30 mg/kg/day. Most of the
studies focused on responses associated with the spleen and thymus. Some of the effects
observed were PPARa-associated, but others are totally or partially independent. There is
evidence for full or partial reversal of effects in those studies that incorporated a recovery phase.
One study of immunotoxicity used the dermal route of exposure (Fairley et al. 2007).
Rat. Loveless et al. (2008) administered 0, 0.3, 1, 10, and 30 mg linear PFOA/kg by oral gavage
to groups of male CD rats (n = 10 per group) for 29 days. The animals received a dose of SRBC
on day 23. A separate group of high-dose rats were injected with water instead of SRBC. Rat
body weight was recorded on days 0, 3, and 6-28. At necropsy, blood was collected for
evaluation of immune system parameters. Cell counts were determined for the thymus and
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spleen. Total spleen and thymocyte cell counts and organ weights in exposed rats were
comparable to control. Microscopic examination of the thymus, mesenteric lymph nodes, and
popluteal lymph nodes revealed no effects in treated rats resulting from PFOA exposure. There
was no difference observed in immunoglobulin (IgM) liters between treated and control rats. The
immunological NOAEL was 30 mg/kg/day.
Mouse. Yang et al. (2000, 2001, 2002a, 2002b) completed a series of studies investigating the
immunotoxic effects of PFOA. In the first study, Yang et al. (2000) examined the liver, spleen,
and thymus effects of several known PPARa activators, including PFOA. The researchers
administered 0.02% PFOA (-40 mg/kg/day) to male C57BL/6 mice in the diet for 2, 5, 7, and
10 days. At the end of the feeding period, mice were sacrificed and the liver, spleen, and thymus
were weighed. Administration of PFOA resulted in a significant increase in liver weight relative
to control even at day 2. Following 5 days of administration, significant decreases in thymus and
spleen weight were noted.
A second component of the Yang et al. study (2000) examined the effect of 0.02% PFOA in
the diet on the cellularity, cell surface phenotype, and cell cycle of thymocytes and splenocytes.
After 7 days, significant decreases in the total number of thymocytes (85%) and splenocytes
(80%) were observed. There is a pattern to the development of thymocytes that should be
considered when evaluating the impact of chemicals on their differentiation. Early thymocytes
formed in the bone marrow do not express CD4 or CDS (CD4"CD8"). In the thymus, they
differentiate and express both CD4 and CDS (CD4+CD8+). They also undergo proliferation and
downregulation of either the CD4 or CDS protein expression to become either a CD4 or CDS
thymocyte (Yang et al. 2000). Following exposure to PFOA, the number of thymocytes
expressing neither CD4 nor CDS decreased by 57%; the number expressing both CD4 and CDS
decreased by 95%; the number expressing only CD4 decreased by 64% while those expressing
only CDS decreased by 72%. As detected by cell cycle flow cytometry analyses, thymocyte
proliferation was inhibited based on the number of cells in each stage of the cell cycle.
T-cells (CDS*) and B-cells (CD19*) in the spleen decreased by 75% and 86%, respectively.
Splenic T-cells are lymphocytes produced in the thymus that carry the CD3+ surface protein
marking them as T-cells for exportation to the spleen. There are several classes of T-cells that are
characterized by surface proteins. Yang et al. (2000) found significant decreases in helper
CD3+T-cells with CD4+ surface proteins (78%) and cytotoxic CD3+T-cells with CD8+ surface
proteins (74%). The authors suggested that, unlike the CD3+T-cells that originate in the thymus,
the decrease in CD19+B-cells of the spleen reflects decreased differentiation and maturation in
the bone marrow where they are formed.
In the final phase of the Yang et al. study (2000), the effects of in vitro exposure of
thymocytes and splenocytes to PFOA were examined. The in vitro studies showed spontaneous
apoptosis occurring in splenocytes and thymocytes after 8 or 24 hours of culturing in the
presence of varying concentrations of PFOA (50, 100, and 200 umol). However, under the
exposure conditions, PFOA did not appear to significantly alter the cell cycle. The only dose
tested (-40 mg/kg/day) was a LOAEL for its effects on the immunoactive products of the
thymus and spleen. Recovery can occur with the cessation of exposure as illustrated by the Yang
et al. study (2001) described below.
Yang et al. (2001) reported on their examination of the immunosuppressive effects of PFOA.
As was the case in their earlier publication (Yang et al. 2000), the 2001 report includes several
components. A diet of 0.02% PFOA (-40 mg/kg/day) was fed to C57BL/6 mice for 2-10 days.
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One group of animals was exposed to PFOA each day until day of sacrifice on days 2, 5, 7, and
10. At sacrifice, body, liver, and spleen weights were recorded. A second group of animals was
dosed according to the same schedule, but dosing ceased after day 7, and the animals were fed
normal diets for 2-10 days to monitor recovery from the effects of exposure. In the recovery
group, animals were sacrificed after 2-, 5-, and 10-day recovery periods.
The mice that received 0.02% PFOA for up to 10 days experienced significant increases in
liver weight compared to controls beginning at day 2. Significant decreases in thymus and spleen
weights were observed starting on day 5. Body weight increased for the first 2 days of the study
and decreased continuously for the remainder of the exposure period. The activity of palmitoyl-
CoA and lauryl-CoA, biomarkers for PPARa activation and peroxisome proliferation, also were
increased significantly and increasingly across the exposure period. The impact of PFOA
exposure was similar to that observed in the Yang et al. study (2000). After administration for
7 days, the number of thymocytes expressing neither CD4 nor CDS decreased by 65% following
exposure to PFOA; the number expressing both CD4 and CDS decreased by 95%; and the
number expressing either CD4 or CDS decreased by 65% and 75%, respectively. T-cell (CD3+)
splenocytes and B-cell (CD19+) splenocytes decreased by 65% and 75%, respectively. As
detected by cell cycle flow cytometry analyses, thymocyte but not splenocyte proliferation was
inhibited.
The animals that participated in the recovery portion of this study rapidly regained their body
weight starting on the second day after withdrawal of PFOA. However, the liver weight failed to
recover even after 10 days. Thymus weight recovery started on day 2 and was completed by day
10. The spleen weights returned to normal by day 2 post-withdrawal. The increases in thymus
and spleen weight during recovery were paralleled by increases in total thymocyte and
splenocyte counts. Thymocyte recovery was apparent on day 5 and complete by day 10, although
during the first two days of the recovery period, further decreases in the CD4+CD8+, CD4+ and
CD8+ cells were observed. Flow cytometry evaluation of the distribution of the cells across the
cell cycle in the recovery group animals demonstrated increases in cell proliferation following
removal of PFOA from the diet. However, final cell counts did not reach the control values for
the thymocyte (CD4+ and CD8+) or splenocyte (CD3+ and CD19+) populations evaluated.
In the second component of the Yang et al. study (2001), C57BL/6 mice were administered
diets consisting of 0.001%-0.05% PFOA (w/w) for 10 days. These doses are equivalent to
approximately 2.0-100 mg/kg/day. There was a dose-related decrease in spleen and thymus
weights and a dose-related increase in liver weights accompanied by a corresponding increase of
palmitoyl-CoA and lauryl-CoA activity. Enzyme activity was significantly increased for all
doses. Spleen and thymus weights were significantly decreased at doses > 0.01% and higher but
not at the lower doses; the increases in liver weights were significantly increased for the 0.02%
and 0.05% doses. With the testing of a broader range of doses, -20 mg/kg/day was found to be a
LOAEL for effects on the thymus and spleen and the ~6 mg/kg/day dose a NOAEL.
Yang et al. (2002a) examined the possible involvement of PPARa in the immunomodulation
exerted by PFOA. This study made use of transgenic PPARa-null mice (Sv/129), which are
homozygous with regards to a functional mutation in the PPARa 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.
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The results showed that, in contrast to wild-type mice, feeding PPARa-null mice PFOA
resulted in no significant decrease in body weight. Liver weight in PPARa-null mice fed the
PFOA diet was significantly increased when compared to control PPARa-null mice, but not
when compared to wild-type PFOA-exposed mice. Peroxisome proliferation, as measured by
fatty acid oxidation, was totally lacking in PPARa-null mice. Also, in contrast to wild-type mice,
feeding PPARa-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 PPARa-null mice
compared to the PPARa-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 PPARa-null mice than they were in the PFOA-exposed wild-type mice, but higher than
in the PPARa-null control mice. PFOA treatment caused no significant change in splenocyte
proliferation in PPARa-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 PPARa. However, there
were some impacts on the thymus and liver that were independent of PPARa receptor activation.
PPARa-null mice still showed increases in liver weight and effects on the thymus (small
decrements in organ weight, thymocyte cellularity, and proliferative cell cycle) following a 7-day
exposure to approximately 40 mg/kg/day PFOA that were independent of PPARa.
Yang and colleagues extended their studies of the immunotoxicity of PFOA in a feeding
study designed to examine the effects of PFOA on specific humoral immune responses in mice
(Yang et al. 2002b). For this study, 0.02 % PFOA was administered to male C57BL/6 mice for
10 days. The animals were then evaluated via plaque-forming cell (PFC) and serum antibody
assays for their ability to generate an immune response to horse red blood cells (HRBCs).
Ex vivo and in vitro splenic lymphocyte proliferation assays also were performed. The results
showed that mice fed normal chow had a strong humoral response to challenge the HRBCs, as
measured by the PFC assay. In contrast, mice fed PFOA responded to HRBC immunization with
no increase in HRBC-specific PFCs, relative to unimmunized controls. However, in experiments
where PFOA-treated mice received normal chow following HRBC immunization, there was a
significant recovery of the numbers of specific PFCs stimulated. The suppression of the humoral
immune response by PFOA was confirmed by analysis of the serum anti-HRBC response.
In ex vivo experiments, splenocytes isolated from control mice responded to both
concanavalinA (ConA) and lipopolysaccharide (LPS) with lymphocyte proliferation, as
measured by thymidine incorporation. However, treating mice with 0.02% PFOA for 7 days
attenuated the proliferation. In a set of in vitro experiments, PFOA (1-200 umol) added to the
culture medium of splenocytes cultured from untreated mice did not cause an alteration of
lymphocyte proliferation in response to LPS or ConA.
DeWitt et al. (2008) expanded the repertoire of studies of the immunological effects of
PFOA by examining various aspects of humoral (antibody production) and cellular immunity.
The first component of their publication had many similarities with the Yang et al. study (2001).
Adult female C57BL/6J mice (n = 40 per endpoint and 8 per group) were exposed to a single
daily dose of 30 mg PFOA/kg/day in distilled water by gavage for 10 continuous days. After 10
continuous days of exposure, half of the mice continued receiving PFOA from day 11 through
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day 15 (constant group) while the other half received distilled water from day 11 through day 15
(recovery group). On day 11,16 mice per group were immunized with sheep red blood cells
(SRBC) and eight mice per group were injected with BSA. Sacrifices took place on day 16
(1 day postexposure period) and day 31 (15 days postexposure period). Vehicle and cage
controls also were included in the study. All groups were monitored for the following effects:
Body weight and organ weights (day 16, day 31)
Serum IgM levels (day 16)
Delayed-type hypersensitivity (DTH) foot-pad response to BSA (day 26)
Serum IgG levels after booster immunization with SRBC on day 20 (day 31)
The results for body and organ weights were comparable to those in the Yang et al. study
(2001). Body weight was significantly decreased from days 8-11 for both PFOA-treated groups
and on day 16 for mice in the constant exposure group. By day 31, there were no body weight
differences between the groups. Relative liver weight was significantly elevated in both 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 postexposure day 1 in both the recovery and constant exposure groups. There were no
significant differences from controls for SRBC-specific IgG levels and for DTH foot-pad
responses to the BSA challenge.
The C57BL/6 mice used for the continuous-dosing versus recovery component of the DeWitt
et al. study (2008) were found to develop ulcerative dermatitis following the PFOA exposure. It
was determined that this effect was a genetic susceptibility in the strain, and they were not used
for the dose-response component of the study; the C57BL/6N strain was used in its place.
Two studies of dose-response were included in the DeWitt et al. (2008) publication. Groups
of 16 female C57BL/6N mice were given 0, 3.75, 7.5, 15, and 30 mg PFOA/kg/day in the
drinking water for 15 days during the first experiment. In the second experiment, the doses were
0, 0.94, 1.88, 3.75, and 7.5 mg PFOA/kg/day administered for 15 days in the drinking water. The
immunological sensitization and postdose monitoring were identical to that used in the constant-
dosing versus recovery experiment.
In the first experiment, body weight was significantly decreased from day 8-16 at 30 mg/kg
PFOA and on day 16 at 15 mg/kg PFOA. As observed previously, liver weights were
significantly elevated at day 16 and day 31 at all doses. Absolute and relative spleen and thymus
weights were significantly decreased at >15 mg/kg PFOA on day 16. With the exception of the
absolute thymus weight at 15 mg/kg PFOA, all spleen and thymus weights were similar to
weights in controls 15 days after dosing. The IgM response to SRBC was significantly reduced at
>3.75 mg/kg PFOA in a direct dose-related manner. The IgG response to SRBC challenge was
slightly but significantly elevated at 3.75 and 7.5 mg/kg PFOA but similar to that of the control
level at the higher doses. Thus, there was a direct response of IgM, but not IgG, to dose across
the dose levels. There was no significant change in the DTH response at any dose. The LOAEL
from the first experiment was 3.75 mg/kg/day dose based on decreased IgM and increased IgG
response to SRBC immunization and increased liver weights (p<0.05).
The second dose-response experiment confirmed the 3.75 mg/kg/day dose as the
immunological LOAEL on the basis of significantly decreased spleen weight, decreased IgM
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levels on day 16, and increased IgG levels on day 31. The immunological NOAEL was
1.88 mg/kg/day. BMD analysis of IgM serum titer data gave a lower bound 95% confidence
limit of 1.75 mg/kg/day on a BMD (one SD) of 3.06 mg/kg/day. Liver weight was significantly
increased at all doses on days 16 and 31. The LOAEL for increased liver weight was 0.94 mg/kg
PFOA.
As mentioned earlier, some of the immunological responses observed in the studies of
immunotoxicity are linked to PPARa activation by PFOA. DeWitt et al. (2015) published results
for a study in female PPARa KO mice (B6.129S4-PpartmlGonzN12 mice) and compared them to
the response of female C57BL/6-Tac wild-type mice. Both T-cell-dependent and T-cell-
independent antibody production were evaluated. The doses used in the study of the T-cell-
dependent responses were 0, 7.5, and 30 mg PFOA/kg/day dissolved in deionized drinking water
for 14 days. On day 11, the animals were injected with SRBCs to stimulate an immune response.
PFOA dosing continued for 4 more days (15 days dosed); the following day, the animals were
sacrificed. Body weight was significantly decreased only in wild-type mice at 30 mg/kg/day.
Relative spleen weights were significantly decreased (P<0.05) in the wild-type but not the KO
mice at 30 mg/kg/day of PFOA. Relative thymus weights were significantly decreased in the
wild-type mice at 7.5 mg/kg/day, but not in the KO mice at either dose or the wild-type mice
receiving 30 mg/kg/day. There was a significant (P<0.05) reduction in the IgM antibody
response to the SRBC injection at 30 mg/kg/day for both the wild-type and KO animals (n = 6),
indicating that the response was not totally related to PPAR-a activation. The NOAEL in wild-
type and KO animals was 7.5 mg/kg/day and the LOAEL 30 mg/kg/day based on decreased T-
cell-dependent IgM antibody response to SRBC.
To evaluate T-cell-independent responses to PFOA, groups of eight C57BL/6N mice were
given doses of 0, 0.94, 1.88, 3.75, and 7.5 mg/kg/day in their drinking water. On day 11, they
were injected with the T-cell-independent antigen dinitrophenyl ficol. At sacrifice (day 16),
blood was collected for analysis of IgM antibody liters. There was a significant decrease
(p<0.05) in the antibody response by 9.4-10.7% in the animals receiving doses > 1.88 mg/kg/day
of PFOA. The NOAEL for the T-cell-independent response to dinitrophenyl ficol was
0.94 mg/kg/day of PFOA and the LOAEL was 1.88 mg/kg/day. Thus, suppression of the T-cell-
independent response occurred at a lower dose (1.88 mg/kg/day) than the dose resulting in
suppression of the T-cell-dependent response (7.5 mg/kg/day).
The authors looked at changes in lymphocyte populations at 10, 13, and 15 days of exposure
in the female C57BL/6N mice and saw no significant dose-dependent changes in lymphocyte
cell types. They concluded that both sets of responses were due to changes in cellular function
rather than lymphocytotoxicity (DeWitt et al. 2015).
Loveless et al. (2008) administered 0, 0.3, 1, 10, and 30 mg linear PFOA/kg by oral gavage
to groups of male CD-I mice (n = 20 per group) for 29 days. The animals received a dose of
SRBC on day 24. A separate group of high-dose mice was injected with water instead of SRBC.
Mice were weighed daily. At necropsy, blood was collected for evaluation of immunity
parameters. Cell counts were determined for the thymus and spleen.
Absolute and relative spleen and thymus weights were significantly decreased at
>10 mg/kg/day. The relative spleen weight of mice dosed with 1 mg/kg/day also was
significantly decreased compared to control animals. Spleen and thymus cell counts were
significantly decreased and minimal to severe lymphoid depletion/atrophy of the thymus was
observed at >10 mg/kg/day. IgM liters were significantly decreased at >10 mg/kg/day. Serum
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CORT was significantly increased at 10 mg/kg/day and elevated (not statistically significant) at
30 mg/kg/day. When IgM and CORT were plotted against each other, a negative correlation
coefficient suggested that increasing CORT levels decreased the ability to make SRBC antibody.
The LOAEL was 1 mg/kg/day based on decreased spleen weight, and the NOAEL was
0.3 mg/kg/day. Mice appeared to be more susceptible than rats to immunosuppression from
PFOA.
Loveless et al. (2008) hypothesized that at least a portion of the thymic response to PFOA
might be related to physiological stress and increased levels of CORT hormones. DeWitt et al.
(2009) investigated this hypothesis by comparing the immunological response of
adrenalectomized (ADX) C57BL/6N female mice to that of sham-operated female mice from the
same strain. The animals were dosed with 0, 3.75, 7.5, and 15 mg PFOA/kg/day in the drinking
water for 10 days. Body weight was recorded on dosing days 0, 4, and 8, plus 2 and 5 days
postdosing. On exposure days 5 and 10, blood and serum were collected for analysis of a broad
array of clinical chemistry parameters, including activity of liver enzymes indicative of cellular
damage (e.g., ALP, AST, ALT, LDH, GOT, and SDH), serum lipids (cholesterol and
triglycerides), and CORT. A baseline measure of CORT was determined from serum samples
collected before dosing began. One day after cessation of exposure, the mice were immunized
with SRBC. Four days later, serum was collected and the levels of CORT and IgM were
determined.
Body weight in the sham-operated mice declined during dosing in the highest dose group but
recovered by 5 days postdosing. In the ADX mice, body weight declined during dosing at 7.5
and 15 mg PFOA/kg/day, but recovered in mice receiving 7.5 mg PFOA/kg/day by 5 days
postdosing. There were significant increases in ALT, AST, LDH, and SDH at the highest dose
for the ADX mice indicative of damage to hepatic cell membranes (Table 3-33).
Serum levels of triglycerides significantly decreased compared to controls, with all doses for
the sham-operated mice on day 5 of dosing but only for the 7.5- and 15-mg/kg/day doses in the
ADX mice. Cholesterol levels were significantly decreased (p<0.05) in the sham-operated mice
at the highest dose, but no differences in cholesterol levels were observed in the ADX mice.
Table 3-33. Selected Clinical Chemistry Parameters in Mice Treated with PFOA for 5 Days
Dose (mg/kg/day)
ALT
AST
LDH
SDH
Sham-Operated
0
3.75
7.5
15
39.52+2.50
43.88+0.93
56.96+6.78
62.57+3.15
121.56+17.96
104.07+10.24
95.55+10.22
89.07+1.30
320.57+29.84
293.92+68.65
262.71+35.60
191.76+22.25
46.43+1.03
39.31+3.32
39.02+7.77
46.87+1.46
ADX
0
3.75
7.5
15
26.96+1.78
29.67+1.62
39.04+2.59
94.23+31.66*
73.53+4.70
76.58+3.38
83.79+8.94
126.47+16.39*
176.50+19.32
222.69+19.18
320.45+53.34
435.57+81.42*
33.05+1.58
37.95+2.35
46.35+1.42
77.61+19.89*
Source: DeWitt et al. 2009
Note: * = p<0.05 versus corresponding sham control or ADX control group.
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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 3-34).
Table 3-34. Selected Clinical Chemistry Parameters in Mice Treated with
PFOA for 10 Days
Dose (mg/kg/day)
ALT
AST
LDH
SDH
Sham-Operated
0
3.75
7.5
15
51.51+14.62
79.26+33.87
135.57+38.18
344.53+235.63
93.30+6.33
123.73+15.20
142.66+15.59
242.92+117.62
333.48+86.86
404.14+59.89
490.44+69.14
595.01+137.37
54.60+16.72
45.50+10.15
80.71+14.59
89.20+26.03
ADX
0
3.75
7.5
15
128.22+24.80
282.23+193.54
89.79+21.54
261.14+75.95
106.00+8.86
217.10+3.48
99.78+12.59
181.40+32.94
236.96+30.23
379.61+80.67
574.65+236.38
614.05+144.95
61.88+8.87
68.78+24.88
52.07+11.98
101.93+24.00
Source: DeWitt et al. 2009
At the end of dosing, corticosteroid levels in the sham-operated animals were greatly
elevated compared to the levels in the control animals at all doses, and the difference was
statistically significant at the highest dose. By 5 days postdosing, the CORT levels had declined
for all doses but were still elevated compared to controls for the 7.5- and 15-mg/kg/day groups.
In the animals lacking their adrenal glands, there were no statistically significant differences in
the hormone levels. IgM levels were significantly lower than controls at the highest dose for the
sham-operated animals and at the two highest dose groups for the ADX mice. However, when
comparing the sham mice to the ADX mice, the only significant difference in IgM was found for
the 7.5-mg/kg/day animals. On the basis of data, it appears that stress-related CORT production
did not have a major impact on the IgM response to the SRBC inoculation.
Son et al. (2009) administered 0, 2, 10, 50, and 250 ppm PFOA (equivalent to 0, 0.49, 2.64,
17.63, and 47.21 mg/kg PFOA) in the drinking water to 4-week-old male ICR mice for 21 days
to determine if PFOA alters T lymphocyte phenotypes and cytokine expression in mice. The
spleen, thymus, and trunk blood were collected at sacrifice. Sections of the spleen and thymus
were processed for histological examination. Splenic and thymic expression of mRNA from
proinflammatory cytokinesincluding tumor necrosis factor-a, interleukin-lp, and interleukin-
6, and the proto-oncogene c-mycwere 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-a, interleukin-lp, interleukin-6, and c-myc expression were significantly elevated at
250ppm in the spleen. Interl eukin-lp expression also was elevated at 50 ppm in the spleen. In the
thymus, c-myc expression was significantly elevated by treatment with 50 and 250 ppm PFOA.
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The splenic and thymic lymphocyte population was altered by PFOA treatment, as shown in
Table 3-35. A 50% decrease was observed in splenic CD8+ lymphocytes at all PFOA doses, and
increases in splenic CD4+ lymphocytes of 43% and 106% at 50 and 250 ppm PFOA,
respectively, were observed. In the thymus, a 110% increase was observed in thymic CD8+
lymphocytes at 250 ppm, but thymic CD4+ lymphocyte populations were not affected and
CD4+CD8+ populations were decreased at 50 and 250 ppm PFOA. The lowest dose tested
(0.49 mg/kg/day) was a LOAEL for CD4' and CD8+ splenocytes.
Table 3-35. Impact of PFOA on Splenic and Thymic Lymphocyte Populations
Dose (mg/kg/day)
Spleen
CD4-CD8-
CD4+CD8-
CD4-CD8+
CD4+CD8+
Thymus
CD4-CD8-
CD4+CD8-
CD4-CD8+
CD4+CD8+
0.49
T
-
4
4
-
-
-
-
2.64
-
-
1
1
-
-
-
-
17.63
-
T
1
-
-
-
-
1
47.21
1
T
1
-
t
-
t
4
Source: Son et al. 2009
Notes:
t Significantly increased compared to control (p < 0.05).
I Significantly decreased compared to control (p < 0.05).
- Not significantly different from control.
Qazi et al. (2009) investigated the impact of PFOA on the innate immune system. Adult male
C57BL/6 (H-2b) mice were administered 0.001% and 0.02% PFOA (~2 or 40 mg/kg) in the diet
(w/w) for 10 days. After the last dose, all mice were sacrificed. Sacrifice was delayed for a
subset of the animals until 2 hours after they had received an LPS injection to stimulate an
immunological response. Blood, peritoneal exudate cells, liver, epididymal fat, spleen, thymus,
and bone marrow were recovered. The blood, peritoneal exudate, bone marrow, and spleen were
evaluated for total and differential white blood cell counts and concentrations of tumor necrosis
factor (TNF-a) and interleukin 6 (IL-6).
Consistent with other studies of the 0.02% dose, there was a significant increase in liver
weight after the 10-day exposure. Body weight, thymus weight, spleen weight, and epididymal
fat depots were decreased. Food consumption in these animals was reduced by 35%, which
might have played a role in the reduced body, organ, and tissue weights. Compared to the
controls, there was a significant decrease in total white cells, lymphocytes, and neutrophils at
0.02% PFOA. This same dose was associated with a decrease in total white cell count in bone
marrow and spleen, and an increase in the proportion found as macrophages in the bone marrow,
spleen, and peritoneal cavity. Although the total number of macrophages was not reduced in the
peritoneal cavity and spleen, it was reduced in the bone marrow. The increase in the proportion
of macrophages reflects a decrease in other white cell populations. There was significant increase
in the concentration of IL-6 in all of the 0.02 % dosed animals, but only those receiving the LPS
injections showed a significant increase in TNF-a. The 0.001% dose (about 2 mg/kg/day) was a
NOAEL.
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The data available on immunological responses of animals following oral exposure to PFOA
are extensive, especially as they apply to mice. A number of the studies used doses of about
40 mg/kg/day. However, studies conducted at lower doses find effects on the spleen and/or
thymus at doses from 0.5 to 2 mg/kg/day. Activation of PPARa appears to augment the response
based on studies in PPARa-null mice but is not necessary (Yang et al. 2002a). There are
differences between mice and rats with mice showing a response at a lower dose (Loveless et al.
2008). Cessation of dosing can reverse some of the observed effects in mice (Yang et al. 2001).
Inhalation Exposure
No data on the effects of inhalation exposure on immunological endpoints were identified in
the literature.
Dermal Exposure
Fairley et al. (2007) carried out a complex study of toxicity and respiratory hypersensitivity
to ovalbumin (OVA) as impacted by dermal exposure to PFOA dissolved in acetone compared to
acetone alone. There were several phases to the study. In the first phase, a range-finding study,
PFOA was applied to each ear of female BALB/c mice (n = 5 per group) at doses of 0, 0.01%,
0.1%, 0.25%, 0.5%, 1.0%, and 1.5% PFOA (equivalent to 0, 0.25, 2.5, 6.25, 12.5, 25, and
50 mg/kg/day) for 4 days. Six days after last inoculation, the animals were sacrificed. The liver,
spleen, and thymus were recovered and weighed. A significant increase in liver weight was
observed at >6.25 mg/kg PFOA. Spleen weight was significantly decreased in mice dosed with
25 mg/kg and 50 mg/kg PFOA, and thymus weight was significantly decreased in mice at the
highest dose (p<0.05). The cell counts in the spleen were significantly decreased compared to
control at all doses and for the highest two doses in the thymus. The LOAEL was
6.25 mg/kg/day based on a statistically significant increase in liver weight (p<0.01), and the
NOAEL was 2.5 mg/kg/day.
In the second phase of the Fairley et al. study (2007), groups of 5-15 animals were dosed
dermally on the ears for 4 days with doses of 0, 0.5%, 0.75%, 1.0%, and 1.5% PFOA (equivalent
to 0, 12.5, 18.75, 25, and 50 mg/kg/day). On days 1 and 10, they were injected i.p. with either
2.0 mg alum or 7.5 ug OVA and 2.0 mg alum in a phosphate-buffered saline solution (lOOuL).
Four days after the last inoculation, the animals were sacrificed and blood was collected by
cardiac puncture. Liver, spleen, and thymus were recovered and weighed; spleen and thymus
cellularities were determined. A significant (p<0.01) increase in liver weight and decrease in
spleen weight and spleen cellularity occurred at all doses. Thymus weight and cellularity were
significantly decreased (p<0.01) at >18.75 and > 25 mg/kg/day, respectively. There were no
significant differences in the CD4+, CD8+, CD4"8"' or CD3e T-cells. CD3e protein is expressed
by both thymocytes and mature T-cells.
Levels of IgE and OVA-specific IgE were measured in the control and dosed animals by
enzyme-linked immunosorbent assay. IgE is the immunoglobulin that is best correlated with
respiratory allergic responses. It functions to stimulate mast cells and basophils to release
histamine and other mediators of inflammation (Saladin 2004). The IgE response was increased
in a dose-related fashion compared to the OVA control for all the PFOA-treated animals; the
increase was significant (p<0.05 or 0.01) at doses >18.75 mg/kg/day. The OVA-specific IgE
response did not demonstrate a direct response to dose, but there was a significant increase
(p<0.05) for the 18.75- and 25-mg/kg/day groups. The OVA-specific response for the three
highest dose groups was practically indistinguishable from the OVA control.
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The dermal LOAEL was 12.5 mg/kg/day based on increased liver weight and decreased
spleen weight and cellularity. No NOAEL was established.
In the third part of the Fairley et al. study (2007), mice (n = 5) were dosed dermally via their
ears for 4 days as described above (0, 12.5, 18.75, 25, and 50 mg/kg/day). On days 19 and 26
after the start of dosing, they were challenged by pharyngeal aspiration of 250 |j,g OVA in the
phosphate-buffered saline vehicle and sacrificed 24 hours after the last challenge. There was a
dose-related decrease in number of splenocytes carrying the B220+ marker (expressed on B-cells,
activated B-cells, and subsets of T- and natural killer- [NK-] cells) compared to the OVA
controls. The changes were significantly different for the 25-mg/kg/day (p<0.05) and
37.5-mg/kg/day (p<0.01) groups.
After the day 19 challenge, the mice (n = 5) were placed in a plethysmography chamber for
measurement of enhanced pause airway respiration (PenH) values. PenH values reflect volume
of air in the lungs. Once in the chamber, they were challenged with nebulized methacholine for
3 mins followed by 2 mins of fresh air. The PenH measures were recorded every 30 seconds over
the next 5 hours. The area under the plasma concentration time curve (AUC) for the PenH
measures was determined after correction for baseline (acetone control, no OVA or PFOA). An
AUC of 1.6 was considered to be a positive response. Twenty-four hours later, blood was drawn
from the abdominal artery and the mice were sacrificed. The lungs were recovered for
histological analysis. An increase in antigen-specific hyperactivity response to PFOA, in both the
PenH values and the number of animals responding, was observed at doses up to about
25 mg/kg/day. The PenH AUC was significantly (p<0.05) increased in mice treated with
25 mg/kg/day PFOA and OVA compared to the OVA control mice, but there was no significant
difference between the OVA control and the animals exposed to 50 mg/kg/day PFOA and OVA.
The LOAEL for the PenH response was 25 mg/kg/day, and the NOAEL was 18.75 mg/kg/day.
Histopathological evaluation of the lungs revealed macrophage and eosinophil infiltration in
response to the challenge with 250 jig OVA by pharyngeal aspiration. The severity of the
response increased with increasing concentrations of PFOA. Eosinophils and macrophages were
found in the interstitial, peribronchiole, and perivascular areas. Neutrophils, lymphocytes, and
some multinucleated giant cells also were observed. Increased secretory matter, sloughing of
epithelium, and secretory cell necrosis were observed in mice exposed to all concentrations of
PFOA and OVA. The response was not observed in the mice exposed to only PFOA. The authors
concluded that dermal exposure to PFOA was immunotoxic and enhanced the airway
hypersensitivity response to OVA suggesting that PFOA may augment the IgE response to
environmental allergens.
In vitro. In a pilot study, Brieger et al. (2011) examined the effects of PFOA on human
leukocytes. Peripheral blood mononuclear cells (PBMCs) were obtained from the blood of 11
voluntary donors (n = 6 females, 5 males). PBMCs were incubated with varying concentrations
of PFOA followed by assays for cell viability, proliferation, and NK cell activity. The human
promyelocytic leukemia cell line, HL-60, was used in cell viability and monocyte differentiation
assays. The various components of the assays employed are identified as follows:
In the cell viability assay, the PBMCs were incubated with 0-500 ug/mL for 24, 48, and
72 hours, and HL-60 cells were incubated with 0-125 ug/mL PFOA for 24 hours.
In the proliferation assay, the PBMCs were incubated with 0-100 ug/mL PFOA for 24
hours, labeled with 6-carboxyfluorescein succinimidyl ester (CFSE), stimulated with
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ConA, a T-cell mitogen (5 ug/mL to half of all samples), and incubated for an additional
72 hours.
For the NK assays, PBMCs were incubated with 0-100 ug/mL PFOA for 24 hours
followed in incubation of 3 hours with K562 target cells (12.5:1 ratio) labeled with
CFSE.
In the monocyte differentiation assay, HL-60 cells were incubated with 0-100 ug/mL
PFOA for 72 hours. Half of each sample was stimulated with 25 nmol calcitrol,
la,25-dehydroxyvitamin D3 (1,25D3) 24 hours into the incubation period. Expression
of CD1 Ib and CD14 were measured as markers of differentiation.
Whole blood was incubated with 0-100 ug/mL PFOA in the presence or absence of
25 ug/mL phytohemagglutinin (PHA), T-cell cytokine secretion stimulator for 48 hours
in quantification assays for the cytokines TNF-a and IL-6. LPS (0 or 250 ng/mL) was
added to whole blood incubated with 0.1-100 ug/mL PFOA either 4 or 24 hours prior to
the end of the 48-hour incubation period to determine TNF-a 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 ug/mL PFOA significantly increased PBMC
viability at the 72-hour endpoint, and 62.5 ug/mL PFOA significantly increased cell viability at
24 hours. Exposure to 250 and 500 ug/mL PFOA significantly decreased cell viability at all time
endpoints. Exposure to PFOA did not affect HL-60 cell viability. A trend towards slightly
augmented proliferation was observed following incubation with PFOA. Of the nine samples
used, cells from six donors had slightly increased proliferation and t had no response. In cells
incubated with ConA and 100 ug/mL PFOA, a nonsignificant decrease in the number of
proliferating cells was observed. PFOA decreased NK cell activity approximately 16% (not
statistically significant). In the presence of l,25Ds, 100 ug/mL PFOA significantly increased the
percentage of HL-60 cells expressing CD1 Ib 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-a and IL-6
levels. In the presence of PHA, a slight dose-dependent decrease in TNF-a and IL-6 was
observed. There was a slight dose-dependent decrease in TNF-a release when LPS was added
4 hours before the end of the incubation period and a slight dose-dependent increase in IL-6
release when LPS was added 24 hours prior to the end of incubation. The authors also looked at
the correlation between basal PFOA concentration and cytokine release. A significant association
was observed between PFOA concentration and the release of LPS-induced TNF-a and IL-6 by
peripheral monocytes. The authors suggested that the trends observed at the lower concentrations
might show an impact on human immunity with a larger population.
Ahuja et al. (2009) examined the effects of PFOA on the production and activation of human
monocyte-derived dendritic cells. These cells are responsible for a primary immune system
response by activating lymphocytes and secreting cytokines. Peripheral monocytes and immature
dendritic cells were incubated with 200 jimol PFOA from day 0 to day 6 or 8 to determine the
impact on phenotype and cytokine secretion. Maturation stimulus (i.e., prostaglandin E2, tumor
necrosis factor, interleukin ip, and IL-6) was added during the last 48 hours of incubation to
induce dendritic cell maturation. Mixed leukocyte reaction assays were conducted to determine if
immature dendritic cells could stimulate T-cells. Cytokine (HLA-DR, CD25, CD80, CD83, and
CD86) expression was measured as a determination of maturity. HLA-DR is a cytokine that
presents antigens to elicit T-cell response. CD25, 80, 83, and 86 are cell receptors that act as co-
receptors in T-cell activation; and interleukin 12p40 and 10 stimulate T-cells. Mature cytokine-
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activated dendritic cells secrete interleukin 12p40 and interleukin 10 as antiinflammatory
cytokines.
In peripheral monocytes incubated with only PFOA from day 0-6 or day 0-8, expression of
HLA-DR and CD86 was increased compared to expression in control cells, indicating that
immature dendritic cells were present. In the mixed leukocyte reaction assay, the ability to
stimulate T-cells was not different between immature dendritic cells generated in the absence or
presence of PFOA.
To determine if PFOA impacted the differentiation of immature dendritic cells to mature
dendritic cells, immature dendritic cells were incubated with 200 jimol of PFOA for 6 days and
the maturation stimulus was added for the final 2 days of incubation. There were no differences
in cytokine (CD25, CD80, CD83, and CD86) expression between cells incubated with PFOA and
control cells. Expression of interleukin 12p40 and interleukin 10 was significantly inhibited by
PFOA in mature cytokine-activated dendritic cells, even in the presence of maturation stimulus
during the last 48 hours of incubation. The result suggested that exposure to PFOA during
generation of dendritic cells affected the phenotype and cytokine production of human dendritic
cells and could lead to immunomodulation in the development of the immune response.
3.3.3 Hormone Disruption
Thyroid. Martin et al. (2007) administered 20 mg PFOA/kg to adult male Sprague-Dawley rats
(n = 4 or 5) for 1, 3, or 5 days by oral gavage and determined the impact of PFOA on hormone
levels. Blood was collected via cardiac puncture and the serum was analyzed for cholesterol,
testosterone, FT4 and total T4, and total T3. RNA extracted from the livers was used for gene
expression profiling, genomic signatures, and pathway analyses to determine a mechanism of
toxicity.
Following a 1-day, 3-day, and 5-day dose, a significant decrease (p<0.05) was observed in
serum cholesterol (~|45-72%), total T4 (~|83%), FT4 (~|80%), and total T3 (~|25-48%).
Serum testosterone was significantly decreased (p<0.05, ~|70%) following a 3-day and 5-day
PFOA dose. PFOA treatment was matched to hepatotoxicity-related genomic signatures, as well
as signatures for hepatocellular hypertrophy, hypocholesterolemia, hypolipidemia, and
peroxisome proliferation. PPARa 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, andDiol, which deiodinates prohormone T4 to bioactivate T3,
were affected by PFOA. Treatment with PFOA resulted in significantly upregulated expression
ofDio3 and downregulated expression ofDiol (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 PPARy 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 PPARy and CAR. The
20-mg/kg/day dose was a LOAEL for the effects monitored after a 5-day exposure.
Reproductive Hormones. Cook et al. (1992) gavage-dosed male CD rats (n = 15 per group) for
14 days with 0, 1, 10, 25, and 50 mg PFOA/kg/day to examine the possibility that an endocrine-
related mechanism might explain Ley dig cell adenomas observed in rats. A separate control
group was pair-fed to the 50-mg/kg/day group. Blood and testicular interstitial fluid were
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collected at necropsy for hormone analysis including testosterone, estradiol, and LH. A separate
group of rats was dosed with 0 and 50 mg PFO A/kg/day for 14 days and challenged with
100 lUs of human chorionic gonadotropin (hCG) or 2 mg naloxone/kg 1 hour prior to necropsy
to induce testosterone concentrations. Blood was collected and analyzed for testosterone and LH.
Serum from rats challenged with 100 Ills hCG also was analyzed for P, 17
a-hydroxyprogesterone, and androstenedione.
The relative liver weight at 10, 25, and 50 mg PFOA/kg/day was significantly increased
(p<0.05). The accessory sex organ unit relative weight was significantly decreased (p<0.05) at
25 and 50 mg PFOA/kg/day compared to those weights in control rats. The relative weights of
the liver, accessory sex organ unit, and ventral prostate were significantly decreased at the
highest dose compared to the pair-fed control.
Serum estradiol was significantly increased at >10 mg PFO A/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 PFO A/kg after challenge with 100 Ills hCG. No differences in testosterone concentration
were observed in the naloxone-challenged rats, and no differences in LH were observed after
either challenge. In the hCG-challenged rats, androstenedione was significantly reduced at 50 mg
PFOA/kg, but no differences in concentrations were observed in P or 17 a-hydroxyprogesterone
between control and treated rats. The authors suggested that the observed decreased serum
testosterone levels could be due to decreased conversion of 17 a-hydroxyprogesterone to
androstenedione as a result of increased serum estradiol levels. The LOAEL was 10 mg/kg based
on increased liver weight and increased serum estradiol levels, and the NOAEL was 1 mg/kg.
Biegel et al. (1995) conducted in vitro, in vivo, and ex vivo studies to determine the effects of
PFOA on Leydig cell function. In the in vitro study, Leydig cells were cultured with ± 2 Ills
hCG (for final 3 hours) and 0, 100, 200, 250, 500, 700, and 1000 |imol PFOA for a total of
5 hours and then analyzed for testosterone concentration. Leydig cells also were incubated
±500 (imol PFOA and analyzed for testosterone and estradiol at 0, 0.5, 1, 3, 6, 12, 24, and
48 hours.
In the in vitro studies, there was no effect of PFOA treatment on testosterone in Leydig cells
cultured without hCG. In cells cultured with hCG, PFOA caused a dose-dependent decrease in
testosterone production. At 100 jimol PFOA plus hCG, the testosterone concentration was
significantly increased compared to cells treated with only 100 jimol PFOA. Cytotoxicity
occurred at > 750 jimol PFOA. In the time course experiment, 500 jimol PFOA significantly
inhibited hCG-stimulated release of testosterone at time points of at least 3 hours compared to
control. Estradiol levels of PFOA-treated Leydig cells at 48 hours were statistically greater than
the control.
Male CD rats were gavage-dosed for 14 days with 0, 0 pair-fed, or 25 mg PFOA/kg and
necropsied on day 15. Blood and testicular interstitial fluid were collected for hormone analysis.
Liver samples were collected for analysis of peroxisomal p-oxidation and microsomal aromatase
activities. Serum estradiol was significantly increased (p<0.05) by 25 mg PFOA/kg when
compared to the ad libitum and pair-fed control rats. Testicular interstitial fluid testosterone
concentration was significantly decreased (p<0.05) and microsomal aromatase activity, and
peroxisomal P-oxidation activity were significantly increased (p<0.05) in PFOA-treated rats
compared to the pair-fed control rats.
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Ley dig cells from the treated rats in the in vivo study were isolated and cultured for analysis
of testosterone concentration for the ex vivo study. An increase of 8.6-fold in testosterone
production (p<0.05) was observed in Leydig cells isolated from PFOA-treated rats. The authors
suggested that the increased serum estradiol levels resulted from liver aromatase induction by
PFOA, and that PFOA could directly affect Leydig cell function.
Liu et al. (1996) treated adult male Crl:CD(BR) rats (n = 15 per group) with 0, 0 pair-fed,
0.2, 2, 20, and 40 mg PFOA/kg for 14 days by oral gavage to determine the impact of PFOA on
aromatase activity. At necropsy on day 15, blood was collected for serum estradiol
determination. Liver samples were collected for determination of microsomal aromatase activity
and total P450 concentration. The testes were collected and testicular aromatase was determined.
The body weight of rats treated with > 20 mg PFOA/kg was significantly decreased (p<0.05)
compared to the control rats. Pair-fed control rats also had significantly decreased body weight
compared to the control rats. Body weight was not different between the pair-fed control rats and
rats dosed with 40 mg/kg PFOA. Absolute and relative liver weights were significantly increased
(p<0.05) at > 2 mg PFOA/kg. Relative testes weight was significantly increased at >20 mg
PFOA/kg, but the differences were due to decreased body weight. There were no differences
observed in testicular aromatase activity. In the remaining analysis, the pair-fed control group
was similar to the ad libitum control group. The protein yield of hepatic microsomes was
significantly increased at >0.2 mg PFOA/kg, and hepatic aromatase activity, total hepatic
aromatase activity adjusted for liver and body weight effects, and serum estradiol were
significantly increased (p<0.05) at >2 mg PFOA/kg. The maximum increase in total hepatic
aromatase activity was 16-fold and the increase was twofold for serum estradiol. A significant
correlation (p<0.0001) was observed between total hepatic aromatase activity and serum
estradiol. The aromatase activity in liver microsomes isolated from control rats and incubated for
2 hours with PFOA was significantly decreased at > 100 jimol. The authors estimated the half
maximal effective concentration (ECso) values for the outcomes, and they are shown in Table
3-36. Liu et al. (1996) concluded that the PFOA-increased protein yields suggested induction of
the ER resulting in aromatase induction, which led to increased serum estradiol. However, PFOA
also inhibited aromatase activity, which would explain why serum estradiol was only increased
up to twofold.
Table 3-36. Estimated ECso Values
Parameters
Hepatic microsome protein yield
Hepatic microsomal aromatase activity
Absolute liver weight
Relative liver weight
Serum estradiol
Terminal body weight
ECso (mg PFOA/kg)
0.53
0.76
1.07
1.56
3.24
11.65
Source: Liuetal. 1996
Note: EC5o= half-maximum response.
A separate component of the Liu et al. study (1996) examined the effect of PFOA on
aromatase activity in cultured hepatocytes and is discussed below. Aromatase is a cytochrome
P450 enzyme localized to the ER that catalyzes the conversion of androgens to estrogens. The
cultured hepatocytes isolated from control male CD rats were incubated with 0-1000 |imol
PFOA and the aromatase activity was evaluated after 18, 42, and 66 hours (Liu et al. 1996).
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Compared to aromatase activity in time-matched control cultures, PFOA caused a decrease in
aromatase activity after 18 and 42 hours incubation with hepatocytes and an increase after the
66-hour incubation period.
In their study examining the impact of PFOA on aromatase activity, Liu et al. (1996) also
examined the impact of PFOA on peroxisome p-oxidation and cytochrome P450 activities. Male
Crl:CD BR (CD) rats (n = 15 per group) were orally dosed with 0, 0 pair-fed, 0.2, 2, 20, and
40 mg PFOA/kg for 14 days. Liver samples were collected for determination of microsomal total
cytochrome P450 concentration and peroxisome P-oxidation activity. Total cytochrome P450
was significantly increased (p<0.05) at >20 mg PFOA/kg and P-oxidation activity was increased
at >2 mg PFOA/kg. The estimated ECsos for total cytochrome P450 and P-oxidation were 18.18
and 2.19 mg PFOA/kg, respectively. The LOAEL was 2 mg/kg based on increased liver weight,
serum estradiol, and hepatic aromatase activity, and the NOAEL was 0.2 mg/kg.
Hines et al. (2009) examined the roles that exposure to PFOA and ovarian hormones might
play in animals exposed during gestation compared to during their adult years. Timed-pregnant
CD-I mice were gavage-dosed in two blocks on GDs 1-17, but not thereafter. Block 1 animals
were dosed with 0, 1,3, and 5 mg PFOA/kg, and block 2 animals were dosed with 0, 0.01, 0.1,
0.3, 1, and 5 mg PFOA/kg/day. At birth, pups were pooled within each block and dose group and
randomly redistributed among the dams (10 pups per litter). Offspring were weaned at 3 weeks,
and a subset of females from each dose group (0, 0.01, 0.1, 0.3, 1, and 5 mg PFOA/kg/day) was
OVX at weaning or the day after weaning. All animals were observed until they reached 18
months of age.
Body weight was recorded weekly for the first 9 months of age, followed by monthly body
weight recordings over the next 9 months. As the animals matured, they were evaluated for the
endpoints listed in Table 3-37. A group of naive 8-week-old adult mice were dosed for 17 days
with 0, 1, and 5 mg PFOA/kg/day to compare the impact of exposure in adult animals to those
occurring during gestation. At 18 months of age, the mice were sacrificed. Blood, retroperitoneal
abdominal fat, interscapular brown fat, organs, and abnormal growths were collected at
necropsy.
Table 3-37. Data Collection for Female Mice Gestationally Exposed to PFOA
Test
Glucose tolerance test
Serum leptin and insulin
Body mass composition
Glucose tolerance test
Food consumption
Serum estradiol
Age at Test
15-16 weeks
21-33 weeks
42 weeks
17 months
17 months
18 months
Dose (mg/kg/day)
0,1,5
0,0.01,0.1,0.3, 1
0,0.01,0.1,0.3, 1
0,0.1, 1,5
0,0.1, 1,5
0,0.01,0.1,0.3, 1,5
Group
Intact
Intact, OVX
Intact
Intact
Intact
Intact
Source: Hines et al. 2009
Body weight of offspring born to dams exposed to 5 mg PFOA/kg was significantly
decreased (p<0.05) on PND 1 and through 18 months of age compared to control offspring body
weight. At weaning, the body weight of offspring born to dams exposed to 1 mg PFOA/kg/day
was significantly decreased (p<0.05) compared to control offspring body weight. A significant
increase (p<0.05) in body weight, due to more rapid weight gain after week 10, compared to
intact control body weight, was observed in intact mice exposed to 0.01-0.3 mg PFOA/kg/day
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during gestation. Body weight of intact mice gestationally exposed to 0.01-0.3 mg PFOA/kg/day
was comparable to body weight of control mice at 18 months.
Due to the increased weight gain observed in intact mice exposed to PFOA during gestation,
glucose tolerance tests were carried out along with determination of serum insulin concentration.
In cases of insulin resistance, plasma glucose and insulin levels are elevated and the insulin
response is lessened. Insulin resistance also has been associated with excess abdominal fat.
Serum leptin levels also were determined as increased leptin levels have been associated with a
leptin-resistance mechanism of action (MOA) for increased weight gain in humans. Body mass
composition was used to determine if there were differences in body fat between the intact
groups, and feed consumption was recorded to determine if consumption played a role in body
weight differences in intact control and intact gestationally exposed mice. Serum estradiol was
measured to determine if PFOA impacted hormone levels at 18 months in intact control and
intact gestationally exposed mice.
Glucose tolerance testing showed no statistically significant differences in baseline glucose
or response to glucose challenge at 15-16 weeks or at 17 months. At 21 and 31 weeks of age, a
significant increase in serum leptin and insulin levels was observed in intact mice exposed to
0.01 and 0.1 mg PFOA/kg/day. No statistically significant difference was observed between the
fat-to-lean ratio of intact control and intact gestationally exposed animals at 42 weeks of age. No
significant difference was observed in food consumption between intact control and intact
gestationally exposed animals at 42 weeks of age. Serum estradiol levels were not different
between intact control and intact gestationally exposed animals at 18 months.
Exposure to PFOA as an adult did not result in body weight differences among the groups at
18 months of age. The body weight of intact mice gestationally exposed to 1 mg PFOA/kg/day
was significantly increased (p<0.05) compared to adult mice exposed to 1 mg PFOA/kg/day. No
other differences in body weight among the groups were observed.
No significant differences among the groups were observed in survival during the 18-month
study. At necropsy, abdominal white fat was significantly decreased (p<0.05) at 1 and 5 mg
PFOA/kg/day in gestationally exposed intact mice compared to intact control mice. Interscapular
brown fat was significantly increased (p<0.05) at 1 and 3 mg PFOA/kg/day in gestationally
exposed intact mice and in gestationally exposed OVX mice at 1 mg PFOA/kg/day. Relative
spleen weight was significantly decreased (p<0.05) at 3 mg PFOA/kg/day in gestationally
exposed intact mice and at 1 and 5 mg PFOA/mg (p = 0.05-0.07) in gestationally exposed OVX
mice. Relative liver weight was not different between the groups. No differences were observed
at 18 months of age in tissue weight in mice exposed to PFOA as adults. At 1 mg PFOA/kg/day,
white and brown fat weight was significantly increased in gestationally exposed intact mice
compared to adult-exposed mice exposed to 1 mg PFOA/kg/day.
The authors concluded that developmental exposure to low doses and high doses of PFOA
resulted in different phenotypes in mice. At low doses, increased weight, increased serum
insulin, and increased serum leptin were observed in adult mice. At high doses the animals
displayed decreased weight in early and late life, decreased white fat, increased brown fat, and
decreased spleen weight. Under the conditions of the study, the developmental LOAEL was 0.01
mg PFOA/kg based on increased weight gain and increased serum insulin and leptin levels. No
developmental NOAEL was established. The adult NOAEL was 5 mg PFOA/kg, and no LOAEL
was established.
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Adrenal Hormones. Thottassery et al. (1992) exposed intact or ADX male Sprague-Dawley rats
to a single dose of 150 mg/kg PFOA in corn oil to determine the role of adrenal hormones on
liver enlargement and peroxisomal proliferation. ADX rats were dosed 2 days after surgery with
PFOA (ADX PFOA), CORT (ADX CORT), or both (ADX CORT PFOA). A group of intact and
ADX rats received only the vehicle and served as controls. The animals were sacrificed 48 hours
after dosing with PFOA or vehicle. Assays were conducted to determine DNA levels and
changes in peroxisomal p-oxidation, catalase, and ornithine decarboxylase activities. An increase
in ornithine decarboxylase activity has been associated with proliferation of many different cell
types. An increase of ornithine decarboxylase in the livers of animals exposed to PFOA would
suggest that the increased liver weight observed in PFOA-exposed animals was the result of
hyperplasia. Ornithine decarboxylase was determined by measuring liberated CO2 from DL-[1-
14C] ornithine hydrochloride in all animals except those in the ADX CORT PFOA group.
Relative liver weight in intact rats treated with PFOA was significantly increased compared
to control (36%, p<0.05). Relative liver weight in rats in the ADX PFOA group was significantly
increased compared to rats in the ADX vehicle group (16%, p<0.05). Relative liver weight in rats
in the ADX CORT PFOA group was significantly increased compared to rats in the ADX CORT
group (32%, p<0.05). Hepatic DNA levels were significantly decreased p<0.001) in intact rats
treated with PFOA and in rats in the ADX CORT PFOA group.
Ornithine decarboxylase activity was significantly increased in the rats in the ADX PFOA
group compared to rats in the ADX group (170.5 pmole CO2/hr/mg protein, versus 30.5 pmole
CO2/hr/mg protein, p<0.001), but no different between the intact rats treated with PFOA and the
intact rats treated with the vehicle.
PFOA increased whole liver peroxisomal P-oxidation activity by a similar amount and was
not different among the groups. In intact rats and rats in the ADX CORT PFOA group, exposure
to PFOA increased whole liver catalase activity, but exposure did not increase activity in the rats
in the ADX PFOA group. Based on the results, the authors concluded that adrenal hormones
were not required to induce peroxisomal P-oxidation activity in PFOA-exposed rats, but are
required to increase catalase activity. They also concluded that the enlarged livers of PFOA-
exposed animals were the result of hypertrophy rather than hyperplasia based on decreased
hepatic DNA content and lack of increased ornithine decarboxylase activity.
3.3.4 Physiological or Mechanistic Studies
Gene Expression. Rosen et al. (2007) examined the gene expression profile in the lung and liver
of mouse fetuses exposed to PFOA. Pregnant CD-I mice were gavage-dosed with 0, 1, 3, 5, and
10 mg PFOA/kg/day on GD 1-17. Dams were sacrificed on GD 18, and three fetuses per litter
were processed for total RNA from portions of the liver and lung. Global gene expression was
analyzed using Affymetrix gene chips.
A dose-related increase was observed in the number of genes altered by PFOA exposure in
both the liver and lung. A greater number of genes in the liver were altered compared the number
of genes altered in the lung. Analysis of the genes by canonical pathway or biological function
showed that most of the altered genes in both the liver and lung were associated with lipid
homeostasis. In the fetal lung, the two highest doses of PFOA altered genes associated with fatty
acid catabolism. In the fetal liver, all doses of PFOA were associated with genes involved in fatty
acid catabolism, lipid transport, cholesterol biosynthesis, bile acid biosynthesis, lipoprotein
metabolism, steroid metabolism, retinol metabolism, inflammation, phospholipid metabolism,
Perfluorooctanoic acid (PFOA) - May 2016 3-129
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glucose metabolism, proteosome activation, and ketogenesis. Although PPARa is known to at
least partly regulate the expression of genes for the pathways or biological functions involved in
lipid homeostasis, PFOA might independently activate other nuclear receptors, influencing the
metabolic responses observed.
Rosen et al. (2008a) described the gene profiles in liver tissue from wild-type 129Sl/SvlmJ
mice (7-8 per group) and PPARa-null mice (129S4/SvJae-PPARatmlGonz/J, 6-8 per group) dosed
for 7 days with 0, 1, and 3 mg PFOA/kg or 50 mg Wyeth 14,643, a PPARa agonist (Wolf et al.
2008a). RNA was isolated from the tissues and gene expression analyzed using Applied
Biosystems Mouse Genome Survey Microarrays. RT-PCR was used to evaluate selected genes.
In both wild-type and PPARa-null mice exposed to PFOA, the number of significant and
fully annotated genes used to evaluate the data for relevance to canonical pathway or biological
function were fewer at 1 mg/kg than at 3 mg/kg PFOA. However, 85% of the altered genes at
1 mg/kg PFOA also were altered at 3 mg/kg PFOA.
PPARa target genes including acyl-CoA oxidase 1 (Acoxl), Mel, Slc27al, Hsdl7b4, Hadha,
Hadhb, and Pdk4 were upregulated in PFOA- and Wyeth 14,643-treated wild-type mice, but not
in PPARa-null mice. Pdk4 was downregulated in PPARa-null mice exposed to PFOA but not in
PPARa-null mice exposed to Wyeth 14,643. Principal components analysis showed that genes
activated in PFOA-treated PPARa-null mice were similar to those in PFOA-treated wild-type
mice at 3 mg PFOA/kg, suggesting that many of the responses were not completely linked to
PPARa.
In wild-type PFOA- and Wyeth 14,643-treated mice, alterations were observed in genes
associated with fatty acid metabolism (mostly upregulated), inflammatory response (mostly
downregulated), cell cycle control (mostly upregulated), peroxisome biogenesis (mostly
upregulated), and proteasome structure and organization (mostly upregulated). In genes
associated with xenobiotic metabolism, the response was different between PFOA- and Wyeth
14,643-treated wild-type mice. Many of the Cyp2 genes were upregulated by PFOA and
downregulated by Wyeth 14,643. In PPARa-null PFOA-treated mice, genes associated with fatty
acid metabolism, inflammation, xenobiotic metabolism, and cell cycle control were altered in a
manner similar to the changes observed in PFOA-treated wild-type mice.
RT-PCR generally revealed good agreement with microarray analysis. However, expression
of Ehhadh, a PPARa-regulated gene, was upregulated in PFOA-treated wild-type mice but not in
PFOA-treated PPARa-null mice in microarray analysis. In contrast, expression of Ehhadh was
upregulated in all PFOA-treated mice in RT-PCR analysis. The authors concluded that PFOA
induces transcriptional changes mediated through PPARa activation, and it also alters gene
expression independently of PPARa. They noted that PFOA had multiple modes of action and
can function as a biologically active xenobiotic in the absence of PPARa.
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 12981/SvlmJ) and PPARa-null (strain 12984/SvJae) mice dosed
with 3 mg/kg/day PFOA for 7 days (from Wolf et al. 2008a);
male wild-type and PPARa-null mice (strain SV129/C57BL/6) gavage-dosed or fed diets
containing Wyeth 14,643 (PPARa agonist);
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female wild-type and CAR-null (strain C57BL/6xl29Sv) gavage-dosed with CAR
activators phenobarbital (PB) or l,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 PPARa-
null mice. A total of 117 genes were regulated similarly in both strains, and 29 upregulated genes
and 11 downregulated genes were unique to PPARa-null mice.
The gene expression profile of wild-type and PPARa-null mice exposed to PFOA for 7 days
or Wyeth 14,643 for 12 hours, or 3 or 7 days were compared. Four groups of altered genes were
identified based on their expression in wild-type and PPARa-null PFOA-exposed mice compared
to genes from Wyeth 14,643-treated mice. The first group consisted of genes (397) regulated by
both PFOA and Wyeth 14,643 in wild-type mice. They had a common direction and magnitude
of change and were characterized as being involved in lipid homeostasis, inflammation, cell
proliferation, or proteome maintenance genes. Group II consisted of genes in wild-type mice (51)
regulated solely by PFOA; most were involved in amino acid metabolism. Of the 81 genes
altered by PFOA exposure in PPARa-null mice (Group III), 62 had similar expression in wild-
type mice and many were involved in lipid metabolism. Regulation of these genes also was
observed in Wyeth 14,643 wild-type mice. Group IV genes (19) were altered by PFOA only in
PPARa-null mice; most were xenobiotic metabolizing enzymes.
By comparing the gene expression patterns between PFOA and Wyeth 14,643, the authors
concluded that:
PPARa is required for a majority of the transcriptional changes observed in the mouse
liver following PF also are regulated by other peroxisome proliferators in a PPARa-
dependent manner; and
PFOA impacts some PPARa-independent genes including ones involved in lipid
homeostasis (upregulated), amino acid metabolism (downregulated), and xenobiotic
metabolism (upregulated).
The transcription profiles of PFOA exposed wild-type and PPARa-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 PPARa-
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 PPARa-null mice. These results
suggest that some genes altered by PFOA exposure in PPARa-null mice are regulated by CAR
but not by Nrf2.
Bjork and Wallace (2009) examined the PPARa-dependent transcriptional activation
potential of PFOA in rodent and human hepatic liver cells. Primary rat and human hepatocytes
and HEPG2/C3A cells were incubated with 0, 5, 10, 20, 30, 50, 100, and 200 umol PFOA for
24 hours. Expression of Acox, Cyp4al (rat), Cyp4all (human), acyl-CoA thioesterase (Cte-rat,
Acotl-human), and DNA damage inducible transcript (Ddit3) were determined by quantitative
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RT-PCR. These genes are inducible by peroxisome proliferators, except Ddit3, which is induced
in the presence of direct or indirect DNA damage. Exposure to > 100 umol 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 umol PFOA, and Cte/Acotl was significantly induced
at >20 umol PFOA in rat hepatocytes only. Expression of Cyp4al/l 1 was significantly induced
in rat hepatocytes at 550 umol and in human hepatocytes at 20-50 umol. 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 PPARa transgenic mouse line (hPPARa). The study design and
whole animal toxicity data are described in section 3.2.2. Male 8-week-old wild-type (mPPARa)
mice, PPARa-null mice, and hPPARa mice were gavage-dosed with 0, 0.1, and 0.3 mg/kg PFOA
(n = 4-6 per group) for 2 weeks and sacrificed 18-20 hours following the last dose. Livers were
collected and analyzed for mRNA (RT-QPCR) and protein levels (western blot analysis) of
PPARa and related genes (retinoid X receptor alpha [RXRa], peroxisomal bifunctional protein
[PH], peroxisomal thiolase [PT], very long chain acyl-CoA dehydrogenase [VLCAD], medium
chain acyl-CoA dehydrogenase [MCAD], and cytochrome P450 4alO [CYP4A10]). RXRa forms
a heterodimer with PPARa to control transcription of genes affecting lipid metabolism.
CYP4A10 also plays a role in lipid metabolism. Treatment with peroxisome proliferators caused
an increase in both PH and PT. MCAD and VLCAD are mitochondrial fatty acid metabolizing
enzymes whose gene expression is mediated by PPARa (Aoyama et al. 1998). The results of
mRNA expression impacted by PFOA exposure are shown in Table 3-38.
Treatment with PFOA did not alter mRNA expression or protein expression of PPARa,
RXRa, or MCAD in mPPARa mice. At 0.1 mg/kg PFOA, mRNA expression of CYP4A10 was
significantly increased (p<0.05) in mPPARa mice compared to control mPPARa 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 mPPARa mice when
compared to control mPPARa mice.
Table 3-38. mRNA Expression of Hepatic PPARa and Related Genes
PPARa
RXRa
PH
PT
VLCAD
MCAD
CYP4A10
mPPARa
0 mg/kg
-
-
-
-
-
-
-
0.1
mg/kg
-
-
-
-
-
-
+t
0.3
mg/kg
-
-
+t
+t
+t
-
+t
PPARa-null
0 mg/kg
NA
-
-
-
-n
-
-
0.1
mg/kg
NA
-
-n
-n
-n
-n
-n
0.3
mg/kg
NA
-
-n
-n
-n
-
-n
hPPARa
0 mg/kg
-*t
-
-
-*t
-
-
-*t
0.1
mg/kg
-*t
-
-n
-
-n
-n
-n
0.3
mg/kg
-*t
-
-n
-
-n
+t
-n
Source: Nakamura et al. 2009
Notes:
- Not different from respective control.
+ Significantly different from respective control.
* Significantly different from mPPARa mice treated with same concentration.
I Decreased expression relative to respective control or mPPARa mice at same concentration.
t Increased expression relative to respective control or mPPARa mice at same concentration.
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As expected, mRNA and protein expression of PPARa was absent in PPARa-null mice.
Treatment with 0.1 and 0.3 mg/kg PFOA did not alter mRNA or protein expression for any genes
investigated compared to control PPARa-null mice. VLCAD mRNA expression and PT protein
expression in control PPARa-null mice was significantly decreased (p<0.05) compared to
mPPARa control mice. VLCAD mRNA and protein expression of PFOA treated PPARa-null
mice was significantly decreased (p<0.05) compared to mPPARa mice treated with the same
doses. Following treatment with 0.1 mg/kg PFOA, MCAD mRNA expression was decreased
(p<0.05) compared to mPPARa mice treated with the same dose. When compared to mPPARa
mice treated with the same dose, mRNA and protein expression of PH and PT was significantly
decreased (p<0.05) in PPARa-null mice, as was CYP4A10 mRNA expression.
Treatment with PFOA did not alter mRNA or protein expression of PPARa, RXRa, PH, PT,
or VLCAD in hPPARa mice compared with their respective controls. Expression of CYP4A10
mRNA also was not altered by PFOA treatment. MCAD mRNA and protein expression were
significantly increased (p<0.5) in hPPARa mice treated with 0.3 mg/kg PFOA compared to
hPPARa control mice. Expression of PPARa mRNA and protein levels were significantly higher
(p<0.05) in all hPPARa mice than in mPPARa mice given the same concentration of PFOA.
Treatment of hPPARa mice with 0.1 and 0.3 mg/kg PFOA caused a decrease (p<0.05) in mRNA
expression of PH, VLCAD, and CYP4A10 compared to mPPARa mice at the same dose. Only
hPPARa mice treated with 0.3 mg/kg PFOA had decreased protein expression of PH and
VLCAD compared to mPPARa mice given the same treatment. An important finding from this
study was the significant downregulation of some genes in PPARa-null and hPPARa mice that
are significantly upregulated by PPARa in the control animals. In the animals with the
humanized PPARa gene or no PPARa gene, there was a response, but the response was the
opposite of what occurred with normal mouse PPARa activation. In the null and humanized
mice, the significantly decreased alterations in gene expression occurred at 0.1 mg/kg/day; this
dose level had no change in expression for all but one gene in the normal mice and increased
expression, rather than decreased expression, at 0.3 mg/kg/day (see Table 3-38).
Treatment with 0.3 mg/kg PFOA caused activation of PPARa in mouse, but not in
humanized PPARa mice. The results suggest that the functional activation of human PPARa
could be weaker than that of mice as expression of human PPARa in mice was greater than the
expression of mouse PPARa. Higher concentrations of PFOA might be needed to cause
activation of human PPARa in hPPARa mice.
To further evaluate the developmental effects potentially mediated by PPARa, groups of
female wild-type, PPARa-null, and PPARa-humanized mice were given 0 and 3 mg PFOA/kg
on GDs 1-17 by oral gavage (Albrecht et al. 2013). The study design and developmental toxicity
data are described in section 3.2.5. Females were either sacrificed on GD 18 (n = 5-8 per group)
or allowed to give birth and then sacrificed, along with their litters (n = 8-14), on PND 20.
Livers from dams, fetuses, and pups were collected for measurement of mRNAs encoding the
PPARa target genes Cyp4alO and Acoxl, the CAR target gene (Cyp2blO), and the PXR target
gene (Cyp3all).
On GD 18, maternal liver samples from treated groups showed increased expression of
Acoxl in wild-type mice and Cyp4alO in wild-type and humanized mice. Expression of
Cyp2blO and Cyp3al 1 were increased following PFOA administration in all three genotypes. On
PND 20, maternal liver samples from treated groups showed increased expression of Acoxl in
wild-type mice; expression of Cyp2blO was unchanged in all groups; and expression of Cyp3al 1
was increased following PFOA administration in all three genotypes.
Perfluorooctanoic acid (PFOA) - May 2016 3-133
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For fetuses on GD 18, liver samples from treated groups showed increased expression of
Acoxl and Cyp4alO in wild-type and humanized mice. Expression of Cyp2blO was unchanged
following maternal PFOA administration in all three genotypes, while expression of CypSal 1
was increased in humanized fetal liver. On PND 20, pup liver samples from treated dams showed
increased expression of Acoxl and Cyp4alO in wild-type mice; expression of Cyp2blO was
increased in all genotypes; and expression of CypSal 1 was increased following maternal PFOA
administration in wild-type and humanized pups. Thus, expression of PPARa target genes that
modulate lipid metabolism was increased in both wild-type and humanized mice coincident with
increased liver weight and microscopic lesions; however, the neonatal mortality was observed
only in wild-type offspring (Albrecht et al. 2013).
Walters et al. (2009) examined the impact of PFOA on mitochondrial biogenesis and gene
transcription in adult male Sprague-Dawley rats orally dosed with 0 or 30 mg/kg PFOA for
28 days. At sacrifice, a portion of the midlobe region of the livers was collected. Liver DNA and
RNA were isolated for RT-PCR of genes in the peroxisome proliferator-activated receptor
gamma coactivator la- (Pgc-la-) mediated pathway of mitochondrial biogenesis: Pgc-la,
estrogen-related receptor a (Erra), nuclear respiratory factor 1 (Nrfl) and Nrf2, transcription
factor A (Tfam), cytochrome c oxidase subunit II and IV (Cox II and Cox IV), NADH
dehydrogenase 2 (Nd2), and NADH dehydrogenase iron-sulfur protein 8 (NdufsS). In
mitochondrial biogenesis, Pgc-la and Erra increase expression of the transcription factors Nrfl
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 NdufsS) 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-la, 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-la, Erra, Nrfl, Nrf2, and
Tfam was significantly increased 1.3-2.2-fold (p<0.05), and expression of Cox II, Cox IV, Nd2,
and NdufsS was significantly increased 2-9-fold (p<0.05) compared to controls. Protein
expression of Pgc-la was increased, and expression of Cox II and Cox IV were decreased in
PFOA-treated rats. Protein expression of Tfam was not affected by treatment with PFOA. The
results suggested that PFOA induced mitochondrial biogenesis at the transcriptional level by
activation of the Pgc-la pathway, confirming the potential for effects on mitochondria but not
clarifying whether those effects are in some way linked to PPARa activation.
Elcombe et al. (2010) examined the expression of some cytochrome P450 isoforms in the
livers of male Sprague-Dawley rats fed diets containing 300 ppm PFOA or 50 ppm Wyeth
14,643 for 1, 7, or 28 days. The isoforms included those involved in activation of PPARa
(CYP4A1), CAR (CYP2B1/2), and PXR (CYP3A1). All three isoforms were induced by PFOA
exposure. CYP2B1/2 and CYP4A1 were induced after 1 day of exposure to PFOA. CYP3A1 was
induced in all PFOA-exposed rats after 7 days of exposure. Treatment with Wyeth 14,643 caused
the induction of CYP4A1 only.
PPAR Activation. Takacs and Abbott (2007) evaluated the potential for PFOA to activate
PPARs, using a transient transfection cell assay. Cos-1 cells, derived from the kidney cells of the
African green monkey, were transfected with mouse or human PPARa, PPAR.p/5, or PPARy
reporter plasmids and exposed to 0.5-100 |imol PFOA or 0.5-100 |imol PFOA and MK-886
(PPARa antagonist) or GW9662 (PPARy antagonist). An antagonist for PPARp/6 was not
available. The three types, PPARa, P/5, and y, are encoded by different genes, expressed in many
Perfluorooctanoic acid (PFOA) - May 2016 3-134
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tissues, and have specific roles during development as well as in the adult. The results are shown
in Table 3-39. PFOA activated PPARa in a dose-dependent manner with a significant increase in
activity observed at 10, 20, 30, and 40 jimol for the mouse receptor and 30 and 40 jimol for the
human receptor compared to the negative control. The presence of the PPARa antagonist
MK-886 prevented the activity increase resulting from PFOA exposure alone in mouse and
human PPARa constructs.
Table 3-39. Activation of Mouse and Human PPAR by PFOA
PPARa
PFOA
(urn)
0
0.5
1
o
J
5
10
15
20
30
40
Mouse
-
-
-
-
-
+
-
+
+
+
Human
-
-
-
-
-
-
-
-
+
+
PPARP/5
PFOA
(urn)
0
10
15
20
30
40
50
60
70
80
Mouse
-
-
-
-
-
+
+
+
+
+
Human
-
-
-
-
-
-
-
-
-
-
PPARy
PFOA
(urn)
0
1
5
10
20
30
40
50
75
100
Mouse
-
-
-
-
-
-
-
-
-
-
Human
-
-
-
-
-
-
-
-
-
-
Source: Takacs and Abbott 2007
Notes:
+ Significant increase in activity between treated and control.
- No difference in activity between treated and control.
Activity of mouse PPAR.p/5 was significantly increased after exposure to 40-80 jimol PFOA
compared to the negative control. Activity of human PPAR.p/5 was not increased by PFOA
exposure. Activity of mouse and human PPARy were not increased by exposure to PFOA. PFOA
was found to activate mouse and human PPARa and mouse PPAR 0/5 under the conditions in
this study.
Biomarkers for Peroxisome Proliferation. Pastoor et al. (1987) dosed male Crl:CD (SD) BR
rats for 1, 3, and 7 days with 0 or 50 mg PFOA/kg/day. Hepatic DNA content, cytochrome P450
content, UDP-glucuronyltransferase, glutathione S-transferase, benzphetamine N-demethylase
activity (marker for smooth ER proliferation), and ethoxyresorufm O-deethylase activity (marker
for cytochrome P450 induction via the aryl hydrocarbon receptor) were measured from rats
dosed 1 and 3 days. Liver microsomes were prepared from rats dosed for 3 days for CAT and
CPT activity assays. CAT served as a marker for peroxisome proliferation and CPT was a
marker for mitochondrial proliferation. Incorporation of [14C]acetate into hepatic lipids was used
to determine the effect of PFOA on hepatic lipid metabolism. Plasma TC and triacylglycerides
was determined from rats dosed for 7 days.
Hepatic DNA content was not increased in treated rats when compared to content in control
rats. Cytochrome P450 was significantly increased (p<0.05) and ethoxyresorufm O-deethylase
activity was significantly decreased (p<0.05) after treatment for 1 and 3 days. Benzphetamine
N-demethylase activity was significantly increased (p<0.05) after treatment with PFOA for
3 days. CAT activity increased 12-fold (p<0.05) and CPT increased twofold (p<0.05) after a
3-day treatment with 50 mg PFOA/kg. No differences were observed among the groups for the
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other enzymes. No differences were observed between rats treated for 7 days and control rats in
plasma TC or triacylglycerol. Although a significant increase (p<0.05) was observed for
[14C]acetate incorporation into triacylglycerols, cholesteryl esters, and polar lipids, there was no
difference in the distribution of the incorporated label between control and treated rats. The
authors concluded that the lack of increased DNA content, proliferation of smooth ER, and
peroxisome proliferation pointed to increased liver weight due to hepatocyte hypertrophy.
Gap Junction Intercellular Communication. Upham et al. (1998, 2009) examined the effects
of perfluorinated fatty acids on gap junction intercellular communication (GJIC) in male Fischer
344 rats fed diets containing 0 or 0.02% PFOA (intake 37.9 mg/kg/day) for 1 week and in
WB-F344 rat liver epithelial cells. The chain lengths of the perfluorinated fatty acids ranged
from 2-10, 16, and 18 carbons. Liver weight in the rats fed diets containing 0.02% PFOA was
significantly increased compared to control rat liver weight. No differences were observed in
serum AST, ALT, and ALP. PFOA significantly inhibited GJIC in the livers of rats after
treatment for 1 week. In WB-F344 cells, GJIC was inhibited by perfluorinated fatty acids with
7-10 carbons within 15 mins of incubation. The inhibition was reversible with full recovery
occurring within 30 mins of PFOA removal from media. Extracellular receptor kinase was
activated by PFOA within 5 mins of incubation in the cells. Preincubation of cells with the
phosphatidylcholine-specific phospholipase C inhibitor D609 partially prevented GJIC inhibition
by PFOA. The authors concluded that PFOA, having an 8-carbon chain, inhibited GJIC by
activation of extracellular receptor kinase and phosphatidylcholine-specific phospholipase C, but
noted that other mechanisms might be involved.
Production of ROS. Takagi et al. (1991) fed male Fischer 344 rats diets containing 0, 10, and
20 mg PFOA/kg for 2 weeks to determine the formation of 8-OH-dG (marker of oxidative DNA
damage). Livers and kidneys were removed at necropsy and DNA was isolated from each organ
and analyzed. The relative liver and kidney weights were significantly increased (p<0.05) in the
treated rats compared to the control. A significant increase in 8-OH-dG liver levels was observed
at > 10 mg PFOA/kg. There were no significant differences in 8-OH-dG kidney levels between
PFOA-treated and control rats. The authors concluded that PFOA could cause organ-specific
oxidative DNA damage.
Hu and Hu (2009) exposed human hepatoma cells, HepG2, to PFOA to evaluate cytotoxic
effects. Cells also were exposed to a mixture of PFOA and PFOS to determine antagonistic or
synergic effects. The cells were exposed to 0, 50, 100, 150, and 200 |imol PFOA or to 0, 50, 150,
and 200 jimol each of PFOA and PFOS. A group of cells also were exposed to 0, 50, 100, 150,
and 200 jimol PFOS. The cells were cultured for 24, 48, and 72 hours. Cell viability, apoptosis,
ROS, mitochondrial membrane potential, antioxidant enzymes, glutathione content, and
differential expression of apoptosis gene regulators p53, Bax, Bcl-2, caspace-3, and caspace-9
genes were evaluated.
Exposure to PFOA or PFOS caused a dose-dependent decrease in viability of HepG2 cells. A
nonsignificant dose-dependent increase in apoptosis was observed in the cells cultured with
PFOA. However, the combination of PFOA and PFOS showed a significant dose-dependent
increase (p<0.05) in apoptosis. Intracellular ROS were significantly increased (p<0.05) in cells
cultured with 100, 150, and 200 |imol PFOA or PFOS. HepG2 cells exposed to the mixture of
100 and 200 jimol 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 jimol PFOA significantly decreased (p<0.05)
glutathione concentration and glutathione peroxidase activity; and 150 jimol PFOA significantly
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increased (p<0.05) the activities of SOD, catalase, and glutathione reductase, and significantly
decreased (p<0.05) glutathione peroxidase activity, and glutathione concentration in HepG2
cells. The trend was the same at 200 jimol PFOA, with the exception of GST activity being
significantly decreased (p<0.05).
Exposure to PFOA did not change p53, Bax, or caspace-3 expression in HepG2 cells.
Expression of Bcl-2 was downregulated and caspace-9 was upregulated in a dose-dependent
manner in HepG2 cells following exposure to 50-200 u/Mol PFOA. The authors proposed that
PFOA and PFOS induced cell apoptosis by overwhelming the homeostasis of antioxidative
systems, increasing ROS, impacting mitochondria, and changing gene expression of apoptosis
gene regulators.
Eriksen et al. (2010) examined ability of PFOA to generate ROS and induce oxidative DNA
damage in human HepG2 cells. Cell were incubated with 0, 0.4, 4, 40, 200, 400, 1,000, and
2,000 umol PFOA and 2',7'-dichlorofluorescein diacetate. Hydrogen peroxide, FhO2, was used
as a positive control. A fluorescence spectrophotometer was used to measure ROS production
every 15 mins during the 3-hour incubation period in all cultures. The comet assay was used to
measure DNA damage in cells exposed to 0, 100, and 400 umol PFOA for 24 hours.
Cytotoxicity was determined by measuring the level of lactate dehydrogenase activity in the cell
medium. Exposure to PFOA caused a dose-independent increase (all doses p<0.05) in ROS
production in HepG2 cells. Compared to ROS production in negative control cells, PFOA
induced a 1.52-fold increase in production. There was no difference in oxidative DNA damage
and lactase dehydrogenase activity between PFOA-treated cells and negative control cells. The
authors concluded that oxidative stress and DNA damage were probably not relevant to potential
adverse effects of PFOA.
Protein Binding. The ability of PFOA to bind to serum proteins for distribution is discussed in
section 2.2. PFAS protein binding also can impact cellular function in cases in which the proteins
in question are transporters (serum albumin and fatty acid binding protein) or enzymes (lysine
decarboxylase) as well as membrane receptors (e.g., members of the PPAR family) and thyroid
hormone receptors. The mechanistic studies of the nuclear PPARa membrane receptors are
described in section 3.3.4.
Ren et al. (2015) examined the relative binding affinities of 16 PFASs for the human thyroid
hormone receptor's a ligand binding domain (TRa-LBD) using a fluorescence competitive
binding assay. Solutions of 1 umol TRa-LBD were prepared in DMSO. Changes in TRa-LBD
tryptophan fluorescence after binding to 10-umol T3 in the absence or presence of the PFAS was
used to determine the binding properties of the PFAS. ICso values were calculated by linear
extrapolation between two responses located in the vicinity of a 50% inhibition level. All the
PFASs had a lower affinity for the receptor than T3. Affinity of PFOA was less than that for
PFDA, PFUnA, PFNA, and PFOS.
ToxCast Assay Results. The Toxicity Forecaster (ToxCast) database is a large high-throughput
screening compilation of public in vitro and in vivo assays on over 9,000 chemicals (USEPA
2015). PFOA was tested in 1,084 assays and was active in 40 (USEPA 2015). Assays with less
than 50% efficacy reported or overfitting issues are not included in the summary of results that
follows.
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Three of the acceptable ToxCast active cytotoxicity assays evaluated the impact of PFOA.
All three of these assays are based on one cell type. If there was no cytotoxicity reported for a
specific cell type, the AC50 (the minimum concentration with 50% cytotoxicity activity) was
used for comparison when reporting the ToxCast results. The lowest recorded AC50 (109 umol)
measured the degradation of microtubules in liver cells at 109 umol and the highest recorded
(123 umol) measured general cytotoxicity in liver cells.
PFOA activated two of the 21 estrogen related assays in ToxCast; both were ESR1-related.
Estrogen and its receptors are essential for sexual development and reproductive function, but
also play a role in other tissues such as bone. PFOA induced estrogen response element and
inhibited ESR1 at concentrations lower than their AC50 values with concentrations of 33.8 umol
and 47.4 umol, respectively. This implies that PFOA could have some estrogenic potential;
however, due to the small fraction of estrogenic assays activated (10%), any activity is likely
weak.
PFOA activated PPARs, PXR, CAR, and retinoic acid receptor (RAR) assays within the
ToxCast program. From the PPAR assays, PFOA induced the DNA sequences for PPARa,
PPARy, and the peroxisome proliferator hormone response element (PPRE) and antagonized the
PPARy receptor. The only PPAR assay AC50 value that was above the cell-specific AC50 was
that for PPARy antagonism at 5.91 umol. However, it is possible that cytotoxicity occurs due to
PPAR induction, or that PPAR cytotoxicity leads to PPAR induction confounding interpretation
of the outcome. PFOA induced DNA sequences for PXR (AC50 9.42 umol) at a concentration
lower than the cell-specific AC50. CAR and RAR alpha antagonism also was observed, but the
concentrations of 17.57 umol and 28.45 umol, respectively, were not below the cell-specific
cytotoxicity value. PPAR, PXR, CAR, and RAR pathways are all nuclear receptors that can form
heterodimers with one another to induce translation of various genes. Some of these genes are
important for development, reproduction, and waste degradation, and could play a role in PFOA-
induced cancer.
The ToxCast program examined Cytochrome P450 (CYP) activation associated with PFOA
exposure. Although PFOA is not metabolically active, it was found to activate four CYPs:
CYP2C18, CYP2C19, and CYP2C9 in human cells and CYP2C11 in rat cells. All of the CYP
assays were activated at concentrations lower than the lowest AC50 (109 umol) but lacked cell-
specific ACSOs. The CYP2C class of CYPs is involved in the metabolism of xenobiotics such as
the following drugs: the antiseizure medication diazepam, beta blocker propranolol, and selective
serotonergic reuptake inhibitor citalopram. Though there is no evidence of metabolism of PFOA
by these CYPs, it is possible that it acts as a competitive or allosteric inhibitor for known
substrates of the CYPs activated. This coupled with PFOA's high affinity for binding to albumin
could significantly alter the PKs of various pharmaceutical bound to serum albumin, thus
potentially playing a role in increasing systemic toxicity of some pharmaceuticals by increasing
the free serum concentration.
PFOA failed to cause toxicity in the in vivo fish model for neurological and developmental
toxicity. This is important because PFOA induces developmental toxicity in mice and rats
in vivo.
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3.3.5 Structure-Activity Relationship
Bjork and Wallace (2009) compared the PPARa-dependent transcript!onal activation
potential of linear perfluorocarboxylic and sulfonic acids in rodent and human hepatic liver cells.
The PFAAs tested included perfluorinated carboxylic acids with carbon chain lengths of 2-8 and
perfluorinated sulfonic acids with chain lengths of 4-8. Primary rat and human hepatocytes and
HEPG2/C3 A cells were incubated with 0 and 25 umol perfluorinated compounds for 24 hours.
Expression of Acox, Cyp4al (rat), Cyp4all (human), Cte/Acotl, andDditS (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/Acotl only in rat
hepatocytes, and the degree of stimulation of gene expression increased with increasing carbon
number. The Cyp4al 1 gene was not expressed or stimulated by any of the PFAAs in
HepG2/C3 A 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 suggest that the PPARa-related
changes in gene expression induced by perfluorinated compounds in primary rat hepatocytes are
directly related to the carbon chain length and appear to be stronger for the carboxylic acids
(i.e., PFOA) than the sulfonates (i.e., PFOS). There was no induction in expression of Acox and
Cte/Acot 1 in either primary or transformed human liver cells in culture. The authors suggested
that the PPARa mediated peroxisome proliferation observed in rodent liver might not be relevant
as an indicator to human risk.
Wolf et al. (2008b) tested PFAAs, including PFOA, to determine if mouse and human
PPARa activity could be induced in transiently transfected COS-1 cell assays. COS-1 cells were
transfected with either a mouse or human PPAR-a receptor-luciferase reporter plasmid and, after
24 hours, were exposed to either negative controls (water or 0.1% DMSO), a positive control
(Wyeth 14,643), or PFOA at 0.5-100 jimol. Other concentrations of PFAAs were used but not
provided in this report. At the end of 24 hours of exposure, the luciferase activity was measured.
The positive and negative controls had the expected results. A lowest observed effect
concentration (LOEC) and no observed effect concentration (NOEC) were determined for each
PFAA. In the study, the mouse PPARa was more responsive than the human. Also, carboxylates
induced higher mouse and human PPARa activity than the sulfonates. In this study, the NOEC
for PFOA was 0.5 jimol in the mouse and 5 jimol in humans; the LOEC was 1 jimol
(0.43 |ig/mL) in the mouse and 10 jimol (4.3 |ig/mL) in humans.
A similar study included additional PFAAs (Wolf et al. 2012). Transfected cells were
incubated with PFAAs at concentrations of 0.5 to 100 jimol, vehicle (water or 0.1% DMSO as
negative control) or with 10 jimol Wyeth 14,643 (positive control) on each plate. Assays were
performed with three identical plates per compound per species with nine concentrations per
plate and eight wells per concentration. Cell viability was assessed using the Cell Titer Blue cell
viability kit and read in a fluorometer. The positive and negative controls had the expected
results. All cells transfected with either human and mouse PPARa responded to the PFAAs.
Again, the carboxylates were stronger inducers than the sulfonates, and the mouse PPARa was
more reactive than the human PPARa. The study also provided the C20max values for each PFAA
(the concentration at which the PFAA produced 20% of the maximal response elicited by the
most active PFAA). For PFOA, this was 6 jimol in mouse PPARa and 7 jimol in human PPARa.
For comparison, PFOS was 94 jimol and 262 jimol, respectively.
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3.4 Hazard Characterization
3.4.1 Synthesis and Evaluation of Major Noncancer Effects
Serum Lipids. Because of the structural similarities between linear perfluorinated acids and the
short- and medium-chain fatty acids, the potential for these chemicals to cause elevated serum
lipids has been an area of considerable interest. High levels of serum lipids (TC and LDL) are
risk factors for cardiovascular disease in humans, including IHD, a condition in which blood
flow to the heart is decreased through the development of atherosclerotic plaque or clots in the
cardiac arteries.
The association between PFOA and serum lipids has been examined in several studies in
different populations. Cross-sectional and longitudinal studies in occupational settings (Costa et
al. 2009; Olsen et al. 2000, 2003; Olsen and Zobel 2007; Sakr et al. 2007a, 2007b; Steenland et
al. 2015) and in the high-exposure community (the C8 Health Project study population) (Fitz-
Simon et al. 2013; Frisbee et al. 2010; Steenland et al. 2009; Winquist and Steenland 2014a)
generally observed positive associations between serum PFOA and TC in adults and children
(aged l-< 18 yrs); most of these effect estimates were statistically significant. Although
exceptions to this pattern are present (i.e., some of the analyses examining incidence of self-
reported high cholesterol based on medication use in Winquist and Steenland 2014a and in
Steenland et al. 2015), the results are relatively consistent and robust. Similar associations were
seen in analyses of LDL, but were not seen with HDL. The range of exposure in occupational
studies is large (means varying between 0.4 and > 12 |ig/mL), and the mean serum levels in the
C8 population studies were around 0.08 jig/mL. Positive associations between serum PFOA and
TC (i.e., increasing lipid level with increasing PFOA) were observed in most of the general
population studies at mean exposure levels of 0.002-0.007 |ig/mL (Eriksen et al. 2013; Fisher et
al. 2013; Geiger et al. 2014a; Nelson et al. 2010; Starling et al. 2014). The interpretation of these
general population results is limited, however, by the moderately strong correlations (Spearman r
> 0.6) and similarity in results seen for PFOS and PFOA. The most recent update of disease
incidence in workers in the C8 Health Project study population identified 35 cases of
nonhepatitis liver disease (with medical validation) (Steenland et al. 2015); no association was
seen with cumulative exposure when analyzed without a lag (HRs by quartile 1.0, 0.58, 1.43,
1.20; trend p = 0.86 for log cumulative exposure), but there was a possible trend in the analysis
using a 10-year lag (HRs by quartile 1.0, 1.46, 2.13, and 2.02; trend p = 0.40).
Cholesterol and/or triglycerides were monitored in only a few of the animal studies, which
did not all measure concurrent serum PFOA levels. Information on serum lipids from animal
studies has received less attention than in the human population because of the fact that
decreases in triglycerides, cholesterol, and lipoprotein complexes are an expected consequence
of PPARa activation in rodents. The PPARa response in animals tends to lower rather than raise
serum cholesterol and associated lipid levels. Peroxisomes are subcellular organelles that
increase beta oxidation of long-chain fatty acids using a beta oxidation pathway that is not linked
to adenosine triphosphate (ATP) production and release the shortened fatty acids to the cytosol
as an endproduct for export in VLDLs or hepatic ATP-production via mitochondrial beta
oxidation (Garrett and Grisham 1999). PPARa activation also stimulates metabolic changes that
lower hepatic cholesterol. The effects of human PPARa activation are much less pronounced
than those in rats and mice.
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Nakamura et al. (2009) and Minata et al. (2010) examined the lipid endpoints relative to the
mouse strain's PPARa status and PFOA exposure. Nakamura et al. (2009) found that mice with a
normal PPARa 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 PPARa receptor
or PPARa-null mice (NOAEL= 0.3 mg/kg/day) and their respective controls. The study by
Minata et al. (2010) used higher doses than Nakamura et al. (2009) and found that TC was
significantly decreased (LOAEL= 10.8 mg/kg/day; whole blood 47 jig/mL) and total
triglycerides significantly increased (LOAEL= 5.4 mg/kg/day; whole blood 21 |ig/mL) in wild-
type mice. In the PPARa-null mice, the TC was significantly decreased for the 5.4- and
10.8-mg/kg/day doses but significantly increased for a 21.6-mg/kg/day dose while total
triglycerides were significantly increased at all doses; these doses corresponded to whole blood
PFOA levels of 13, 36, and 71 |ig/mL, respectively. Rosen et al. (2007) found that PFOA
activated genes for fatty acid catabolism, cholesterol biosynthesis; bile acid biosynthesis; and
lipoprotein, steroid, and glucose metabolism in fetal livers. When comparing the response in
PPARa wild-type to null mice (Rosen et al. 2008b), 62 of 81 activated genes were the same for
both strains and were ones involved with lipid metabolism.
Martin et al. (2007) identified a 45-72% decrease in serum cholesterol after treatment of
male Sprague-Dawley rats with 20 mg PFOA/kg/day for up to 5 days (serum PFOA 245 |ig/mL
after 3 days), and Loveless et al. (2008) reported decreased TC, HDL, and non-HDL in male CD
rats after doses of 0.3 and 1 mg/kg/day for 28 days. Triglycerides were decreased in the rats at
>0.3 mg/kg/day. De Witt et al. (2009) found a dose-dependent decrease in triglyceride levels in
female C57BL/6N mice exposed to 0, 7.5, and 15 mg PFOA/kg bw in drinking water for
10 days. In male CD-I mice, TC, HDL, and triglycerides were decreased at 10 and 30 mg/kg/day
(Loveless et al. 2008). In pregnant female ICR mice, triglyceride, TC, and free fatty acids were
significantly decreased at 10 mg/kg (Yahia et al. 2010). Elcombe et al. (2010) found a significant
decrease in cholesterol in male Sprague-Dawley rats following a 7- or 28-day exposure to
300 ppm PFOA in the diet with a resulting serum level of 252 |ig/mL at 28 days. Accordingly,
there is not a high degree of concordance between the lipidemic effects of PFOA as noted in
human epidemiology studies and those seen in animals.
Filgo et al. (2015) found a trend for increased liver Ito (fat) cell atrophy and lesion severity
across the doses in CD-I and SV-129 mice at 18 months. PFOA exposure occurred only through
the dam during gestation and lactation in this study. This observation suggests that liver steatosis
could be a concern late in life for animals exposed to PFOA gestationally and during their early
postnatal period. However, the 18-month fat accumulation could also be related to normal aging
and/or dietary fat intakes across the animal's lifetime (Quist et al. 2015). Tan et al. (2013) found
that the fat content of the diet was an important variable in determining the impact of PFOA
(5 mg/kg/day) on liver and serum lipids. Intake of an HFD plus PFOA increased liver
triglycerides and serum free fatty acids compared to an RFD plus PFOA but had no impact on
liver cholesterol concentrations. Serum cholesterol was not monitored.
Hepatic Effects. Both the human and animal studies suggest effects on the liver as indicated by
increases in liver enzymes. The results of the occupational studies provide evidence of an
association with increases in serum AST, ALT and GGT, with the most consistent results seen
for ALT. The associations were not large and could depend on the covariates in the models, such
as BMI, use of lipid-lowering medications, and triglycerides (Costa et al. 2009; Olsen et al. 2000,
2003; Olsen and Zobel 2007; Sakr et al. 2007a, 2007b). Two population-based studies of highly
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exposed residents in contaminated regions near a fluorochemical industry in West Virginia have
evaluated associations with liver enzymes, and the larger of the two studies reported associations
of increasing serum In ALT and In GGT levels with increasing serum PFOA concentrations
(Emmett et al. 2006; Gallo et al. 2012). A cross-sectional analysis of data from NHANES,
representative of the U.S. national population, also found associations with In PFOA
concentration with increasing serum ALT and In GGT levels. Serum bilirubin was inversely
associated with serum PFOA in the occupational studies. A U-shaped exposure-response pattern
for serum bilirubin was observed among the participants in the C8 Health Project, which might
explain the inverse associations reported for occupational cohorts. Overall, an association of
serum PFOA concentration with elevations in serum levels of ALT and GGT has been
consistently observed in occupational, highly exposed residential communities, and the U.S.
general population. The associations are not large in magnitude, but indicate the potential to
affect liver function.
The data from animal studies for increases in ALT and AST support the findings in human
epidemiology studies; however, the animal studies for both aminotransferases lacked serum
PFOA measurements for comparison with the human serum data. Concurrent with the evidence
in animals of damage to liver cells, levels of some membrane transport proteins were altered. In
mice, the increased expression of MRP3 and MRP4 (Maher et al. 2008) and the decreased
expression of OATPs (Cheng and Klaassen 2008) favor excretion of PFOA into the bile.
Competition of PFOA with bile acids for transport could alter the excretion of the cholesterol
derivatives excreted in bile.
In animal studies, serum levels of ALT and/or AST were significantly increased indicating
apoptosis or necrosis of liver cells (Butenhoff et al. 2012; Minata et al. 2010; Son et al. 2008).
Increased levels of ALT were observed at a LOAEL of 2.65 mg/kg/day in ICR mice by Son et al.
(2008). Yahia et al. (2010) reported significantly increased ALT, GGT, AST, and ALP in PFOA-
exposed (10 mg/kg) pregnant ICR mice. Total protein, albumin, and globulin were significantly
decreased in the same mice.
No evidence of liver damage has been found in the human epidemiology studies with the
exception of the few enzyme changes discussed above. In most PFOA animal studies
(e.g., monkeys, rats, and mice), short-term and chronic exposure caused a dose-related increase
in liver weight as at least one of the co-occurring effects (Butenhoff et al. 2002, 2004a, 2012;
DeWitt et al. 2009; Elcombe et al. 2010; Minata et al. 2010; Pastoor et al. 1987; Perkins et al.
2004; Son et al. 2008; Wolf et al. 2008a). Increased liver weights were observed in mice that are
both active and null for PPARa activation (Albrecht et al. 2013; Minata et al. 2010; Wolf et al.
2008a). The histological characteristics of the liver differed in the mice with and without the
PPARa receptor, but the liver weight increase was the same. Liver effects were seen in mice
with an active PPARa 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 PPARa-
null mice were distinctly different from their respective controls (Minata et al. 2010; Wolf et al.
2008a). In the case of the wild-type PFOA-exposed mice, there was less rough ER than in
controls and more lipid-like vacuoles scattered throughout the cytoplasm. The PFOA-exposed
PPARa-null mice had proliferation of smooth ER and limited rough ER and Golgi bodies
compared to their controls. The PPARa-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
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cytoplasm. The vacuoles in the PPARa-null PFOA-exposed mice were hypothesized to be filled
with PFOA as a consequence of its uptake into the cell without dispersion or assimilation.
Similarly, Albrecht et al. (2013) observed centrilobular hepatocellular hypertrophy in mouse
dams given 3 mg/kg on GDs 1-17, but the morphological features differed slightly between
wild-type, PPARa-humanized, and PPARa-null mice. In wild-type mice, hypertrophy was
characterized primarily by centrilobular hepatocytes with increased amounts of densely
eosinophilic and coarsely granular cytoplasm consistent with increased peroxisomes. In null
mice, hypertrophy was generally less prominent than seen in wild-type mice, and affected
hepatocytes had pale eosinophilic, finely granular-to-amorphous cytoplasm. The morphological
features of centrilobular hepatocytes in humanized mice were intermediate between those
observed in wild-type and null mice. The lesion was graded as mild in wild-type mice, minimal
in null mice, and minimal or mild in humanized mice. An additional finding in PFOA-treated
null and humanized mice, but not in wild-type mice, was the presence of few clear, discrete
vacuoles within the cytoplasm of centrilobular hepatocytes.
Hepatocellular hypertrophy and an increased liver-to-body weight ratio are common findings
in rodents when PPARa activation leads to peroxisome proliferation and these effects are
considered nonadverse in wild-type strains when they occur. Hepatic necrosis, effects on bile
ducts, and other signs of liver damage unrelated to PPARa activation observed in conjunction
with the increased liver weight and hepatpcellular hypertrophy are sufficient to justify the liver
weight and hypertrophy as adverse (Hall et al. 2012). Low-level necrotic cell damage was
observed in the Perkins et al. (2004) rat study and in the Loveless et al. study (2008) in CD rats
at 10 mg/kg/day and CD1 mice at 1 mg/kg/day. In the Perkins et al. study (2004), there was a
slight increase in coagulative necrosis at 1.94 and 6.5 mg/kg/day when compared to the control
and lower doses. Some hepatocellular necrosis also was observed in conjunction with
hepatocellular hypertrophy and increased liver weight in Fl male rats from the Butenhoff et al.
(2004a) two-generation study at 3 mg/kg/day.
Minata et al. (2010) reported degenerative histological changes in the bile ducts of PPARa-
null mice at doses >10.6 mg/kg/day and Loveless et al. (2008) observed bile duct hyperplasia in
CD1 mice at doses >10 mg/kg/day. PPARa-null mice had an increased hepatocyte PCNA
labeling index at a dose of 10 mg/kg/day (Wolf et al. 2008a). When considering the studies in
animals with and without the active PPARa receptor, it is clear that PFOA has some effects of
potential toxicological significance that appear to be independent of PPARa activation.
Kidney and Other Organ Effects. Overall, studies of occupational cohorts (Costa et al. 2009), a
highly exposed community (Steenland et al. 2010; Watkins et al. 2013), and the U.S. general
population (Shankar et al. 2011) that evaluated uric acid levels or eGFR as measure of kidney
function found associations with decreased function. Reverse causality as an explanation cannot
be ruled out in studies using serum PFOA as a biomarker of exposure, as a low GFR would
diminish the removal of PFOA from serum for excretion by the kidney.
Some studies in animals have shown effects on the kidney, mainly increased organ weight in
male rats, but the studies lacked concurrent PFOA serum levels and histological examination of
the kidney tissues. In general, kidney effects in rats occurred at doses similar to those resulting in
liver effects.
Increases in absolute and relative-to-body kidney weights occurred in rats given 5 mg/kg/day
(lowest dose tested) for 28 days (Cui et al. 2009). In a two-generation study, FO and Fl males
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had significantly increased absolute kidney weight at 1 and 3 mg/kg/day, but significantly
decreased organ weight at 30 mg/kg/day. Organ weight-to-terminal body weight ratios for the
kidney were statistically significantly increased at >1 mg/kg/day. Kidney weight-to-brain weight
ratios were significantly increased at 1, 3, and 10 mg/kg/day, but decreased at 30 mg/kg/day,
following the trends in absolute weights (Butenhoff et al. 2004a). In the high-dose group,
absolute and relative kidney weight changes occurred in a pattern typically associated with
decrements in body weight. However, in the lower dose groups kidney weight, consistently
displayed an increase (absolute and relative to body and brain weights), suggesting an induction
of transporters for renal removal of the foreign molecule. The differential expression of
transporters in the kidney of rats has been shown to be under hormonal control with males
having lower levels of export transporters compared with females (Kudo et al. 2002).
In both the Cui et al. (2009) and Butenhoff et al. (2004a) studies, PFOA was administered by
daily gavage. No changes in kidney weight were found with dietary administration with a
resulting dose of 14.2 mg/kg/day to male rats for 2 years (Butenhoff et al. 2012).
In general, effects on organs other than the liver tend to occur at doses higher than those that
affect the liver. Lung effects, including pulmonary congestion, have been observed in male
Sprague-Dawley rats (LOAEL = 5 mg/kg/day) (Cui et al. 2009). Increased thickness and
prominence of the adrenal zona glomerulosa and vacuolation in the cells of the adrenal cortex
were observed in male rats fed 10 mg/kg/day for approximately 56 days (Butenhoff et al. 2004a).
Thyroid Effects. Three large studies provide support for an association between PFOA exposure
and incidence or prevalence of thyroid disease in female adults or children, but not in males
(Lopez-Espinosa et al 2012; Melzer et al. 2010; Winquist and Steenland 2014b). In addition,
associations between PFOA and TSH were also seen in pregnant females with anti-TPO
antibodies (Webster et al. 2014). However, generally null associations were found between
PFOA and TSH or thyroid hormones (T4 or T3) in people who have not been diagnosed with
thyroid disease.
Effects of PFOA on thyroid hormones in animals are generally not as well characterized as
those of PFOS. Butenhoff et al. (2002) evaluated the toxicity of PFOA in male cynomolgus
monkeys during 6 months of oral administration and reported that levels of total T3 and FT3 in
circulation were reduced significantly in the 30/20 mg/kg/day treatment group. The effect seen as
early as 5 weeks after initiation of treatment, 2 weeks after the dose was lowered to 20
mg/kg/day. Recovery of T3 deficits was noted upon cessation of chemical treatment once the
serum level of PFOA returned to baseline 90 days later. Serum total T4, FT4, and TSH were not
altered throughout the study. The preferential effects of PFOA on serum T3 and a lack of a TSH
compensatory response are similar to those observed with PFOS.
Martin et al. (2007) showed that serum total T4 and FT4 were markedly and abruptly
depressed (~ 80%) in adult male rats 1 day after oral gavage treatment with PFOA (20 mg/kg);
serum T3 was also reduced (25%), although to a lesser extent. These findings were confirmed
when both male and female rats were given PFOA (10 mg/kg) daily for 3 weeks and serum
thyroid hormones were monitored (Lau, personal communication) (Martin et al. 2007). Serum
total T4 and FT4 were profoundly depressed (>85%) and T3 less so (~ 25%) in male rats, but
serum TSH levels were not altered significantly. These hormonal changes were noted when
serum PFOA level reached about 67 ug/ml. The dose-response relationship of serum total T4
with PFOA exposure has yet to be fully evaluated and the lowest effective dose remains
unknown.
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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 could be
relevant to humans because of the long PFOA human half-life.
Displacement of T4 from binding to TTR has been proposed as a possible mechanism to
account for the hypothyroxinemia in rats. However, although PFOA binds to human TTR, its
binding affinity is only one-fifteenth of that of the natural ligand T4 (Weiss et al. 2009). Based
on a toxicogenomic analysis of rat liver after an acute exposure to PFOA, Martin et al. (2007)
suggested a possible role of peroxisome proliferators in the thyroid hormone imbalance, although
this hypothesis has yet to be explored in detail.
Hyperglycemia. Several human epidemiology studies have examined PFOA in relation to
diabetes (incidence or prevalence) or measures of hyperglycemia These studies do not show a
pattern of results that suggest an association between PFOA and diabetes or hyperglycermia in
occupational settings (Costa et al. 2009; Olsen et al. 2000, 2003; Sakr et al. 2007a; Steenland et
al. 2015), in the high-exposure community population (MacNeil et al. 2009), or in the general
population (Lin et al. 2009; Nelson et al. 2010).
Hines et al. (2009) found no differences in glucose tolerance tests at 15-16 weeks and at
17 months of age in PFOA-exposed CD-I mice, but did observe significantly increased serum
leptin and insulin levels at 21 and 31 weeks of age, suggesting that the insulin resistance
mechanistic pathway could be affected by PFOA. Conversely, Quist et al. (2015) found no dose-
related impact on serum leptin in CD-I pups from the Hines et al. study (2009) when examined
on PND 91 for the mice on an RFD and on an HFD fasted for 4 hours before serum collection. In
the animals on a HFD that did not fast before serum collection, there was a trend towards a dose-
related decrease in serum leptin. Thus, the fat content of the diet and the timing of serum
collection are important variables that can influence study results relative to leptin levels and
indicators of insulin resistance.
Nervous System Effects. The data pertaining to neurotoxicity (including neurodevelopmental
effects) of PFOA are limited, but do not indicate the presence of associations between PFOA and
a variety of outcomes. Fei et al. (2008b) found no association between maternal serum PFOA
concentrations and fine motor skills, gross motor skills, and cognitive abilities of children aged 6
and 18 months. Fei and Olsen (2011) found no association between behavioral or coordination
problems in children aged 7 years and prenatal PFOA exposure. Epidemiology studies of
children derived from the NHANES and C8 populations found a weak statistical association
between serum PFOA with parental reports of ADHD (Hoffman et al. 2010; Stein et al. 2013).
One animal study (Johansson et al. 2009) suggests a potential effect on habituation and
activity patterns in NMRI mice treated on PND 10 with a single dose of PFOA and evaluated at
and 2 and 4 months of age (LOAEL=0.58 mg/kg). The in vivo observations were supported by
changes in the expression of a variety of neurologically active brain proteins in the treated pups
(Johansson et al. 2009). The offspring of C57BL/6/Bkl dams fed 0.3 mg PFOA/kg/day
throughout gestation had detectable levels of PFOA in their brains at birth (Onishchenko et al.
2011). Behavioral assessments of the offspring starting at 5 weeks of age revealed gender-related
differences in exploratory behavior patterns. In the social group setting, the PFOA-exposed
males were more active and PFOA-exposed females were less active than their respective
controls. The PFOA-exposed males also had increased activity counts compared to control males
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in circadian activity experiments. The results of an in vitro study of hippocampal synaptic
transmission and neurite growth in the presence of long-chain perfluorinated compounds showed
that 50 and 100 umol PFOA increased spontaneous synaptic current and had an equivocal impact
on neurite growth (Liao et al. 2009a, 2009b). These data suggest a need for additional studies of
the effects of PFASs, including PFOA, on the brain.
Reproductive and Developmental Effects. There have been numerous human studies
examining PFOA exposure and reproductive and/or developmental effects in both humans and
animals. A series of studies in the high-exposure C8 Health Project study population have
reported associations between PFOA exposure and pregnancy-induced hypertension or
preeclampsia (Darrow et al. 2013; Savitz et al. 2012a, 2012b; Stein et al. 2009). Each of these
studies provides evidence of an association between PFOA exposure and risk of pregnancy-
induced hypertension or preeclampsia, with the most robust findings from the methodologically
strongest study (Darrow et al. 2013).
The association between PFOA and birth weight has been examined in numerous human
studies. Most studies measured PFOA using maternal blood samples taken in the second or third
trimester or in cord blood samples. Studies on the high-exposure C8 community population
(Darrow et al. 2013; Nolan et al. 2009; Stein et al. 2009; Savitz et al. 2012a, 2012b) have not
observed associations between PFOA and either birth weight among term births or the risk of
low birth weight among all (singleton) births. In contrast, several analyses of general populations
indicate a negative association between PFOA levels and birth weight (Apelberg et al. 2007; Fei
et al. 2007; Maisonet et al. 2012), while others did not attain statistical significance (Chen et al.
2012; Hamm et al. 2010; Monroy et al. 2008; Washino et al. 2009). A meta-analysis of many of
these studies found a mean birth weight reduction of 19 g (95% CI: -30, -9) per each 1-unit
(ng/mL) increase in maternal or cord serum PFOA levels (Johnson et al. 2014). However, when
low GFR was accounted for in PBPK simulations by Verner et al. (2015), the association
reported between PFOA and birth weight is less than that found in their meta-analysis of the
epidemiology data. The study authors reported that the actual association might be closer to a 7-g
reduction (95% CI: -8, -6). Verner et al. (2015) also showed that, in individuals with low GFR,
there are increased levels of serum PFOA and lower birth weights. This suggests that a portion of
the association between PFOA and birth weight could be confounded by low maternal GFR
under conditions such as preeclampsia and pregnancy-induced hypertension. While there is some
uncertainty in the interpretation of the observed association between PFOA and low GFR with
birth weight, given the available information, the association between PFOA exposure and
reduced birth weight observed for the general population is plausible. In humans with low GFR,
the impact on body weight is likely due to a combination of the low GFR and the serum PFOA.
Two studies examined development of puberty in females in relation to prenatal exposure to
PFOA as measured through maternal or cord blood samples in follow-up of pregnancy cohorts
conducted in England (Christensen et al. 2011) and in Denmark (Kristensen et al. 2013). The
results of these two studies are conflicting, with no association (or a possible indication of an
earlier menarche seen with higher PFOA) in Christensen et al. (2011), and a later menarche seen
with higher PFOA in Kristensen et al. (2013). Another study examined PFOA exposure
measured concurrently with the assessment of pubertal status (Lopez-Espinosa et al. 2011). An
association between later age at menarche and higher PFOA levels was observed, but the
interpretation of this finding is complicated by the potential effect of puberty on the exposure
biomarker levels (i.e., reverse causality).
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Limited data suggest a correlation between higher PFOA levels (>0.02 jig/mL) in females
and decreases in fecundity and fertility (Fei et al. 2009; Velez et al. 2015), but there are no clear
effects of PFOA on male fertility endpoints (0.0035-0.005 |ig/mL) (Joensen et al. 2009, 2013).
Knox et al. (2011) found that the odds of having experienced menopause were significantly
higher in the highest PFOA quintile group relative to the lowest PFOA group. Two studies found
delayed puberty in females (Kristensen et al. 2013; Lopez-Espinosa et al. 2011), but reverse
causality needs to be considered. However, Christensen et al. (2011) found no association
between puberty and PFOA exposure in children of the Avon Longitudinal Study of Parents and
Children in the United Kingdom. Removal of PFOA with the start of monthly menstruation and
the cessation of this route with menopause or hysterectomy are additional factors that can
influence serum PFOA levels that are the result of the developmental milestones rather than a
cause (Taylor et al. 2014; Wong et al. 2014). Costa et al. (2009) found no association between
serum PFOA concentration and estradiol or testosterone in workers at a PFOA production plant.
Measures of postnatal development and behavior in children were not associated with PFOA
levels in the mother (0.001-0.0057 |ig/mL) (Andersen et al. 2010, 2013; Fei et al. 2008b; Fei and
Olsen 2011; H0yer et al. 2015a, 2015b). Fei et al. (2008b) found no association between
maternal PFOA concentration and fine motor skills, gross motor skills, and cognitive skills in
offspring at 6 and 18 months of age. Fei and Olsen (2011) also found no association between
prenatal PFOA exposure and behavior or coordination problems in children aged 7 years. The
age at which children reached developmental milestones did not show any relationship to
maternal plasma PFOA concentration. Halldorsson et al. (2012) found that low-dose
developmental exposures to PFOA resulted in obesogenic effects in female offspring at 20 years.
Among the animal studies, there was no effect of PFOA on reproductive or fertility
parameters in rats (Butenhoff et al. 2004a; York et al. 2010), but effects on male fertility were
observed in male mice (Lu et al. 2015). In mouse gavage studies, decreased body weight and
decreased neonatal survival were observed at > 1 mg/kg/day, increased full litter resorptions and
increased stillbirths were observed at > 5 mg/kg/day, increased time to parturition was observed
at >10 mg/kg/day, and decreased maternal weight gain was observed at >20 mg/kg/day for
exposures lasting from GD1-17 (Abbott et al. 2007; Lau et al. 2006; White et al. 2007; Wolf et
al. 2007).
Postnatal development also has been studied extensively in rats and mice as discussed below.
A separate group of studies in mice focused on mammary gland development in dams and
female offspring. Both species showed some indication of potential developmental toxicity.
Doses that elicited a response were higher in rats compared with in mice. The species differences
in dose response are likely related to half-life differences of hours for the female rat and days-to-
weeks for the female mouse.
Reduced postnatal growth leading to developmental delays was observed in both rats and
mice. A two-generation diet study in rats resulted in significantly decreased body weight gain
prior to weaning and delayed sexual maturity in the first generation males and females at
30 mg/kg/day (Butenhoff et al. 2004a). For treatment beginning on PND 21, delayed vaginal
opening was also observed in BALB/c mice at >1 mg/kg/day and in C57BL/6 mice at
>5 mg/kg/day, although body weight was not decreased until doses of >10 mg/kg/day in both
strains (C. Yang et al. 2009). Cross-fostering studies in mice showed that gestational PFOA
exposure maximized decreased postnatal body weight, delayed eye opening, delayed body hair
growth, and decreased survival in the offspring (Wolf et al. 2007). Restricted exposure studies
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showed that gestational exposure to PFOA over differing gestational time periods led to differing
offspring effects (Wolf et al. 2007). The longer the gestational exposure, the greater the body
weight deficit in the male and female pups over PND 2-22. In males, the difference in body
weight persisted until PND 92. Delayed eye opening and body hair growth were observed at
5 mg/kg/day in offspring exposed GD 7-17 or 10-17, but decreased postnatal survival was
observed at the same dose in offspring exposed GD 15-17.
Two developmental studies compared wild-type mice with PPARa-null mice, but results are
inconclusive. One study revealed that the litter resorptions were independent of PPARa
expression (>5 mg/kg), while decreased neonatal survival (0.6 mg/kg) and delayed eye opening
(1 mg/kg) were dependent upon PPARa 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 PPARa-humanized mice, suggesting that the mouse PPARa could be involved in the
etiology of PFOA-induced neonatal mortality.
Qualitative assessment found delayed mammary gland development of female CD1 mouse
pups following maternal doses > 0.01 mg PFOA/kg in Macon et al. (2011) and Tucker et al.
(2015). Macon et al. (2011) also found significant differences from controls in quantitative
measures of longitudinal and lateral growth and numbers of terminal end buds at 1 mg/kg/day.
However, Albrecht et al. (2013) found no significant differences in the average length of
mammary gland ducts and the average number of terminal end buds per mammary gland per
litter in female pups of PPARa wild-type, PPARa-null, or hPPARa sv/129 mice following a
maternal dose of 3 mg/kg, using an approach to scoring that lacked a qualitative component
adjustment such as that used by Macon et al. (2011).
The approach to scoring mammary gland development was not consistent across studies, and
little information was provided on the qualitative components of the scores. This makes
comparisons across studies difficult. Statistical significance was attained at higher dose levels for
the quantitative portion of the Macon et al. (2011) scoring protocol than for the qualitative
component of the score. The process used to score the qualitative developmental score by Macon
et al. (2011) was not described. Tucker et al. (2015) found that CD-I mice were considerably
more sensitive to effects on mammary gland development (LOAEL = 0.01 mg/kg/day) than
C57BL/6 mice (NOAEL 0.1 mg/kg/day). Scoring was conducted using the Macon et al (2011)
approach.
White et al. (2011) used doses of 0 and 1 mg PFOA/kg/day to FO dams throughout gestation
with and without the addition of drinking water containing 5 ppb PFOA beginning on GD 7 and
continuing the contaminated drinking water during the production of two more generations; no
persistent significant differences were found in the body weights of the pups in the Fl and F2
generations for the pups receiving 1 mg/kg/day, indicating a poor correlation between mammary
duct branching patterns and the ability to support pup growth during lactation. The 5 mg/kg/day
dose did have an impact on body weight. Albrecht et al. (2013) also found no significant impacts
on pup body weight in their one-generation assay at a dose of 3 mg/kg/day. Despite the
diminished ductal network assessed in the qualitative mammary gland developmental score of
the dams in White et al. (2011), milk production was sufficient to nourish growth in the exposed
pups as reflected in the body weight measurements compared to controls. The MoA for PFOA-
induced delayed mammary gland development is unknown and requires further investigation.
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At doses of 5 and 10 mg/kg/day, mammary gland development was delayed in BALB/c mice
(C. Yang et al. 2009). In C57BL/6 mice, mammary gland development was accelerated at
5 mg/kg/day, but delayed at 10 mg/kg/day, indicating strain differences in pubertal mammary
gland development following a dose of 5 mg/kg/day. Y. Zhao et al. (2010) showed that 5 mg
PFOA/kg/day stimulates mammary gland development in C57BL/6 mice by promoting steroid
hormone production in the ovaries and increasing mammary gland growth factor levels.
Immune Effects. Associations between prenatal, childhood, or adult PFOA exposure and risk of
infectious diseases (as a marker of immune suppression) have not been consistently seen,
although there was some indication of effect modification by gender (i.e., associations seen in
female children but not in male children) (Fei et al. 2010a; Granum et al. 2013; Looker et al.
2014; Okadaetal. 2012).
The WHO guidelines for immunotoxicity risk assessment recommend measures of vaccine
response as a measure of immune effects, with potentially important public health implications
(WHO 2012). Three studies have examined associations between maternal and/or child serum
PFOA levels and vaccine response (measured by antibody levels) in children (Grandjean et al.
2012; Granum et al. 2013) and adults (Looker et al. 2014). The study in adults was part of the
high-exposure community C8 Health Project; a reduced antibody response to one of the three
influenza strains tested after receiving the flu vaccine was seen with increasing levels of serum
PFOA; these results were not seen with PFOS. The studies in children were conducted in general
populations in Norway and in the Faroe Islands. Decreased vaccine response in relation to PFOA
levels was seen in these studies, but similar results also were seen with correlated PFASs
(e.g., PFOS).
Several animal studies demonstrate effects on the spleen and thymus as well as their cellular
products (B lymphocytes and T-helper cells) in several strains of mice. Studies by Yang et al.
(2000, 2001, 2002b) and DeWitt et al. (2008) were conducted using relatively high PFOA doses
(-30-40 mg/kg/day). In each study, the PFOA-treated animals exhibited significant decreases in
spleen and thymus weights as well as in splenocyte and thymocyte populations at various stages
of differentiation. Recovery usually occurred within several days of cessation of PFOA dosing.
However, when the response of C57BL/6 Tac PPARa mice was compared to wild-type of the
same strain, the KO mice showed no response of both spleen and thymus weights at
30 mg/kg/day, whereas there was a response in the wild-type strain (DeWitt et al. 2015). Both
strains showed an increase in IgM in response to a SRBC injection. The 30-mg/kg/day dose was
the LOAEL for the KO mice and 7.5 mg/kg/day was the response level for the wild-type strain.
Thus the suppression of the immune system is not totally a PPARa-related response. In a similar
experiment (Yang et al. 2002a), no significant changes in spleen weight or cellularity were
observed in PPARa-null mice as compared to wild-type mice, but there was a small and
significant decrease in thymus weight and cellularity compared to controls.
DeWitt et al. (2008) used different functionality assays in their study in C57B1/6 mice. The
IgM response to SRBC was suppressed by 20% when mice were immunized immediately after
exposure to the initial dose of 30 mg PFOA/kg/day ceased. However, there was no significant
increase in the response to BSA 4 days post-PFOA exposure or in the IgG response to SRBC
15 days post-PFOA exposure. These results are indicative of recovery once PFOA exposures
have ceased.
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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,
C57B1/6N. The study design examined the spleen and thymus weights, splenocyte and
thymocyte numbers, and IgM response of the immune system to the immunological challenges
as described above. The LOAEL was 3.75 mg/kg/day based on a significant decrease in IgM
response, and the NOAEL was 1.88 mg/kg/day.
In one component of the Yang et al. study (2002b), the functional impact of changes in
spleen and thymus were evaluated through the response of treated mice to HRBCs. The control
mice responded to the HRBC exposure with an increased plaque-forming response; however, the
PFOA-treated mice did not have an increased plaque-forming response when tested (Yang et al.
2002b). In addition, when blood from PFOA-treated mice was evaluated posttreatment, there
was no increase in lymphocyte proliferation in response to the addition of Con-A and LPS to the
test media. The control mice responded with the expected lymphocyte proliferation after the
addition of Con-A and LPS antigens.
Loveless et al. (2008) looked at the IgM response to SRBC in male CD rats and CD-I 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 liters were not altered by PFOA treatment in rats. In mice, however,
thymus and spleen weights, thymus and spleen cell counts, and IgM liters were decreased at
>10 mg PFOA/kg/day. CORT also was increased in mice at the same doses.
The data collected from the immunotoxicity studies support a MoA through which PFOA
interferes with splenocyte and thymocyte precursor cells in the bone marrow as well as
maturation of the cells once they have been transported to their respective organs. Examination
of cell populations at different stages of development reveals lower numbers of the CD4"CD8"
cells formed in bone marrow as well as decreased populations of splenocyte and thymocyte cells
at different stages of expressing the surface proteins that mark them as functional beta
lymphocytes (thymus) or T-helper cells (spleen) (Son et al. 2009). Although the studies that
measured the splenocyte and thymocyte populations were carried out at doses higher than the
3.75 mg/kg/day LOAEL observed by DeWitt et al. (2008), the fact that the IgM response to an
antigenic material was decreased at that dose indicates an inability to produce antibodies at
adequate levels when exposed to a challenge.
Loveless et al. (2008) hypothesized that the observed effects on serum lymphocytes could be
the result of adenocorticotropic steroids in a response to stress. A study by DeWitt et al. (2009)
in which the immunological response of ADX mice treated with PFOA were compared to sham-
operated controls did not support the Loveless et al. (2008) hypothesis.
Data from PPARa-null mice suggest that rodents might be more susceptible to the
immunosuppressive impacts of PFOA than humans. However, the fact that there were still
effects on the thymus weight and cellularity even in the PPARa-null mouse strain indicate the
potential for an inadequate humoral response in exposed populations.
3.4.2 Synthesis and Evaluation of Carcinogenic Effects
Evidence of carcinogenic effects of PFOA in epidemiology studies is based primarily on
studies of kidney and testicular cancer. These cancers have relatively high survival rates
(e.g., 2005-2011 5-year survival rates 73% and 95%, respectively, for kidney and testicular
cancer based on NCI Surveillance, Epidemiology and End Results data). Thus studies that
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examine cancer incidence are particularly useful for these types of cancer. The high-exposure
community studies also have the advantage, for testicular cancer, of including the age period of
greatest risk, as the median age at diagnosis is 33 years. The two occupational cohorts in
Minnesota and West Virginia (most recently updated in Raleigh et al. 2014 and Steenland and
Woskie 2012) do not support an increased risk of these cancers, but each of these is limited by a
small number of observed cases (six kidney cancer deaths, 16 incident kidney cancer cases, and
five inciden testicular cancer cases in Raleigh et al. 2014; and 12 kidney cancer deaths and 1
testicular cancer death in Steenland and Woskie 2012). Two studies involving members of the
C8 Health Project showed a positive association between PFOA levels (mean at enrollment
0.024 |ig/mL) and kidney and testicular cancers (Barry et al. 2013; Vieira et al. 2013); there is
some overlap in the cases included in these studies. No associations were found in the general
population between mean serum PFOA levels up to 0.0866 |ig/mL and colorectal, breast,
prostate, bladder, and liver cancer (Bonefeld-J0rgensen et al. 2014; Eriksen et al. 2009; Hardell
et al. 2014; Innes et al. 2014); none of these studies examined kidney or testicular cancer.
Two animal carcinogenicity studies indicate that PFOA exposure can lead to liver adenomas
(Biegel et al. 2001), Leydig cell adenomas (Biegel et al. 2001; Butenhoff et al. 2012), and
PACTs (Biegel et al. 2001) in male Sprague-Dawley rats. Liver adenomas were observed in the
Biegel et al. study (2001) at an incidence of 10/76 (13%) at 20 mg/kg/day. The incidence in the
control group was 2/80 (3%). Although no liver adenomas were observed in Butenhoff et al.
(2012), carcinomas were identified in the male controls, males in the low-dose group
(2 mg/kg/day), and male and female rats in the high-dose group (20 mg/kg/day). The differences
from control were not significant in either study, but the carcinoma incidence among the
Butenhoff et al. (2012) high-dose males (10/50) was similar to that for the adenomas in the
Biegel et al. study (2001) (10/76). Liver lesions were identified in the males and females at the
1- and 2-year sacrifices (Butenhoff et al. 2012). An increased incidence of diffuse
hepatomegalocytosis and hepatocellular necrosis occurred at 20 mg/kg/day. At the 2-year
sacrifice, hepatic cystoid degeneration (characterized by areas of multilocular microcysts in the
liver parenchyma) was observed in 8, 14, and 56% in males of the control, 2-, and 20-mg/kg/day
dose groups, respectively. Hyperplastic nodules in male livers were increased in the high-dose
group (6% versus 0% in control rats).
Filgo et al. (2015) examined the livers of three strains of mice exposed only during
gestation/lactation for tumors when they were sacrificed at 18 months. Liver tumors were found
in each dose group, but tumor types varied and the data did not display any evidence of dose
response. The animals were survivors from two different projects and the number per dose group
was small. Thus, the data are not adequate for determining whether PFOA is a carcinogen in
mice.
Testicular LCTs were identified in both the Butenhoff et al. (2012) and Biegel et al. (2001)
studies. The tumor incidence was 0/50 (0%), 2/50 (4%), and 7/50 (14%) for the control, 2.0-, and
20-mg/kg/day dose groups, respectively (Butenhoff et al. 2012). The Biegel et al. study (2001)
included one dose group (20 mg/kg/day); the tumor incidence was 8/76 (11%) compared to 0/80
(0%) in the control group. LCT incidence at 20 mg/kg/day was comparable between the two
studies (11 and 14%).
PACTs were only observed in the Biegel et al. study (2001). The incidence was 8/76 (11%;
7 adenoma, 1 carcinoma) at 20 mg/kg/day while none were observed in the control animals.
Although no PACTs were observed by Butenhoff et al. (2012), pancreatic acinar hyperplasia was
observed at 2 and 20 mg/kg/day at incidences of 2/34 (6%) and 1/43 (2%), respectively, which
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lacked dose response. Reexamination of the pancreatic lesions in Butenhoff et al. (2012) and
Biegel et al. (2001) resulted in the conclusion that 20 mg/kg/day increased the incidence of
proliferative acinar cell lesions in both studies. Some lesions in the Biegel et al. study (2001) had
progressed to adenomas.
The initial findings from the Butenhoff et al. study (2012) were equivocal for mammary
fibroadenomas in female rats. However, a reexamination of the tissues by a PWG found no
statistically significant differences in the incidence of fibroadenomas or other neoplasms of the
mammary gland between control and treated animals (Hardisty et al. 2010). The PWG used the
diagnostic criteria and nomenclature of the Society of Toxicological Pathologists for the
reexamination. Under those criteria, there was an increase in the number of tumors documented
in the control group, especially fibroadenomas originally classified as lobular hyperplasia. The
reclassification led to a loss of significance when the tumors in the treated animals were
compared to tumors in the control animals.
Ovarian tubular hyperplasia and adenomas also were observed in female rats (Butenhoff et
al. 2012). Mann and Frame (2004) reexamined the ovarian lesions using an updated
nomenclature system, which resulted in some of the hyperplastic lesions being reclassified. The
ovarian lesions originally described as tubular hyperplasia or tubular adenomas were regarded as
gonadal stromal hyperplasia and/or adenomas. After the reclassification, there were no
statistically significant increases in hyperplasia (total number), adenomas, or
hyperplasia/adenoma combined in treated groups compared to controls.
Mutagenicity studies of PFOA using the S. typhimurium (Friere et al. 2008; Lawlor 1995,
1996) andE. coli (Lawlor 1995, 1996) system have resulted in negative results in the presence
and absence of activation. One mutagenicity study (Lawlor 1995, 1996) in S. typhimurium gave a
positive result, but it was not reproducible. Clastogenicity studies in CHO by Murli (1996b,
1996c) were positive with activation for chromosomal abnormalities and polyploidy and
equivocal in the absence of activation. Micronucleus assays by Murli (1995, 1996d) were
negative.
A significant increase in 8-OH-dG liver levels, a biomarker for oxidative stress, was
observed at > 10 mg PFOA/kg in the liver but not the kidney of Fischer 344 male rats by Takagi
et al. (1991). Work with HepG2 cells by Hu and Hu (2009) suggested that PFOA could induce
apoptosis by overwhelming the homeostasis of antioxidative systems, increasing ROS, impacting
mitochondria, and changing expression of apoptosis gene regulators. Eriksen et al. (2010)
observed a PFOA-induced increase in ROS production in HepG2 cells, but no PFOA-induced
oxidative DNA damage or cytotoxicity.
3.4.3 Mode of Action and Implications in Cancer Assessment
The modes of lexicological/carcinogenic action of PFOA are not clearly understood.
However, available data suggest that the induction of tumors is likely due to nongenotoxic
mechanism involving membrane receptor activation, perturbations of the endocrine system,
and/or the process of DNA replication and cell division. PFOA lacks the ability to react with and
modify DNA, although its electrostatic properties would permit interaction with chromosomal
histone proteins with a net positive surface charge.
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Rat Liver Tumors. PPARa agonism has been proposed as a potential MOA for the liver
carcinomas and adenomas in rats following chronic PFOA exposure (Maloney and Waxman
1999; Klaunig et al. 2003, 2012). In the PPARa agonism MOA, binding of PFOA to the PPARa
receptor results in increased peroxisome proliferation and cell replication. PPARa is primarily
expressed in the liver, but also is present in the kidney, intestines, heart, and brown adipose
tissue (Hall etal. 2012).
Peroxisomes are single-membrane organelles found in a number of plant and animal cells
that have the capacity to carry out beta oxidation of long-chain fatty acids and other substrates
through hydrogen peroxide-generating pathways and without the generation of ATP (Goodrich
and Sul 2000). Peroxisomes metabolize the long-chain fatty acids via both beta and omega
oxidation pathways (Fielding 2000), but are unable to metabolize fatty acid chains of eight
carbons or less (Garrett and Grisham 1999). The shorter chain fatty acids are exported to the
cytosol and taken up by mitochondria for further degradation via beta oxidation with resultant
production of acetyl-CoA and ATP.
When a chemical binds to and activates the PPARa 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 PPARa receptor, including unsaturated CIS fatty acids, metabolites of arachidonic acid, and
the prostaglandin metabolite PGJ2 (Spector 2000). PPARa activation is accompanied by
upregulation of many genes associated with catabolism of fatty acid and cholesterol biosynthesis
and lipid transport (Hall et al. 2012; Rosen et al. 2008a).
There are four key events in the PPARa-agonist MOA for liver tumors (Klaunig et al. 2003,
2012) (see Figure 3-1). The first key event is activation of PPARa. Increased palmitoyl-CoA
oxidase activity is used in many studies as a biomarker for PPARa activations. Other associated
indicators are hepatocellular hypertrophy and increased liver weight. However, these indicators
alone are not sufficient to establish a PPARa MOA because they also are caused by chemicals
that have no influence on PPARa.
The primary data that demonstrate PFOA activation of the PPARa receptor are those from
Rosen et al. (2008a, 2008b) that examined the transcript profiles in the livers of wild-type and
PPARa-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 PPARa gene activator Wyeth 14,643
and PPARa-null mice. Based on the analysis of gene regulation, it was clear that PPARa
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 PPARa agonist.
Multiple studies in both rats and mice provide evidence that PFOA induces peroxisome
proliferation in the liver (Elcombe et al. 2010; Minata et al. 2010; Pastoor et al. 1987; Wolf et al.
2008b; Yang et al. 2001). PFOA also was found to activate mouse and human PPARa using a
transient transfection cell assay (Takacs and Abbott (2007). Maloney and Waxman (1999) also
demonstrated that 5-10 umol PFOA (2 to 4 mg/L) activated mouse PPARa using COS1 cells
(kidney fibroblast-derived cells) transfected with a luciferase reporter gene.
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Key Events in the Mode of Action for PPARa -
Agonist Induced Rodent Liver Tumors
Causative Events
PPARa Agonist
Acttatkx, of PPARa Associate Events'
Cell Proliferation
De ere ase d A pop t osi s
i
Preneoplastie Foci
*
Qonal Expansion
*
Liver Tumors
Ex pressio n of Pe roxi som al Ge nes
Increase in Peroxisomes (number
& size)
Mhoudi the Bane othff bidogicalewnte (ag. Kupffer cell mediatedevents, inhibition of gsp jure lore 1 the
measuremene of per«i!»mepraliferafonandpetDxisomalerE>fne activity (in parfcJar asyt-CoA) are sudelyusedas
reliable markersofPFftRa scdvakn.
Source: USEPA 2005c
Figure 3-1. PPARa Agonist MoA for Liver Tumors
In rodents, hepatic physical and biochemical changes observed after activation are highly
correlated with liver tumors leading to the hypothesis that a > 3-fold increase in peroxisomes and
> 1.5 fold increase in liver weights in short-term studies are sufficient to cause liver cancers in
long-term studies (Hall et al. 2012). The temporal and dose-response relationship of measures of
peroxisome proliferation, hepatocellular hypertrophy, liver weight, and liver histopathology were
examined in male Sprague-Dawley rats following 4, 7, and 13 weeks of administration of
dietary PFOA at doses ranging from 0-6.5 mg/kg/day (Perkins et al. 2004). There was no
evidence of peroxisome proliferation, hepatocellular hypertrophy, or liver weight increases at
0.06 mg/kg/day. However, at 13 weeks, the 6.5-mg/kg/day dose had an increase in palmitoyl-
CoA oxidase activity (an indicator for peroxisomes) that was 3.4 times greater than that for the
pair-fed control. The absolute liver weight was 1.6 times greater than the pair-fed control. At the
lower 1.94 mg/kg/day doses, the increases were 2.8 and 1.4 for the palmitoyl-CoA and liver
weight, respectively
There are indications that PFOA also acts through PPARa-independent mechanisms
associated with CAR and PXR receptors. Martin et al. (2007) examined the genomic signature
from PFOA-treated Sprague-Dawley rats (up to 5-day exposure) using microarray expression
profiling, pathway analysis, and quantitative PCR. The animal responses were consistent with
PPARa agonism, but there was also evidence of PPARy agonism (downregulation of cholesterol
synthesis) and activation of CAR- and PXR-related genes. CAR activation can lead to hepatocyte
proliferation and hepatocarcinogenesis in animals. However, the human CAR receptor is
relatively resistant to mitogenic effects and less likely to induce cancers through this mechanism
(Hall et al. 2012). In rodents, the PXR receptor can interact with PPARa in the coordination of
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hepatocyte proliferation, but there are differences in the amino acid composition of the ligand
binding domain of the mouse receptor and the human receptor (10% homology) (Hall et al.
2012). Accordingly, although the line of evidence is strongest for PPARa activation as the
initiator for the downstream events in the PFOA cancer MO A, there can be involvement from
other membrane receptors other than PPARa.
The second step in the PPARa MoA calls for evidence for increased cell proliferation and
decreased apoptosis. Few studies examined the occurrence of these events with PFOA. Son et al.
(2008) saw evidence of decreased apoptosis in liver and kidney cells stained for caspaceS in
4-week-old male ICR mice treated for 21 days at a dose of about 20 mg/kg/day. However,
Elcombe et al. (2010) failed to see a significant decrease in male Sprague-Dawley rats with a
28-day exposure to a diet containing 300 ppm (-20 mg/kg/day) PFOA (comparable to the high
dose in both cancer studies). In wild-type 129S4/SvlmJ mice, Minata et al. (2010) observed
increased apoptosis in hepatocytes, arterial walls, and bile duct epithelium and in the bile duct
epithelium of PPARa-null mice at 10.8 and 21.6 mg/kg PFOA. Thus, the apoptosis data for
PFOA are not consistently supportive of the key step in this proposed MoA (i.e., a decrease in
apoptosis).
Using a BrdU labeling technique, Elcombe et al. (2010) observed significant increases in cell
proliferation in male Sprague-Dawley rats after 1, 7, and 28 days of exposure to a 300-ppm
PFOA dietary dose. The highest increase was observed after 7 days of treatment (a fivefold
increase) and declined to a twofold increase after 28 days of dosing. The liver results from the
Biegel et al. (2001) mechanistic study were negative for cell proliferation in male Sprague-
Dawley rats exposed to the same dietary concentration (20 mg/kg/day) and sacrificed at 1, 3, 6,
9, 12, and 15 months. However, based on the Elcombe et al. (2010) observations, the timing of
the interim sacrifice would have missed the peak of the proliferative response. The Butenhoff et
al. study (2012) identified hyperplastic nodules in 3/50 high-dose males and 2/50 high-dose
females at 20 mg/kg/day; 5/50 males and 1/50 females had hepatocellular carcinomas.
The study by Wolf et al. (2008a) looked at the labeling index in 129Sl/SvlmJ mice and
PPARa-null mice and found a difference in their dose response. In the wild-type mice, the
labeling index was increased at all doses > 1 mg/kg/day; however, in PPARa-null mice, the
labeling index was increased only at the highest dose, 10 mg/kg/day.
There were no studies identified that focused specifically on preneoplastic foci and clonal
expansion of altered cells after PPAR activation. Minata et al. (2010) observed a dose-dependent
increase in eosinophilic cytoplasmic changes consistent with peroxisome proliferation in liver
parenchyma, but found no focal necrosis at doses < 21.6 mg/kg/day in wild-type 129S4/SvlmJ
mice.
Klaunig et al. (2003, 2012) concluded that, based on the available data, PFOA-induced liver
tumors in Sprague-Dawley rats can be attributed to a PPARa MOA since there are data available
addressing most of the key steps in this proposed MoA. However, some data gaps exist for key
events and other mechanisms that might be involved. Overall, the tumor response observed in the
available studies was not strong and did not demonstrate a dose-related response in males (3/49,
1/50, and 5/50 hepatocellular carcinomas in the control, 2-mg/kg/day, and 20-mg/kg/day dose
groups, respectively) and a single carcinoma in females at the high dose. Biegel et al. (2001) did
not identify any liver carcinomas (0/76) in males at their 20-mg/kg dose, but there were 10/76
males with adenomas. This is consistent with a hyperplastic tissue response rather than the loss
of cell cycle control characteristic of cancer. The data from the Butenhoff et al. (2012) and
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Biegel et al. (2001) studies suggest that PFOA is not a potent hepatic carcinogen based on the
low tumor incidence and finding of hyperplastic nodules.
Leydig Cell Tumors (LCT). LCTs were observed in both the Butenhoff et al. (2012) and Biegel
et al. (2001) studies. The LCT incidence was 0/49, 2/50, and 7/50 at doses of 0, 2, and
20 mg/kg/day, respectively, in Butenhoff et al. (2012) and 2/78 (pair-fed control) and 8/76 at
20 mg/kg/day in Biegel et al. (2001).
A large number of nongenotoxic compounds of diverse chemical structures have been
reported to induce LCTs in rats, mice, or dogs. LCTs also occur in humans but are relatively rare
at about 1-3% of human testicular tumors, which also are infrequent (1%) (Carpino et al. 2007).
A workshop report (Clegg et al. 1997) on the MOA for LCT classified the chemicals that caused
LCT in animal studies into seven MOA categories. The postulated MO As support the following
hormonal steps to the process:
1. A xenobiotic chemical inhibits the production of testosterone, leading to low serum
levels.
2. Low serum testosterone levels signal the hypothalamus to produce gonadotropin
releasing hormone (GnRH).
3. GnRH signals the pituitary to release LH.
4. LH signals the Leydig cells to produce testosterone.
5. LH causes Leydig cell proliferation.
Several of the available PFOA studies support an impact of PFOA on decreased testosterone
production. Studies conducted by Cook and colleagues (Biegel et al. 1995; Cook et al. 1992; Liu
et al. 1996) found that adult male rats administered PFOA by gavage for 14 days had decreased
serum testosterone and increased serum estradiol levels (Cook et al. 1992). These endocrine
changes correlated with its potency to induce LCTs (Biegel et al. 2001).
Subsequent experiments demonstrated that PFOA increased levels of estradiol by inducing
cytochrome P450 CYP19 (aromatase). Aromatase converts androgens to estrogens, including the
conversion of testosterone to estradiol. PFOA directly inhibits testosterone production when
incubated with isolated Leydig cells and ex vivo studies demonstrate that this inhibition is
reversible (Biegel et al. 1995). However, in the mechanistic bioassay by Biegel et al. (2001),
serum testosterone and LH levels were not significantly altered at the levels of PFOA that
resulted in LCTs (20 mg/kg/day).
This inhibition of testosterone biosynthesis can be mediated by PPARa (Gazouli et al. 2002).
Support for PPARa-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 PPARa wild-type and humanized PPARa mice
but not in PPARa-null mice. Similarly, disruption of testosterone biosynthesis by lowering the
delivery of cholesterol into the mitochondria and decreasing the conversion of cholesterol to
pregnenolone and androstandione in the testis was noted in wild-type and humanized PPARa
mice. These effects were not seen in PPARa-null mice. Decreased serum testosterone was noted
after oral exposure to PFOA in studies by Biegel et al. (1995, 2001) and Cook (1992).
The induction of LCTs by PFOA also can be attributed to a hormonal mechanism whereby
PFOA either inhibits testosterone biosynthesis and/or lowers testosterone by increasing its
conversion to estradiol through increased aromatase activity in the liver. Both of these
mechanisms appear to be mediated by PPARa. However, data are not currently sufficient to
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demonstrate that the other key steps in the postulated MOA are present in PFOA-treated animals
following exposures that lead to tumor formation. Studies are needed to demonstrate the increase
of GnRH and LH in concert with the changes in aromatase and estradiol. There was also no
indication of increased Ley dig cell proliferation at the doses that caused adenomas in the Biegel
et al. study (2001). Thus, additional research is needed to determine if the hormone testosterone-
estradiol imbalance is a key factor in development of LCTs as a result of PFOA exposure.
Two studies involving members of the C8 Health Project showed a positive association
between PFOA levels (mean at enrollment 0.024 jig/mL) and kidney and testicular cancers
(Vieira et al. 2013; Barry et al. 2013). This contributed to the EPA conclusion that PFOA can be
classified as having suggestive evidence for carcinogenicity.
Pancreatic Acinar Cell Tumors. The 2-year bioassay by Biegel (2001) identified PACTs in 7/6
rats receiving a 20-mg/kg dose for 2 years compared to 1/79 in the pair-fed controls. As with
LCTs, the MOA for PACTs is not understood. These tumors are most commonly identified in
rats, but do occur in other animal species (e.g., mice, hamsters) and in humans (Wisnoski et al.
2008). Males are more susceptible to pancreatic tumors than females. Two hypothetical MO As
have been proposed and are as follows (Klaunig et al. 2003, 2012; Obourn et al. 1997):
There is a change in the bile acid flow or composition that leads to cholestasis, thereby
causing an increase in CCK activating a feedback loop resulting in proliferation of the
secretory pancreatic acinar cells. CCK is a peptide hormone that stimulates the digestion
of fat and protein, causes the increased production of hepatic bile, and stimulates
contraction of the gall bladder. An FIFD, trypsin inhibition, and changes in bile
composition are proposed initiators for this sequence of events.
Increased levels of testosterone support the growth of acinar cell preneoplastic foci,
leading to the development of carcinomas.
There is minimal information on the relationship of PFOA exposure to either of the proposed
MO As. Obourn et al. (1997) studied the impact of PFOA on CCK and trypsin using in vitro
assays and found that PFOA was not an agonist for the CCKA receptor that activates CCK
release. PFOA also had no inhibitory action on trypsin at levels 1,000 times greater (0.31 ug/mL)
than the positive control.
The Obourn et al. study (1997) also looked at Wyeth 14,643, a peroxisome proliferator, in
these same assays and found results similar to those for PFOA. When they conducted an in vivo
study with 100 ppm Wyeth 14,643, they found a small but significant increase (p<0.05) in bile
flow per unit body weight, a decrease (p<0.05) in bile flow per unit liver weight, and a small
decrease (p<0.05) in the total bile acid concentration following a 6-month dietary exposure.
There is the potential for PFOA to change the composition of bile because of its competition
with bile acids for biliary transport. In mice, increased expression of MRP3 and MRP4
transporters (Maher et al. 2008) and decreased expression of OATPs (Cheng and Klaassen 2008)
favor excretion of PFOA into the bile. Minata et al. (2010) found the levels of PFOA in bile from
wild-type male mice to be considerably higher than those in PPARa-null mice, suggesting a link
to PPARa. In the same study, male wild-type and PPARa-null mice were orally dosed with ~0,
5.4, 10.8, and 21.6 mg/kg/day of PFOA for 4 weeks. Total bile acid was significantly increased
at the highest dose in PPARa-null mice suggesting that, in the presence of PFOA, activation of
PPARa increases PFOA excretion, a scenario that could possibly decrease the flow of bile acids
competing for the same transporters. In the Butenhoff et al. study (2012), there was a lack of
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PACT tumors but an increase in proliferative lesions of the acinar cells. One hypothesis offered
for the difference in results was differences in the diets used in the two studies (Chang et al.
2014).
PFOA appears to suppress testosterone production through the induction of aromatase
(Biegel et al. 1995; Cook et al. 1992; Liu et al. 1996) and to increase the estradiol. Accordingly,
the second proposed MOA for PACTs does not appear to apply to PFOA.
The data on a PPARa-linked MoA are strongest for the liver tumors. Some data also provide
a link of PPARa to the Ley dig cell and PACT tumors observed in the rat 2-year bioassays. They
are not as strong and identify a need for additional research justifying the suggestive evidence
finding. However, when integrated with the metabolic inertness of PFOA in animals and
humans, a linear response to dose is not likely. This is consistent with the tumor data. Thus a
nonlinear MOA is likely and the remaining challenge is to identify the critical event in each
MOA that leads to development of the tumors.
Other Potential Modes of Action. There are other potential MO As that could apply to PFOA.
They include interruption of intercellular communication, mitochondrial effects, and hormonal
effects. None of these mechanisms are considered to be key steps in the MO As discussed above.
GJIC, a process by which cells exchange ions, messages, and other small molecules, is
important for normal growth, development, and differentiation as well as for maintenance of
homeostasis in muticellular organisms. Because tumor formation requires loss of homeostasis
and many tumor promoters inhibit GJIC, it has been hypothesized that GJIC might play a role in
carcinogenesis (Trosko et al. 1998). PFOA has been demonstrated to inhibit GJIC in liver cells
in vitro and in vivo (Upham et al. 1998, 2009). However, inhibition of GJIC is a widespread
phenomenon, and the effect by PFOA was neither species- nor tissue-specific. In addition it was
reversible. Thus, the significance of GJIC inhibition in regard to the mode of carcinogenic action
of PFOA is unknown.
Several chemicals structurally related to PFOA have been shown to manifest their toxicity by
interfering with mitochondria biogenesis and bioenergetics. Walters et al. (2009) found evidence
supporting mitochondrial proliferation in Sprague-Dawley rats receiving 30 mg/kg/day of PFOA
for 28 days as reflected in measurements of mitochondrial DNA, transcription factors, and other
biomarkers for mitochondrial effects. Dietary PFOA also was demonstrated to uncouple
oxidative phosphorylation in mitochondria of the liver from rats exposed via their diet (Keller et
al. 1992). At high concentrations, PFOA caused a small increase in resting respiration rate and
slight decreases in the membrane potential. The observed effects were attributed to a slight
increase in nonselective permeability of the mitochondria membranes caused by PFOA's
surface-active properties (Starkov and Wallace 2002). Quist et al. (2015) found evidence of
mitochondrial proliferation in the liver of CD-I mice pups from dams exposed to 1 mg/kg/day
during gestation and lactation when tissues were examined using transmission electron
microscopy at PND 21 and 91.
3.4.4 Weight of Evidence Evaluation for Carcinogenicity
The findings for cancer in humans provide support for an association between PFOA and
kidney and testicular cancers; however, the number of independent studies examining each of
these is limited. The support comes from high-exposure community studies examining cancer
incidence and covering children and young adults (Barry et al. 2013; Vieira et al. 2013); there is
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some overlap in the cases included in these studies. The two occupational cohorts in Minnesota
and West Virginia (most recently updated in Raleigh et al. 2014 and Steenland and Woskie
2012) do not support an increased risk of kidney or testicular cancer, but are limited by a very
small number of observed cases. None of the general population studies examined these cancers,
but associations were not seen in the general population studies addressing colorectal, breast,
prostate, bladder, and liver cancer, with mean serum PFOA levels up to 0.0866 |ig/mL
(Bonefeld-J0rgensen et al. 2014; Eriksen et al. 2009; Hardell et al. 2014; Innes et al. 2014).
The only chronic bioassays of PFOA were conducted in rats (Butenhoff et al. 2012; Biegel et
al. 2001). The two studies support a positive finding for the ability of PFOA to be tumorigenic in
one or more organs of male, but not female, rats. There are no carcinogenicity data from a
second animal species. There are some data that provide support for the hypothesis that the
PPARa agonism MOA is wholly or partially linked to each of the observed tumor types. The
data support a PPARa MOA for the liver tumors and thus are indicative of lack of relevance to
humans. PPARa activation also could play a role in the other tumor types observed, but more
data to support intermediate steps in the proposed MO As are needed.
The mutagenicity data on PFOA are largely negative, although there is some evidence for
clastogenicity in the presence of microsomal activation and at cytotoxic concentrations. Given
the chemical and physical properties of PFOAincluding the fact that it is not metabolized,
binds to cellular proteins, and carries a net negative electrostatic surface chargethe clastogenic
effects are likely the result of an indirect mechanism. PFOA has the potential to interfere with
the process of DNA replication because of its protein binding properties and the fact that histone
proteins, spermine and spermidine, carry a net positive surface charge. Involvement of ROS in
the MOA as a result of PFOA alone is unlikely because of its metabolic stability. Conditions
leading to ROS would be a function of metabolic responses perturbed by PFOA rather than
PFOA alone. A compound that is not metabolized will not be able to covalently alter the
structure of DNA or intercalate because of electrostatic repulsion between the aromatic base pi
bond electrons with the partial negative charges on the PFOA fluoride atoms. Due to its protein
binding properties, PFOA could have an impact on one or more of the proteins involved in the
process of DNA replication or cell division (cytoskeletal proteins); however, no mechanistic
studies were identified that examined the biochemical effects of PFOA on DNA replication or
cell division. There are no data that support the clastogenic MOA.
Despite the limitations in the data for the LCTs and PACTs, under the U.S. EPA Guidelines
for Carcinogen Risk Assessment (USEPA 2005a) there is suggestive evidence of carcinogenic
potential of PFOA in humans.
3.4.5 Potentially Sensitive Populations
Human biomonitoring studies do not suggest major differences between serum PFOA levels
in males and females. However, the worker populations that are those most likely to demonstrate
such differences because of their higher exposures were predominantly male.
Some animal species have gender differences that affect toxicity of PFOA. Sexually mature
female rats excreted almost all of a 10-mg/kg dose of PFOA within 48 hours compared to only
19% excreted by male rats. Male hamsters excrete PFOA faster than female hamsters, and
female rabbits excrete PFOA slightly faster than male rabbits. Male and female mice excrete
PFOA at approximately the same rate (Hundley et al. 2006). Studies of the transporters involved
in the toxicokinetics of PFOA demonstrate that they are differentially impacted by the presence
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of male and female sex hormones influencing tissue persistence (Cheng et al. 2006; Kudo et al.
2002). As studied in rats (Kudo et al. 2002), the male sex hormones increased half-life
(decreased excretion) of PFOA while the female hormones were associated with shorter half-
lives (increased excretion). The gender differences in mice are not as pronounced as those in rats.
Work by Cheng et al. (2006) and Cheng and Klaassen (2009) demonstrated that these hormones
impact transporters in the liver and kidney.
In studies in which both male and female rats were used, the males were more sensitive to
toxicity than were the female rats (Butenhoff et al. 2004a). Mice displayed similar sensitivities
following PFOA exposure (Kennedy 1987). In the monkey studies, the number of animals per
gender per dose group was too small to reveal a difference related to gender.
Unfortunately, much work remains to be done to determine whether the gender difference
seen in rats is relevant to humans. Similarities are possible because the long half-life in humans
suggests that they might be more like the male rat than the female rat. There is a broad range of
half-lives in human epidemiology studies, suggesting a variability in human transport and
binding capabilities resulting from genetic variations in transporter structures and, consequently,
in function. Genetic variation in human OATs and OATPs has been identified as described in a
review by Zai'r et al. (2008).
Neonates, Infants, and Fetuses
The developing fetus might be sensitive to effects of PFOA. The observed effects on birth
weight in animals are supported by evidence of an association between PFOA and low birth
weight in humans (Johnson et al. 2014). There is some uncertainty related to the interpretation of
the small change in birth weight observed in humans. Specifically, it has been suggested that low
GFR also can impact birth weight (Morken et al. 2014). Verner et al (2015) conducted a meta-
analysis based on PBPK simulations and found that, in individuals with low GFR, there are
increased levels of serum PFOA as well as lower birth weights. Thus, while there is some
uncertainty in the interpretation of the observed association between PFOA and low GFR and
birth weight given the available information, the data indicate that PFOA exposure does impact
birth weight in the general population. In humans with low GFR (which includes females with
pregnancy-induced hypertension or preeclampsia) who also are exposed to PFOA, the effect on
body weight is likely due to a combination of both.
Several animal studies have examined potential MoAs for developmental effects following
prenatal exposure to PFOA. PFOA exposure during development in rats and mice resulted in
increased resorptions (mouse), increased fetal skeletal variation (rat, mouse), decreased neonatal
survival (rat, mouse), decreased postnatal body weight (mouse), delayed eye opening and body
hair growth (mouse), delayed vaginal opening (mouse), accelerated preputial separation (mouse),
and delayed mammary gland development (mouse dam and offspring) (Abbott et al. 2007;
Butenhoff et al. 2004a; Lau et al. 2006; Macon et al. 2011; Tucker et al. 2015; White et al. 2007,
2009, 2011; Wolf et al. 2007). Other long-term effects observed in the surviving offspring
included increased body weight gain, serum leptin, and serum insulin levels along with changes
in adipose tissue (Hines et al. 2009). The MO As for these developmental effects are unknown,
but several potential MoAs have been investigated.
Wolf et al. (2007) restricted mouse prenatal PFOA exposures to 3-11-day periods during
gestation to determine if PFOA was affecting a certain stage of organogenesis resulting in the
observed developmental effects. Decreased postnatal survival was observed at the highest dose
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used (20 mg/kg/day). Eye opening and body hair growth were delayed in offspring exposed for
the longest periods of time (GD 7-17 and GD 10-17) and might have been the result of a higher
cumulative dose or greater sensitivity during early gestation. A cross-fostering paradigm was
used to determine if the developmental effects were the result of gestational exposure, lactational
exposure, or a combination of both. Postnatal survival was decreased in offspring exposed
through gestation and lactation (5 mg/kg/day). Eye opening and body hair growth were delayed
and body weight was reduced in offspring exposed during gestation (5 mg/kg/day), and gestation
and lactation (3 and 5 mg/kg/day). No developmental delays in eye opening and body hair
growth were observed in offspring exposed via lactation only, indicating that, for these
developmental endpoints, PFOA alters growth regulation in the developing fetus that persists as
growth continues postnatally.
Both gestational and lactational exposures contribute to the impact of PFOA on body weight
during early life as illustrated by cross-fostering control unexposed female pups with those dosed
with PFOA. Three cross-fostering combinations were evaluated by White et al. (2009): control
pups nursed by exposed dams, exposed pups nursed by control dams, and exposed pups nursed
by exposed dams. Two doses were evaluated: 3 and 5 mg/kg/day, but the body weight data was
only provided for the 5-mg/kg/day dose group for PND 1-10. PFOA exposures significantly
reduced pup body weights and increased liver weights. The body weight deficits compared to
control were greatest for the gestation and lactation exposure combination and lowest for the
lactation-only group.
Abbott et al. (2007) examined activation of PPARa as a factor in the developmental toxicity
of PFOA. Wild-type and PPARa-null mice experienced full litter resorptions following
gestational (GD 1-17) PFOA exposure (>5 mg/kg/day), indicating that the mechanism of PFOA-
induced resorptions was independent of PPARa expression. These resorptions could be due to
insufficient trophoblast cell type differentiation and/or increased trophoblast cell necrosis (Suh et
al. 2011). Postnatal survival was significantly decreased in wild-type offspring but not in
PPARa-null offspring, indicating that PPARa expression was required for postnatal lethality
(Abbott et al. 2007). Eye opening was significantly delayed in wild-type offspring, but not in
PPARa-null offspring, although a trend was observed in those offspring for later eye opening.
The results indicated that PPARa expression was important for eye opening, but other PPARa-
independent factors also might play a role in its mechanism. Takacs and Abbott (2007) showed
that PFOA can activate mouse PPAR.p/5, which is expressed in developing tissue, and suggested
that inappropriate activation of PPARp/5 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 5 mg PFOA/kg/day on GD
1-17 and GD 8-17 had significantly delayed mammary gland development (full of alveoli,
visible adipose tissue, not well differentiated) at PND 10, which is at the peak of lactation in
rodents. The delayed dam mammary gland development could play a role in the observed
reduced offspring body weight gain if the quantity or quality of the milk is altered by PFOA
(Abbott et al. 2007; Lau et al. 2006; White et al. 2007; Wolf et al. 2007).
Restricted gestational exposure and cross-fostering studies showed that delayed offspring
mammary gland development observed PND 1-63 occurred regardless of exposure duration or
timing (gestation versus lactation exposure; maternal dose of 1 mg/kg/day). The developmental
delays persisted even as the internal PFOA dose decreased (Macon et al. 2011; White et al. 2007,
2009, 2011). More studies need to be conducted to elucidate the MOA for dam and offspring
Perfluorooctanoic acid (PFOA) - May 2016 3-161
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mammary gland effects and its potential functional consequences for lactating humans. White et
al. (2011) conducted a multigeneration study of the effects of PFOA on mammary gland
development and found no dose-related effects on the pup body weights nourished by dams with
lower mammary gland scores than the controls. Tucker et al. (2015) demonstrated that a dose-
response for developmental mammary gland effects varies by more than an order of magnitude,
depending on the strain of mouse studied. CD-I mice are more sensitive than C57BL/6 mice
(Tucker etal. 2015).
Mammary gland development also was affected by peripubertal exposure to PFOA (C. Yang
et al. 2009, Y. Zhao et al. 2010). Low doses (5 mg/kg/day) of PFOA from 3 to 7 weeks of age
caused accelerated mammary gland development in C57BL/6 mice, but delayed mammary gland
development in BALB/c mice, suggesting strain-related differences.
Experiments examining the mechanism for accelerated mammary gland development showed
that PFOA promotes steroid hormone production in the ovaries and increases the levels of
several mammary gland growth factors in C57BL/6 wild-type and PPARa-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-I mice
resulted in significant weight gain and increased serum insulin and leptin levels of the offspring
in mid-life. The increased leptin levels, as well other hormone perturbations, might place PFOA
into the environmental endocrine disrupter obesogen category similar to diethylstilbestrol
(Newbold et al. 2007). However, in a study by Quist et al. (2015) using the mature animals from
the Hines et al. study (2009), there was no dose-related impact on serum leptin in CD-I pups
gestationally exposed across a dose range of 0-1 mg/kg/day when examined on PND 91, except
in the group given an HFD and not fasted before serum collection. For those animals, there was a
dose-related decrease in leptin. Other mice on an HFD that were fasted for 4 hours before serum
collection in the same study lacked a dose-related leptin response. In humans, increased leptin
levels are associated with increased body fat and suggestive of a leptin-resistance MOA for being
overweight (Considine et al. 1996). A similar relationship might occur in prenatally PFOA-
exposed mice; however, the Quist et al. study (2015) suggests that the fat content of the diet and
the time of serum collection are important variables that need to be considered. Studies
determining MO As need to be conducted to determine relevance of the mammary gland effects
to animal and human health.
Diet might influence the risk associated with PFOA exposures. Animal studies demonstrate
an increased risk for liver steatosis in animals on an HFD (Quist et al. 2015; Tan et al. 2013) and
possibly for insulin resistance (Hines et al. 2009). The epidemiology data are not supportive of a
correlation with insulin resistance, but the observations of elevated serum triglycerides,
especially among a highly exposed population, could be viewed as a risk factor for steatosis.
Most of the epidemiology studies did not evaluate dietary factors as part of the study design for
either birth weight or serum lipids (e.g., cholesterol, triglycerides, LDL).
Perfluorooctanoic acid (PFOA) - May 2016 3-162
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4 DOSE-RESPONSE ASSESSMENT
4.1 Dose-Response for Noncancer Effects
An RfD or reference concentration (RfC) is used as a benchmark for the prevention of 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, and so forth. The RfD is expressed in terms of mg/kg/day and the RfC is expressed in
milligrams per cubic meter (mg/m3). The RfD and RfC are estimates (with uncertainties
spanning perhaps an order of magnitude) of the daily exposure to the human population
(including sensitive subgroups) that is likely to be without an appreciable risk of deleterious
effects during a lifetime.
4.1.1 RfD Determination
The derivation of the RfD for PFOA presented a number of challenges due to the
toxicokinetic complexity of PFOA, variability in half-life between species, and metabolic
inertness of PFOA in living organisms. The toxicokinetic features of PFOA lead to differences in
half-lives across species and in the case of rats, and possibly humans, differences between
genders. Toxicokinetics also influence intraindividual and lifestage variability in response to
dose. Additionally there were inconsistencies across the epidemiology studies and the effects
observed in animal studies, and a number of animal studies lacked a NOAEL. Each of these
factors highlights the importance of having measures of internal dose for quantification of an
RfD and supports the utilization of a PK model as a component of the dose-response assessment.
Human Data. Key studies examined occupational and residential populations at or near 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 database. Health outcomes assessed include blood lipid and clinical
chemistry profiles, reproductive parameters, thyroid effects, diabetes, immune function, birth and
fetal and developmental growth measures, and cancer.
Epidemiology studies examined workers at PFOA production plants, a high-exposure
community population near a production plant in the United States (i.e., the C8 cohort), and
members of the general population in the United States, Europe, and Asia. These studies
examined the relationship between serum PFOA concentration (or other measures of PFOA
exposure) and various health outcomes. Exposures in the highly-exposed C8 community are
based on the concentrations in contaminated drinking water and serum measures. Exposures
among the general population typically included multiple PFAS as indicated by serum
measurements. The correlation among these compounds is often moderately strong
(e.g., Spearman r > 0.6 for PFOA and PFOS in the general population). Mean serum levels
among the occupational cohorts ranged approximately 1-4 |ig/mL and in the C8 cohort they
ranged 0.01-0.10 jig/mL. Geometric mean serum values for the NHANES general population
(> age 12; 2003-2008) were 0.0045 ug/mL for males and 0.0036 ug/mL for females (Jain 2014).
These studies have generally found positive associations between serum PFOA concentration
and TC (i.e., increasing lipid level with increasing PFOA) in the PFOA-exposed workers at mean
serum levels 0.4 to >12 |ig/mL and the high-exposure community at mean serum about
0.08 |ig/mL; similar patterns are seen with LDLs but not with FtDLs. These associations also
Perfluorooctanoic acid (PFOA) - May 2016 4-1
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were seen in most of the general population studies (mean serum 0.002-0.007 |ig/mL), but
similar results were seen with PFOS and the studies did not adjust for these correlations.
Associations between PFOA exposure and elevations in serum levels of ALT and GGT, were
consistently observed in occupational cohorts, the high-exposure community, and the U.S.
general population at serum PFOA concentrations also associated with increased TC. The
associations are not large in magnitude, but they indicate the potential to affect liver function.
Thyroid disease incidence was associated with PFOA in women and girls in the high-
exposure C8 study population and in women with background exposure at mean serum
concentrations of 0.026-0.123 |ig/mL. Changes in thyroid hormones were not consistently
associated with PFOA concentration.
Associations between PFOA exposure and risk of infectious diseases (as a marker of immune
suppression) have not been found, but a decreased response to vaccines in relation to PFOA
exposure was reported in studies in adults in the high-exposure community population (median
0.032 |ig/mL) and in studies in children in the general population (mean 0.004 |ig/mL). In the
latter studies, it is difficult to distinguish associations with PFOA from those of other correlated
PFAAs. Increased risk of ulcerative colitis was reported in the high-exposure community study
and in a study limited to workers in that population.
Studies in the high-exposure community reported an association between serum PFOA at
approximately 0.01-0.02 |ig/mL and risk of pregnancy-related hypertension or preeclampsia.
This outcome has not been examined in other populations. An inverse association between
maternal PFOA (measured during the second or third trimester) or cord blood PFOA
concentrations and birth weight was seen in several studies, but the magnitude was small. It has
also been suggested that low GFR can impact birth weight (Morken et al. 2014). Verner et al.
(2015) conducted a meta-analysis based on PBPK simulations and found that some of the
association reported between PFOA and birth weight could be partially attributable to low GFR.
However, the study authors demonstrated that in individuals with low GFR there also are
increased levels of serum PFOA. Thus, while there is some uncertainty in the interpretation of
the observed association between PFOA and low GFR and birth weight given the available
information, the data indicate that PFOA exposure does impact birth weight in the general
population.
The epidemiology studies have not found associations between PFOA and diabetes,
neurodevelopmental effects, or preterm birth and other complications of pregnancy.
Developmental outcomes including delayed puberty onset in girls has been reported; however, in
the two studies examining prenatal PFOA exposure in relation to menarche, conflicting results
were found (i.e., no association or a possible indication of an earlier menarche seen with higher
maternal PFOA levels in one study and a later menarche seen with higher maternal PFOA levels
in the other study).
Animal DataLong Term Studies. Some of the effects in animal studies are associated with
activation of the PPARa receptor leading to peroxisome proliferation. These include increased
liver weight; decreases in serum triglycerides, cholesterol, and lipoproteins; and increases in
ALT, AST, or both. However, although the mechanisms for other effects, such as decreased
body weight, immunological effects, and developmental delays are unknown, they might be
relevant to human health risk assessment.
Perfluorooctanoic acid (PFOA) - May 2016 4-2
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As an initial step in the dose-response assessment, EPA identified a suite of animal studies
with NOAELs, LOAELs, or both that identified the studies as candidates for development of a
chronic RfD. These studies are listed in Table 4-1. The candidate studies were selected based on
their NOAEL, LOAEL, or both; a duration of > 7 weeks; use of a control; and two or more
doses. Table 4-1 does not include the data from human epidemiology studies because, although
they include information on serum levels, they do not identify exposure sources or external
doses.
Table 4-1. NOAEL/LOAEL Data for Oral Subchronic and Chronic Studies of PFOA
Species
Monkey
Male
Monkey
Female
Monkey
Male
Rat
Male
Rat Male
FO
generation
Rat Male
Fl generation
Rat Female
FO generation
Rat Female
Fl generation
Rat
Male and
Female
Study
Duration
90 days
90 days
26 weeks
13 weeks
84 days
16 weeks
127 days
10 weeks
2 years
NOAEL
mg/kg/day
none
3
none
0.64
none
none
30
10
1.3 (m)
1.6 (f)
LOAEL
mg/kg/day
3
10
3
1.94
1
1
none
30
14.2 (m)
16.1 (f)
Critical Effects (s)
t relative pituitary weight
J, absolute and relative heart weight
t absolute liver weight
(hepatocellular hypertrophy) and
mean liver-to-body weight
percentages
t absolute and relative liver weight
with hepatocellular hypertrophy
accompanied by a slight, but not
significant, increase in hepatic
coagulative necrosis
t absolute and relative liver and
kidney weight accompanied by J,
body weight
J, body weights and weight gains; t
absolute and relative liver weights,
liver hypertrophy; f absolute and
relative kidney weights
No significant effects observed
Delay in sexual maturity, J, body
weight and weight gain
M: J, body weight gain;
histopathology lesions in liver, testes,
and lungs.
F: I body weight gain
Reference
Goldenthal 1978
Goldenthal 1978
Butenhoff et al.
2002
Perkins et al.
2004
Butenhoff et al.
2004a; York et
al. 2010
Butenhoff et al.
2004a; York et
al. 2010
Butenhoff et al.
2004a; York et
al. 2010
Butenhoff et al.
2004a; York et
al. 2010
Butenhoff et al.
2012
When examining the effects associated with the LOAELs summarized in Table 4-1, changes
in relative liver weight, absolute liver weight, or both appear to be a common denominator for
monkeys and rats (Butenhoff et al. 2002, 2004a; Perkins et al. 2004) with or without other
hepatic indicators of adversity. Serum PFOA levels, where available, associated with increased
liver weight were 81 and 41 |ig/mL for the male monkey and rat, respectively. However, the
increases in liver weight and hypertrophy are effects associated with activation of cellular
PPARa receptors, making it difficult to determine whether this change is totally a reflection of
the PPARa activation or PFOA toxicity and meet the Hall et al. (2012) criteria for establishing
adversity for a PPARa-activating chemical. Studies in PPARa null mice and animals with a
Perfluorooctanoic acid (PFOA) - May 2016
4-3
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human PPARa receptor (Li et al. 2011; Minata et al. 2010; Nakamura et al. 2009; Wolf et al.
2008b), along with studies of hepatic gene activation by PFOA (Albrecht et al. 2013; Bjork and
Wallace 2009; Nakamura et al. 2009; Rosen et al. 2008a, 2008b), suggest that the increase in
liver weight is at least partially due to cellular impacts that are not controlled by PPARa
receptors. However, it remains difficult to separate the impact of PPARa activation from the
direct effects of PFOA in the candidate studies.
According to Hall et al. (2012), increases in liver weight can be considered adverse when
accompanied by cellular necrosis, inflammation, fibrosis of the liver, and/or macrovesicular
steatosis. There was some evidence of hepatic necrosis in the studies of Perkins et al. (2004) and
in the male Fl generation adult rats from the Butenhoff et al. study (2004a), but the incidences
were not statistically significant or described in detail. To the extent that adverse lesions reflect
sensitivity in the animals impacted, they are used in the assessment to reflect that the liver
hypertrophy and increased liver weight are adverse in individual animals where they are
accompanied by necrosis.
Body weight effects were seen in several studies (Butenhoff et al. 2004a, 2012) and are a
more toxicologically-relevant endpoint, especially in the cases where they were not accompanied
by decreased food intake and when found in neonates (Butenhoff et al. 2004a). There were
developmental delays for males and females in the two-generation study published by Butenhoff
et al. (2004a). Testicular effects were observed by Butenhoff et al. (2012) and in the chronic one-
dose study by Biegel et al. (2001). There was evidence of increased kidney weight in male Fl
Sprague-Dawley rats (Butenhoff et al. 2004a) confounded by decreases in body weight at higher
doses, but at lower doses the kidney weight effect is likely a reflection of tissue adaptation as a
result of the requirement for upregulation of tubular transporters to facilitate urinary excretion
using transporters developed for excretion of endogenous and dietary substrates rather than
PFOA.
Four of the longer term studies in Table 4-1 lack a NOAEL and have LOAELs that range
1-3 mg/kg/day. The NOAELs for the remaining 7 studies range from 0.64 (male rats) to
30 mg/kg/day (female rats). Male monkeys and rats appear to respond at doses that are lower
than their female counterparts. No long-term studies in mice were identified. Since NOAELs and
LOAELs are to some extent the product of concentration or dose level selection, examination of
the dose information in Table 4-1 suggests that several of the data sets that have serum data to
inform modeling of internal doses have the potential to be co-critical in the dose-response
evaluation.
Animal DataShort Term Studies. A number of studies identified adverse effects following
low dose exposures over durations of 7 to 38 days. The studies fall into two clusters, those
evaluating developmental or reproductive effects and those with a focus on immunological
effects. The critical shorter-term studies in rats and mice are summarized in Table 4-2. Although
the exposure duration is shorter in developmental studies than the two-generation study, the
developmental studies are important in quantification of dose-response because the exposures
occur during critical windows of development and predicate effects that can occur later in life.
Perfluorooctanoic acid (PFOA) - May 2016 4-4
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Table 4-2. Shorter-term and Developmental Oral Exposure Studies
Species
Study
Duration
NOAEL
(mg/kg/day)
LOAEL
(mg/kg/day)
Critical Effect(s)
Reference
Rat
Male
Male
29 days
14 days
1
0.2
10
2
Increased absolute and relative
liver weight, focal liver necrosis
t liver weight, t serum estradiol
and hepatic aromatase
Loveless et al.
2008
Liuetal. 1996
Mouse
Female
offspring
Male&
Female
Female
Male
Male
Female
Female
Female
CD1
Female CD-I
Female
C57BL6
Female
17 days
38 days
38 days
29 days
28 days
17 (pups)
718 (dams)
days
17 days
CDs 1-17
17 days
CDs 1-17
17 days
17 days
15 days
none
1
0.01
0.3
none
none
none
none
none
0.1
1.88
1
5
0.03
1
5
1
3
0.3
0.01
0.3
3.75
Delayed mammary gland
development in dams during
lactation
t Ito cell hypertrophy at 18
months. Dosing occurred during
gestation and lactation only.
t TC at PND 91 for fasted and
nonfasted animals receiving a HFD
but not those receiving the
standard fat content control diet.
Exposure occurred only during
gestation and lactation.
t absolute and relative liver
weight, I relative spleen weight,
moderate-severe liver hypertrophy
with single cell and focal necrosis
Significantly J, fertility based on
pregnant females per male mouse,
and J, pup birth weight.
t absolute maternal liver weight, J,
ossification (calvarin, enlarged
fontanel), accelerated onset of
puberty in male offspring.
t absolute and relative maternal
liver weight, delayed offspring eye
opening and body hair growth, f
offspring relative liver weight, J,
offspring body weight, delayed
mammary gland development
(female offspring).
Delayed mammary gland
development
Delayed mammary gland
development at PND 56. Exposure
occurred only during gestation.
Delayed mammary gland
development at PND 61. Exposure
occurred only during gestation.
I IgM (1 day post-dose), increased
IgG (15 days post-dose), ^absolute
and relative spleen weight (1 day
post-dose)
White et al.
2011
Filgo et al.
2015
Quist et al.
2015
Loveless et al.
2008
Luetal. 2015
Lau et al. 2006
White et al.
2009; Wolf et
al. 2007
Macon et al.
2011
Tucker et al.
2015
Tucker et al.
2015
DeWitt et al.
2008
Perfluorooctanoic acid (PFOA) - May 2016
4-5
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Species
Male
Male
Male
Female
Female CD 1
Female CD 1
Study
Duration
15 days
14 days
14 days
1 1 days
CDs 7-17
8 days
CDs 10-17
17 days
CDs 1-17
NOAEL
(mg/kg/day)
7.5
2.5
0.2
none
none
none
LOAEL
(mg/kg/day)
30
5
2
5
0.01
0.3
Critical Effect(s)
I sheep red blood cell IgM
response in PPAR null mice
indicate the response not
completely PPAR dependent
I sperm count, changes in
testicular morphology, evidence of
t free radical oxidation
t liver weight, serum estradiol and
hepatic aromatase activity
t maternal and pup relative liver
weight, delayed offspring eye
opening and hair growth, J, male
offspring body weight, delayed
mammary gland development
(female offspring)
Delayed mammary gland
development on PND 2 1 (female
offspring)
Delayed mammary gland
development on PND 14 (female
offspring)
Reference
DeWitt et al.
2015
Liu etal. 2015
Liu etal. 2015
White et al.
2009; Wolf et
al. 2007
Macon et al.
2011
Macon et al.
2011
All but two of the short term studies used mice as the target species. Mice differ from rats in
that the toxicokinetics of the males and females are similar. The half-life of PFOA in male rats is
much longer than that in females, favoring higher serum levels in males after equivalent
exposures. The difference in the excretion kinetics is a consequence of differences in renal
transporters between male and female rats that appear to be under hormonal control. Several of
the short term studies include serum data to support PK modeling of internal dose-response
(DeWitt et al. 2008; Lau et al. 2006; Macon et al. 2011).
As was the case with the longer-term studies, increased liver weight was a common effect
among the shorter-term studies. Increases in absolute or relative liver weights were reported in
six of the studies that provided dose-response data from short term exposures (Lau et al. 2006;
Liu et al. 1996, 2005; Loveless et al. 2008; White et al. 2009; Wolf et al. 2007) (Table 4-2). In
some of the remaining studies, liver weight was not monitored as a variable. The Loveless et al.
study (2008) identified significant focal liver necrosis in rats at the 10 mg/kg/day LOAEL, and
both single cell and focal liver necrosis in mice at a LOAEL of 1 mg/kg/day. This might indicate
that mice are more susceptible to necrosis than rats. Hepatic necrosis was reported in the longer
duration Perkins et al. study (2004) of male rats and in the male Fl generation adult rats from the
Butenhoff et al. study (2004a), but hepatic necrosis was present in few animals and not evaluated
for statistical significance.
The co-occurring effects at the LOAEL were effects on spleen, thymus, liver, and/or
developmental endpoints. Four of the studies involved exposures that occurred only during
gestation and lactation and resulted in effects that were observed in the mature offspring. The
hepatic and serum cholesterol effects in Quist et al. (2015) at PND 91 at a LOAEL of
0.03 mg/kg/day were present only in animals with elevated intakes of dietary fat. In adult
animals with the same gestation/lactation only exposures, Filgo et al. (2015) identified a LOAEL
of 5 mg/kg/day for accumulation of fat deposits in the liver Ito cells (steatosis). The study did not
Perfluorooctanoic acid (PFOA) - May 2016
4-6
-------
provide information on the intakes of dietary lipid that could be compared with the data from
Quist et al. (2015) to determine whether the effects were correlated with postnatal dietary fats or
from the exposure during gestation and lactation.
A NOAEL of 2.5 mg/kg/day and a LOAEL of 5 mg/kg/day was reported for sperm counts
and testicular morphology after a!4-day exposure by Liu et al. (2015). No impact on male
fertility was observed in Sprague-Dawley rats in the two-generation Butenhoff et al. study
(2004a), in contrast to the Lu et al. study (2015) where male fertility was decreased in mice at a
dose of 5 mg/kg/day for 28 days. A 14-day exposure to 2 mg/kg/day (Liu et al. 2015) led to
significantly increased serum estradiol and increased hepatic aromatase activity.
The developmental impacts of PFOA exposure ranged from delayed mammary gland
development in pups (Albrecht et al. 2013; Macon et al. 2011; Tucker et al. 2015; White et al.
2009, 2011; Wolf et al. 2007) to delays in attaining developmental milestones such as
ossification, eye opening, and hair growth (Lau et al. 2006; Wolf et al. 2007). The LOAEL for
developmental effects on mammary glands in female offspring from dams given 0.01 mg/kg/day
for 8 days from Macon et al. (2011) is of unknown biological significance. In the same study, no
effects on offspring body weight were found at maternal doses up to 3 mg/kg/day for 17 days
(Macon et al. 2011). Data from White et al. (2011) showed no significant effects on body weight
gain in pups nursing from dams treated with 1 mg/kg/day despite these dams having less fully-
developed mammary glands compared to controls. Similarly, no differences in response to
lactational challenge were seen in PFOA-exposed dams with morphologically delayed mammary
gland development (White et al. 2011).
The studies that looked at the delays in other developmental milestones including eye
opening, hair growth, and bone ossification all lacked a NOAEL. In Lau et al. (2006), the
LOAEL was 1 mg/kg/day for reduced ossification of the proximal phalanges (forelimb and
hindlimb). In the Wolf et al. (2007) study, delays in eye opening and hair growth occurred at a
LOAEL of 5 mg/kg/day for gestational exposures of both 1-17 days and 7-17 days. Attainment
of sexual maturity in males from the Lau et al. study (2006), rather than being delayed, was
accelerated, at the LOAEL of 1 mg/kg/day.
The LOAEL for the mammary gland developmental effects in female offspring from dams
given 0.01 mg/kg/day for 8 days from Macon et al. (2011). In the same study, no effects on
offspring body weight were found at maternal doses up to 3 mg/kg/day for 17 days (Macon et al.
2011). Data from White et al. (2011) showed no significant effects on body weight gain in pups
nursing from dams treated with 1 mg/kg/day despite these dams having less fully-developed
mammary glands compared to controls. Similarly, no differences in response to lactational
challenge were seen in PFOA-exposed dams with morphologically-delayed mammary gland
development (White et al. 2011). Given that milk production was sufficient to nourish growth in
exposed pups, there is uncertainty related to the functional impact of this endpoint and thus it
was not considered quantitatively.
A NOAEL of 2.5 mg/kg/day and a LOAEL of 5 mg/kg/day were reported for sperm counts
and testicular morphology after a 14-day exposure by Liu et al. (2015). No impact on male
fertility was observed in Sprague-Dawley rats in the two-generation Butenhoff et al. study
(2004a) in contrast to the Lu et al. study (2015) where male fertility was decreased in mice at a
dose of 5 mg/kg/day for 28 days. A 14-day exposure to 2 mg/kg/day (Liu et al. 2015) lead to
significantly-increased serum estradiol and increased hepatic aromatase activity.
Perfluorooctanoic acid (PFOA) - May 2016 4-7
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The studies by DeWitt et al. (2008, 2015) demonstrate effects of PFOA on spleen and
thymus weights and the immunoglobulin response to SRBC or dinitrophenyl ficol in wild-type
and PPARa null mice. The DeWitt et al. (2015) data indicate that some but not all of the
response is related to PPARa activation. As supported by the epidemiology data, suppression of
the immune system response to a challenge because of PFOA exposure is an area of concern for
humans as well as animal species.
Six of the twelve studies lacked a NOAEL. For those studies with a NOAEL, the value
ranged 0.01-7.5 mg/kg/day, while the LOAELs ranged 0.01-30 mg/kg/day. The range of values
across studies is narrow, with overlap between the NOAELs and LOAELs. In all instances, the
durations of exposure in shorter-term studies were less than 39 days, suggesting that
physiological responses to PFOA occur early in the exposure continuum and at doses, but not
necessarily average serum levels, comparable to those observed in the long term studies.
4.1.1.1 PK Model approach
In linking chemical exposure to toxic endpoints, careful consideration of PKs is crucial. This
is especially true for PFOA, where inter-species and gender variation in CL half-life can vary by
several orders of magnitude. If the toxicological endpoints are assumed to be driven by internal
concentrations, the internal exposure needs to be calculated and considered across species.
Differences in PKs (e.g., male rats excrete PFOA more slowly than females) and differences
across species produce differences in the external dose needed to achieve the same internal dose.
The use of the animal data and the available PK model allows for the incorporation of species
differences in saturable renal resorption, dosing duration, and serum measurements for doses
administered to determine HEDs based on average serum concentration and CL.
Candidate studies for RfD development were restricted to those of adequate duration
(preferably > 7 weeks), multiple dose groups, use of a concurrent control, and with serum data
amenable for modeling that showed the most sensitive effects following exposure to PFOA.
Those studies included the subchronic study by Perkins et al. (2004), the two-generation study by
Butenhoff et al. (2004a), both conducted in rats, and the Butenhoff et al. study (2002) in
monkeys. Also included are the developmental studies of Lau et al. (2006) and Wolf et al.
(2007)/White et al. (2009), and the DeWitt et al. study (2008) of immunotoxicity in mice that
showed effects of lifetime concern despite their briefer exposure durations. Together these
studies encompassed the range of doses evaluated and the LOAELs observed in other studies that
lacked serum data.
The Butenhoff et al. study (2002) was included as it is the only longer-term study in a
nonhuman primate and had serum PFOA data available. The dose of 3 mg/kg/day was a LOAEL
for increased liver weight in the absence of clear adverse effects. The small number of animals
per group (2-4) made evaluation of accompanying liver effects difficult to evaluate against the
Hall et al. (2012) criteria. Thus, while included in the model for comparison of serum levels, data
from Butenhoff et al. (2002) will not be considered further in the quantitative RfD assessment.
PFOA has dose-dependent kinetics. Although repeated doses rapidly result in quasi-
equilibrium blood concentrations, a single dose results in a much longer half-life than is
consistent with a rapid approach to quasi-equilibrium (Andersen et al. 2006; Lou et al. 2009).
Using a simple, linear PK model (e.g., a one- or two-compartment model) to predict internal
dose resulted in estimated HEDs from repeated exposures that were greater than the external
doses supported by the experimental data (Butenhoff et al. 2004a; Lou et al. 2009). Application
Perfluorooctanoic acid (PFOA) - May 2016 4-8
-------
of a saturable renal resorption model (Andersen et al. 2006) predicted average serum values at
the time of sacrifice and the duration necessary to reach steady state.
PK data (serial blood concentrations following treatment with known quantities of PFOA)
were collected for three species: cynomolgus monkey (Butenhoff et al. 2004b), Sprague-Dawley
rat (Kemper 2003), and mice. Data were available for two strains of mice and these were
analyzed separately: CD1 (Lou et al. 2009) and C57BL6/N (DeWitt et al. unpublished, cited in
DeWitt et al [2008]). Due to the pronounced difference in the PKs of male and female rats, the
two genders were fit separately. For mice, only female data were used. For monkeys a limited
amount of female data was used jointly with male data, assuming the only difference between the
genders for monkey was bodyweight.
For each study with a toxicological endpoint and LOAEL, the AUC and final serum
concentrations were determined for the exposure duration investigated in that study. These
values are summarized in Table 4-3 for rats, Table 4-4 for mice, and Table 4-5 for monkeys. In
order to make a rough assessment of the validity of the model predictions, a final serum
concentration was predicted for each treatment so that it could be compared to measured serum
values. The predicted final serum concentration is the estimate for serum concentration at the
time of sacrifice. They differed by a factor of four when strains were different and closer to a
factor of two when predicted using parameters from the same strain. Because these predictions
do not perfectly match the measured serum concentrations, there remains uncertainty about the
exposure estimates, and this uncertainty has not been fully characterized.
Table 4-3. Predicted Final Serum Concentration and Time-Integrated Serum
Concentration (AUC) for Studies in Rats
Study
Perkins et
al. 2004
Butenhoff
et al. 2004a
Species/
Strain
Rat (M)
ChR-CD
Rat (M)
Sprague-
Dawley
Exposure
Duration
13 weeks
Diet
FOM: 10
wkpre
mating-
mating
Gavage
Oral
Doses
mg/kg/day
0.06
0.64
1.94
6.50
1
3
10
30
Measured
Final Serum
value
Hg/ml
7.1(1.2)
41(13)
70 (16)
138 (34)
NT
NT
51.5s
45.3
Species/
Strain Used
for
Prediction
Rat (M)
Sprague-
Dawley
Rat (M)
Sprague-
Dawley
Predicted
Final Serum
Value
Hg/ml
3.8 (0.0955)
34.8 (0.865)
79.5 (3.84)
139(13.1)
49.9(1.53)
102 (6.5)
153 (17.3)
169 (27.7)
Predicted AUC
mg/L*h
7230(181)
69100 (1540)
168000 (6520)
326000 (27100)
92500 (2600)
204000 (10900)
345000 (34200)
412000 (70500)
Notes: Numbers in parentheses indicate SD
M = male; s = serum; NT = not tested
Since the Kemper (2003) data were not tied to toxicological endpoints and were only used in model development, they are not
included in this table.
Perfluorooctanoic acid (PFOA) - May 2016
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Table 4-4. Predicted Final Serum Concentration and Time-Integrated Serum
Concentration (AUC) for Studies in Mice
Study
White et al.
2009; Wolf
et al. 2007
DeWitt et
al. 2008
Lau et al.
2006
Species/
Strain
Mouse (F)
CD-I
Mouse (F)
C57BL/6N
Mouse (F)
CD-I
Exposure
Duration
days
CDs l-17a
Gavage
CDs 7-17
Gavage
15 days
Drinking
water
CDs 1-17
Gavage
Oral
Doses
mg/kg/d
3
5
5
0.94
1.88
3.75
7.5
15
30
1
o
J
5
10
20
40
Measured
Serum
Value
Hg/ml
NT
NT
24.8
NTb
NTb
35.3
42.8
50.0
162.6
21. 9c
40.5 c
71.9 c
116 c
181 c
271 c
Species/
Strain Used
for
Prediction
Mouse (F)
CD-I
Mouse (F)
CD-I
Mouse (F)
C57BL/6N
Mouse (F)
CD-I
Predicted
Final Serum
Value fig/ml
25 (2.22)
25.6 (2.26)
29 (2.55)
29.7(1.58)
51.9(1.89)
70.2 (2.57)
81.4(3.91)
94.7(11.8)
117(29.3)
57.6 (3.82)
87.2 (7.93)
95.2(7.41)
106 (5.84)
121(11)
148 (30.2)
Predicted
AUC mg/L*h
33,700 (1,860)
40,700(2,170)
25,400 (1,320)
7,300 (541)
13,800(951)
22,400 (1,290)
30,500 (1,540)
40,100(4,720)
56,000 (12,300)
16,400 (606)
33,600 (1,930)
40,700(2,180)
49,600 (1,980)
61,400 (5,050)
80,100 (12,700)
Notes: Numbers in parentheses indicate SD
NA = not applicable; could not be determined
F = female; GD = gestation day; NT = not tested
a Sacrificed on PND 22.
bDeWitt et al. (2008) had 0.94 and 1.88 mg/kg/day dose groups in a second experiment.
0 The Lau et al. (2006) serum data were provided by the author for animals treated GDs 1-17.
Table 4-5. Predicted Final Serum Concentration and Time-Integrated Serum (AUC)
in Studies of Monkeys
Study
Butenhoff et
al. 2002,
2004b
Species/
Strain
Monkey
(M)
Cyno-
molgus
Exposure
Duration
26 weeks
Oral capsule
Oral Doses
mg/kg/day
3(n=3)
10 (n = 4)
30/20 (n = 3)
Measured Serum
value fig/ml
117.9(87.6-141)
77.35(55.4-96.5)
283.2 (61.7-489)
Species/
Strain Used
For
Prediction
Monkey
Cyno-
molgus
Predicted
Final
Serum
Value
Hg/ml
89.1 (12.4)
121 (14)
149(31)
Predicted
Exposure (AUC)
mg/L*h
380,000 (50,100)
553,000 (62,800)
710,000 (144,000)
Notes: Numbers in parentheses indicate SD
M = male
The average serum concentration for the LOAEL or NOAEL was determined through
numeric simulation. The average or mean value has the advantage of normalizing the serum
concentration across the exposure durations to generate a uniform metric for internal dose in
situations where the dosing durations varied and serum measurements were taken immediately
prior to sacrifice. The averaged serum concentration is a hybrid of the AUC and the maximum
serum concentration. Compared across studies, PFOA average serum concentration appears to be
a stable reflection of internal dosimetry.
Table 4-6 provides the AUC from the model, the dosing duration from each of the modeled
studies, and the resultant average serum concentration. The data from the monkey study
(Butenhoff et al. 2002, 2004b) were not used because of the small number of animals evaluated
Perfluorooctanoic acid (PFOA) - May 2016
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and the wide variability in the responses. The internal doses associated with the effect levels
(LOAELs) differ by less than an order of magnitude (13.1-96.2 mg/L), while the AUC values
differ by over two orders of magnitude (5,360-380,000 mg/L*h). Given the differences in
external doses, the projected serum levels are proportionally quite similar.
Table 4-6. Average Serum Concentrations Derived from the AUC and the Duration of
Dosing
Study
DeWitt et al. 2008: mice; | IgM
response to SRBC
Lau et al. 2006: mice
reduced pup ossification (m, f),
accelerated male puberty
Perkins et al. 2004: rats; t liver
weight/necrosis
Wolf etal. 2007: mice; CDs 1-17
|Pup body weight3
Wolf et al. 2007: mice; CDs 7-17
|Pup body weight3
Butenhoffetal. 2004a:
J, relative body weight/t relative
kidney weight and t kidney :brain
weight ratio in FO and Fl at sacrifice
Dosing
duration
days
15
17
91
17
11
84
NOAEL
mg/kg/day
(AUC
mg/L*h)
1.88
(13,800)
None
0.64
(69,100)
None
None
None
NOAEL
(Av serum
mg/L)
38.2 (2.63)
None
31.6(0.073)
None
None
None
LOAEL
mg/kg/day
(AUC
mg/L*h)
3.75
(22,400)
1
(16,400)
1.94
(168,000)
3
(33,700)
5
(25,400)
1
(92,500)
LOAEL
(Av serum
mg/L)
61.9(3.58)
38.0(1.4)
77.4 (2.98)
77.9 (4.3)
87.9 (4.57)
45.9(1.29)
Notes: Significance p < 0.05 or< 0.01
m = male; f = female; SRBC = Sheep Red Blood Cell
a serum from pups on PND 22
Table 4-6 identifies serum values of 38, 45.9, and 61.9 mg/L as the lowest concentrations
associated with adverse effects in the Lau et al. (2006), Butenhoff et al. (2004a), and DeWitt et
al. (2008) studies, respectively. Thus, it appears that the LOAELs are roughly consistent across
gender, species, and treatment with respect to average serum concentration. Assuming that MoA,
susceptibility to toxicity, or both do not vary and that PKs alone explains variation, it is
reasonable to expect similar concentrations to cause similar effects in humans.
The Andersen et al. (2006) model, used to make the predictions in Tables 4-3, 4-4, and 4-5
can be solved analytically to predict the steady-state concentration (Css) resulting from a fixed
infusion DR (in units of |imol/h):
DR
Css = free * Qfil
1 +
T
1 rr
Qfil * kT + DR,
Although the assumption of a constant infusion exposure simplifies the actual dose regimen
used, this assumption permits rapid calculation and analysis of Css. Using this equation, one can
calculate a range of Css values for each DR. The range of Css values are derived from the species-
specific combinations of parameters from the Bayesian analysis of the available PK data. This
result for Css depends nonlinearly on DR. The PFOA toxicity studies used discrete, daily doses;
these doses were converted into rates by dividing daily dose by 24 hours. In Table 4-7, the Css is
compared with the average serum concentration predicted. The fraction of Css is calculated,
indicating that the studies range 36%-91% of Css.
Perfluorooctanoic acid (PFOA) - May 2016
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For human exposure to PFOA, one needs to rely on steady-state calculations since there is a
lack of both the sufficient PK and exposure knowledge to make more complicated estimates. The
average serum concentrations of the LOAEL in Table 4-7 are all within roughly one order of
magnitude (12.4-87.9 mg/L).
Table 4-7. Comparison of Average Serum Concentration and Steady-State Concentration
Study
Perkins et al. 2004
Butenhoffetal. 2004a
Wolf etal. 2007; CDs 7-17
Wolf etal. 2007; CDs 1-17
DeWitt et al. 2008
Lau et al. 2006
Dosing
duration
days
91
84
11
17
15
17
LOAEL
mg/kg/day
1.94
1
5
3
3.75
1
Css (mg/L) for
constant infusion
of LOAEL
84.4(3.81)
52.5 (1.72)
107 (6.8)
95.9 (6.73)
84.1(4.5)
67.8 (4.39)
Average Serum
Cone, for Study
(mg/L)
77.4 (2.98)
45.9 (1.29)
87.9 (4.57)
77.9 (4.3)
61.9(3.58)
38(1.4)
Fraction of Css
(Average / Css)
0.913 (0.00746)
0.874 (0.00776)
0.82(0.0117)
0.813 (0.0148)
0.736 (0.0233)
0.561 (0.0277)
Notes: Average serum concentrations from PK simulations of toxicity study treatment regimens and Css were both predicted
using species-specific parameter distributions (i.e., draws from the Markov Chain determined by analyzing the available PK data
for the appropriate species). The number in parentheses is the SD.
The predicted average serum concentrations can be converted into an oral equivalent dose by
recognizing that, at steady state, CL from the body should equal dose to the body. CL can be
calculated if the rate of elimination (derived from the half-life) and the Vd are both known. In
making the calculation based on CL, it is important also to consider whether the exposure to
PFOA has lasted long enough to reach steady state. Four of the endpoints modeled represent
serum values that are greater than 80% of steady state, but none represent steady-state
concentrations. Those endpoints representing lower percentages of steady state require
consideration of the uncertainty resulting from use of a projection that is not representative of a
steady-state concentration (UFs) when establishing an RfD for a chronic lifetime exposure
endpoint.
Measures of half-life in humans have been determined for both workers and the general
population (section 2.6.2). Olsen et al. (2007) gives the human half-life as 3.8 years for PFOA in
an occupationally-exposed U.S. cohort. Bartell et al. (2010) determined a value of 2.3 years
based on the decline in serum levels among members of the general population exposed via
drinking water in the area near the DuPont Works plant in Washington, West Virginia after the
drinking water concentrations decreased. EPA chose to use the Bartell et al. (2010) half-life
value because it is the one most relevant to scenarios where exposures result from ingestion of
contaminated drinking water by members of the general population.
Thompson et al. (2010) gives a Vd of 0.17 L/kgbw (section 2.6.3). The Vd is defined as the total
amount of PFOA in the body divided by the blood or serum concentration. The Vd was calibrated
using human serum concentrations and exposure data from NFLANES and assumes that most
PFOA intake came from contaminated drinking water. The value for Vd was calibrated so that the
model prediction of elevated blood levels of PFOA was consistent with the values from NHANES.
The Vd value determined by Thompson et al. (2010) did not consider PFOA contributions from
sources other than drinking water. However this estimate is not radically different from the 0.198
L/kgbw determined for the monkey in the study by Butenhoff et al. (2004b).
The half-life (ty2) and Vd are utilized to calculate the CL for PFOA according to the following
equation assuming first order kinetics for CL (Medinsky and Klaassen 1996):
Perfluorooctanoic acid (PFOA) - May 2016
4-12
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CL = Vdx (In 2 H-1%) = 0.17 L/kgbwX (0.693 H- 839.5 days) = 0.00014 L/kgbw/day
Where:
Vd = 0.17L/kg
In 2 = 0.693
tvi = 839.5 days (2.3 years x 365 days/year = 839.5 days)
These values combined give a CL of 1.4 x 10"4 L/kg bw/day.
Scaling the derived average serum concentrations (in mg/L) for the NOAELs and LOAELs
in Table 4-6 yields the predicted oral HED in mg/kg/day for each corresponding serum
measurement. The HED values are the predicted human oral exposures necessary to achieve
serum concentrations equivalent to the LOAEL (and NOAEL where available) in the animal
toxicity studies. Note that this scaling assumes linear first-order human kinetics in contrast to the
nonlinear phenomena observed at high doses in animals. It is justifiable at the lower dose
NOAEL and LOAEL concentrations from the animal studies that that demonstrate the first-
order, linear response to dose necessary to calculate CL.
Thus, HED = average serum concentration (in mg/L) x CL
Where:
Average serum is from model output in Table 4-8
CL = 0.00014 L/kg bw/day
Table 4-8. HEDs Derived from the Modeled Animal Average Serum Values
Study
De Witt etal. 2008: mice; |
IgM response to SRBC
Lau et al. 2006: mice
reduced pup ossification
(m,f), accelerated male
puberty
Perkins et al. 2004: rats;
fliver weight/necrosis
Wolf etal. 2007: mice; CDs
1-17
|pup body weight
Wolf, etal. 2007: mice;
CDs 7-17
|pup body weight1
Butenhoff etal. 2004a: |FO
body weight/t absolute and
relative kidney weight
Macon etal. (2011)
GDs 1-17 J,mammary gland
development2
Dosing
duration
days
15
17
91
17
11
84
17
NOAEL
mg/kg/d
1.88
None
0.64
None
None
None
NOAEL
Av serum
mg/L
38.2
31.6
HED
mg/kg/d
0.0053
0.0044
LOAEL
mg/kg/d
3.75
1
1.94
3
5
1
0.3
LOAEL
(Av serum)
mg/L
61.9
38.0
77.4
77.9
87.9
45.9
12.4
HED
mg/kg/d
0.0087
0.0053
0.0108
0.0109
0.0123
0.0064
0.0017
Notes: Significance p < 0.05 or < 0.01
m = male; f = female; SRBC = Sheep Red Blood Cell
1 serum from pups on PND 22
2 serum from pups on PND 7
Perfluorooctanoic acid (PFOA) - May 2016
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4.1.1.2 RfD Quantification
The subset of studies amenable for use in derivation of HED based on average serum
measurements from the PK model is based solely on results from studies that have dose and
species-specific serum values for model input, as well as exposure durations of sufficient length
to achieve values near to steady-state projections or applicable to developmental endpoints with
lifetime consequences following short term exposures.
As explained previously, human data identified significant relationships between serum
levels and specific indicators of adverse health effects but lacked the exposure information for
dose-response modeling. For this reason none of the human studies provided an appropriate POD
for RfD derivation. The pharmacokinetically-modeled average serum values from the animal
studies are restricted to the animal species selected for their low dose response to oral PFOA
intakes. Extrapolation to humans adds a layer of uncertainty that needs to be accommodated in
deriving the RfD.
HED PODs. The HEDs derived from Perkins et al. (2004), DeWitt et al. (2008), Lau et al.
(2006), and Butenhoff et al. (2004a) were each examined as the potential basis for the RfD. Only
Perkins et al. (2004) and DeWitt et al. (2008) identified a NOAEL from which the HED could be
derived. The Lau et al. (2006) and Wolf et al. (2007)/White et al. (2009) LOAEL HEDs included
developmental effects in the offspring as accompanied by the increased liver weight that is an
accepted biomarker for PFOA exposure. These are developmental exposure studies that carry
lifetime consequences for a less-than-lifetime exposure. The Butenhoff et al. study (2004a)
included significant decreased body weight (not confounded by reduced food intake) in FO males
accompanied by increased kidney weight (consistent with the need for renal tubular transport) as
co-critical at the LOAEL. The DeWitt et al. study (2008) has a LOAEL for decreased IgM,
increased IgG, and increased absolute and relative spleen weight after a 15-day exposure.
Table 4-9 provides the calculations for potential RfDs using the HEDs derived from PK
modeling based on the serum values collected at animal sacrifice. The table applies UFs to each
POD and illustrates the array of potential RfD outcomes. Each POD is impacted by the doses
utilized in the subject study, the endpoints monitored, and the animal species/gender studied.
Thus, the array of outcomes, combined with knowledge of the individual study characteristics,
helps to inform selection of an RfD that will be protective for humans.
The potential lifetime RfD values in Table 4-9 differ by about an order of magnitude
(0.00002-0.00015 mg/kg/day) but so do the UFs applied to the POD. These results demonstrate
the ability of the model to normalize the animal data across species, strain, gender, and exposure
duration. The UFs applied in the derivation of the potential RfDs alter the first-order, direct
relationship between the animal serum measurements associated with the animal studies and the
resultant RfD. Accordingly, the resultant RfD cannot be extrapolated to a corresponding human
serum value equivalent to the RfD using the CL value applied when calculating the HED from
the animal serum.
Perfluorooctanoic acid (PFOA) - May 2016 4-14
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Table 4-9. The Impact of Quantification Approach on the RfD Outcomes for the HEDs
from the PK Model Average Serum Values
POD
PK-HEDNOAEL Perkins
rats; t liver weight/necrosis
PK-HEDLOAEL Wolf GD 1 -1 7
mice; |pup body weight
PK-HEDLOAEL Wolf GD 7-1 7
mice; |pup body weight3
PK-HEDNOAEL DeWitt
mice; J, IgM response to
SRBC
PK-HEDLOAEL Lau
mice reduced pup
ossification (m,f),
accelerated male puberty
PK-HEDLOAEL Butenhoff
|FO body weight/t absolute
and relative kidney weight
Value
mg/kg/day
0.0044
0.0109
0.0123
0.0053
0.0053
0.0064
UFH
10
10
10
10
10
10
UFA
3
3
o
J
o
J
o
J
3
UFL
-
10
10
10
10
UFs
-
-
-
10
UFD
-
-
-
UFtotal
30
300
300
300
300
300
Candidate
RfD
mg/kg/day
0.00015
0.00004
0.00004
0.00002
0.00002
0.00002
Notes: m = male; f = female; SRBC = Sheep Red Blood Cell
a serum from pups on PND 22
The Perkins et al. (2004) and Butenhoff et al. (2004a) studies were conducted in male
Sprague-Dawley rats with durations of 91 days via diet and 84 days via gavage, respectively.
Both were associated with increased relative liver weight accompanied by some hepatic necrosis
in a small number of animals. The Butenhoff et al. study (2004a) also observed a significant
decrease in body weight compared to controls for the FO male rats at the end of the 84-day
exposure. The studies by Lau et al. (2006) and Wolf et al. (2007)/White et al. (2009) were
conducted in pregnant female mice with a 17-day average exposure via gavage, resulting in
increased liver weights in the dams and low body weights and developmental delays in offspring.
Uncertainty Value Application
A UF for intraspecies variability (UFn) of 10 is assigned to account for variability in the
responses within the human populations because of both intrinsic (toxicokinetic genetic, life
stage, health status) and extrinsic (life style) factors that can influence the response to dose. No
information to support a UFn other than 10 was available to characterize interindividual and age-
related variability in the toxicokinetics or toxicodynamics among humans other than variability
in serum levels measured among populations residing in common geographical locations with
presumably fairly similar exposures.
A UF for interspecies variability (UFA) of three was applied to account for uncertainty in
extrapolating from laboratory animals to humans (i.e., interspecies variability). The 3-fold factor
is applied to account for toxicodynamic differences between the animals and humans. The FtEDs
were derived using average serum values from a model to account for PK differences between
animals and humans.
A UF for LOAEL to NOAEL extrapolation (UFL) of 10 was applied to all PODs other than
the Perkins et al. study (2004) to account for use of a LOAEL for the POD. The POD for the
Perkins et al. study (2004) is a NOAEL.
Perfluorooctanoic acid (PFOA) - May 2016
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A UF for extrapolation from a subchronic to a chronic exposure duration (UFs) is one,
because the PODs are based on average serum concentrations and determined to represent > 80%
of steady state for each study (81%-91%) except for the Lau et al. (2006) developmental study
(56%). The Lau et al. (2006) developmental HED was not adjusted for lifetime exposures,
because the average serum values associated with the developmental studies are more protective
than those for the longer-term studies of systemic toxicity. A UFs of 10 was applied to the
DeWitt et al. study (2008) serum-derived HED reflecting 73% of steady state because the data
suggest that longer-term exposures to the same dose have the potential to increase serum values
beyond the levels indicated by the 15-day exposure to mice. In addition, the NOAEL for
immunological effects (0.94 mg/kg/day) was a LOAEL for effects on liver weight in the absence
of histological evaluation on both days 16 and 31 following a 15-day exposure. Thus, there is a
potential that lifetime exposures at serum steady state could impact the liver, increasing the risk
for tissue damage.
A database UF (UFo) of one was applied to account for deficiencies in the animal study
database for PFOA. There are extensive human data from epidemiology studies in the general
population, as well as in worker cohorts. The epidemiology data provide strong support for the
identification of hazards observed following exposure to PFOA in the laboratory animal studies
and the human relevance of the critical effects. However, uncertainties in the use of the available
epidemiology data precluded their use at this time in the quantification of an RfD. There are
extensive data from short term, subchronic, chronic, reproductive, developmental, and
mechanistic studies in laboratory animals that support the endpoints used in calculating the
potential RfD. The potential RfD is the one applicable to those most at risk from exposure to
PFOA present in drinking water, the fetus and young infants. The alternative identical RfDs are
values that could be more appropriate for other exposure scenarios and endpoint concerns.
4.1.2 RfD Selection
The candidate RfDs in Table 4-9 range 0.00002-0.00015 mg/kg/day. The RfD of
0.00002 mg/kg/day calculated from HED average serum values from Lau et al. (2006) was
selected. The RfD based on Lau et al. (2006) is derived from reduced ossification of the proximal
phalanges (forelimb and hindlimb) and accelerated puberty in male pups (4 days earlier than
controls) as the critical effects. The selected RfD from the Lau et al. study (2006) is supported by
the RfD for effects on the response of the immune system (DeWitt et al. 2008) to external
challenges as observed following the short-term 15-day exposures to mature mice and effects on
organ and body weights in Fl adult males observed following chronic exposure.
Decreased pup body weights were also observed in studies conducted by Wolf et al.
(2007)/White et al. (2009) and Lu et al. (2015) using mice receiving external doses within the
same order of magnitude (1,3, and 5 mg/kg/day respectively) as those in the chosen study for the
RfD. The selected RfD from the reproductive and developmental study is supported by the
longer-term RfD for effects on the response of the immune system (DeWitt et al. 2008) to
external challenges as observed following the short-term exposures to mature mice and the
effects on kidney weight observed at the time of sacrifice in the FO parental males in the
Butenhoff et al. study (2004a) that provided the modeled serum data.
Support for the selected RfD is provided by other key studies with NOAELs and LOAELs
similar to those used for quantification, yet lacked serum data that could be used for modeling.
There were effects on liver weight and hepatic hypertrophy in the Perkins et al. (2004) and
Perfluorooctanoic acid (PFOA) - May 2016 4-16
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DeWitt et al. (2008) studies that were not considered in the identification of the study LOAEL
because of a lack of data to demonstrate adversity as determined by the Hall criteria (Hall 2012).
The LOAEL for evidence of hepatic necrosis and other signs of tissue damage in the Fl male
rats from the Butenhoff et al. study (2004a) was 3 mg/kg/day, and the NOAEL was 1 mg/kg/day.
In the Loveless et al. study (2008), for male rats 1 mg/kg/day was a NOAEL for increased
relative liver weight and focal liver necrosis was seen at a LOAEL of 10 mg/kg/day, while in
male mice the NOAEL was 0.3 mg/kg/day for the increased liver weight and focal liver necrosis
was at a LOAEL of 1 mg/kg/day following a 29-day exposure. In the study by Tan et al. (2013)
the degree of damage to the liver at 5 mg/kg/day became more severe with increased necrosis,
inflammation, and steatosis when animals were given an HFD. The HED modeled from the
average serum value in mice for the LOAEL of 3 mg/kg/day from Wolf et al. (2007)/White et al.
(2009) was 0.0109 mg/kg/day, about twice that of the 0.0053 mg/kg/day for the rats in the Lau et
al. study (2006) at a LOAEL of 1 mg/kg/day. Both studies lacked a NOAEL.
Using the PK model of Wambaugh et al. (2013), average serum PFOA concentrations were
derived from area under the curve (AUC) considering the number of days of exposure before
sacrifice. The predicted serum concentrations were converted as described above to oral HEDs in
mg/kg/day for each corresponding serum measurement. The POD for the derivation of the RfD
for PFOA is the HED of 0.0053 mg/kg/day that corresponds to a LOAEL that represents
approximately 60% of steady-state concentration. An UF of 300 (10 UFn, 3 UFA, and 10 UFL)
was applied to the HED LOAEL to derive an RfD of 0.00002 mg/kg/day.
There are extensive human data from epidemiology studies on the general population, as well
as worker cohorts. The epidemiology data provide support for the human relevance of the
hazards identified in the laboratory animals. However, they lack the quantitative information on
the human exposures (doses and durations) responsible for the human serum levels. Although
some associations show a relationship between effects and serum measures, the serum measures
are lower than the PODs from the animal studies and some associations are confounded by
reverse causality. Data supporting a first-order kinetic relationship between dose/duration and
serum concentrations are needed before the human data can be used in a manner comparable to
the process utilized in the RfD derivation.
4.1.3 RfC Determination
Limited data from human epidemiology and animal toxicity studies were available with
which to evaluate the potential health effects resulting from continuous inhalation exposure to
PFOA. The available data base, summarized below for human and animal data, does not provide
the minimum data needed for derivation of the RfC as discussed in USEPA (1994b). Thus, the
RfC for PFOA is not recommended or derived.
Human Data. Studies have examined occupational and residential populations at or near 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 database. While inhalation is an important route of exposure to workers,
drinking water was identified as a contributor to exposure in the general population. In all of the
epidemiology studies, wide ranges of serum levels were reported that are likely a reflection, not
only of intra-human toxicokinetic variability, but also of diversity in external exposure sources
and routes of exposure. Thus, the data cannot be clearly applied to quantification of dose-
response via inhalation.
Perfluorooctanoic acid (PFOA) - May 2016 4-17
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Animal Data. Inhalation toxicity data in laboratory animals were limited to acute, single, and
repeated exposures for PK studies, and a developmental toxicity study in rats. No subchronic or
chronic inhalation toxicity studies in animals were available for assessment. Generally, adverse
effects observed following inhalation exposure to PFOA were similar to effects following
exposure to an irritating dust. For male rats exposed to PFOA as a dust in air, the 4-hour LCso
was 980 mg/m3, with adverse clinical signs of body weight loss, irregular breathing, red
discharge around the nose and eyes, and corneal opacity and corrosion (Kennedy et al. 1986,
2004).
Distinct toxicokinetic differences between male and female rats were found following single
and repeated inhalation exposures. Sprague-Dawley rats were exposed nose-only to PFOA
aerosols of 0, 1, 10, or 25 mg/m3 for 6 hours or for 6 hours/day, 5 days/week, for 3 weeks
(Hinderliter 2003). Absorption was indicated in both males and females after single and repeated
exposures with plasma PFOA concentrations proportional to exposure concentration. The Cmax
values were approximately 2-3 times higher in males than in females and persisted for up to
6 hours in males compared to just 1 hour in females. Similarly, the elimination of PFOA was
rapid by females at all exposure levels, and by 12 hours after exposure the plasma levels had
dropped below the analytical LOQ (0.1 |ig/ml). In males, the plasma concentration remained
approximately 90% of the peak concentration at all exposure levels at 24 hours after exposure,
and steady state was reached following repeated exposures. While these results clearly show
toxicokinetic differences between male and female rats, toxicity data were not included, limiting
use of the information in a quantitative risk assessment.
In a developmental toxicity study, pregnant Sprague-Dawley rats were exposed whole-body
to PFOA dust concentrations of 0, 0.1, 1, 10, or 25 mg/m3 for 6 hours/day on GDs 6-15 (Staples
et al. 1984). Dams were either sacrificed on GD 21 or allowed to litter and rear their offspring
until PND 35. Maternal toxicity at 10 and 25 mg/m3 consisted of wet abdomens,
chromodacryorrhea, chromorhinorrhea, a general unkempt appearance, lethargy (high-
concentration group only), and decreased body weight and food consumption. Five out of
24 dams died during treatment at 25 mg/m3. Significantly increased mean liver weight (p<0.05)
was seen at 25 mg/m3. No effects were observed on the maintenance of pregnancy or fetal and
pup survival. At 25 mg/m3, mean offspring body weight was lower than that of controls on GD
21 and throughout lactation.
4.2 Dose-Response for Cancer Effects
As discussed in section 3.4.5, there is equivocal evidence that PFOA exposure might be
associated with an increased risk for cancer from the human epidemiology database and animal
studies. Only one study in highly exposed worker populations showed a positive association
between death from cancer and PFOA exposure. In that study, a significant increase in mortality
due to kidney cancer was found for workers in the highest quartile of cumulative PFOA
exposure; the estimated average mean serum level was 0.35 |ig/mL (Steenland and Woskie
2012). Mortality from cancer in PFOA workers was not elevated in several other studies
(Leonard et al. 2008; Lundin et al. 2009; Raleigh et al. 2014). Serum levels were not available in
studies reporting only mortality. No association was found between PFOA level and cancer
incidence rate in workers with mean serum of 0.113 |ig/mL (Steenland et al. 2015).
No associations were found in the general population between mean serum PFOA levels up
to 0.0866 |ig/mL and colorectal, breast, prostate, bladder, and liver cancer (Bonefeld-J0rgensen
Perfluorooctanoic acid (PFOA) - May 2016 4-18
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et al. 2014; Eriksen et al. 2009; Hardell et al. 2014; Innes et al. 2014). In contrast, two studies
involving members of the C8 Health Project showed a positive association between PFOA levels
(mean at enrolment 0.024 |ig/mL) and kidney and testicular cancers (Barry et al. 2013; Vieira et
al. 2013).
The only chronic bioassays of PFOA were conducted in rats (Biegel et al. 2001; Butenhoff et
al. 2012). The two studies support a positive finding for the ability of PFOA to be weakly
tumorigenic in one or more organs of male but not female rats. There are no carcinogenicity data
from a second animal species. The study by Butenhoff et al. (2012) examined male and female
rats; the Biegel et al. study only evaluated males. The tumor types observed were:
Liver (Butenhoff et al. 2012).
Leydig Cell (Biegel et al. 2001; Butenhoff et al. 2012).
Pancreatic Ascinar Cell (Biegel et al. 2001).
The dose response information and tumors incidence data from the Butenhoff et al. (2012)
and Biegel et al. (2001) studies are summarized in Table 4-10. The data are limited in that only
Butenhoff et al. tested more than one dose. Only one tumor-type (Leydig cell adenoma)
demonstrated a dose-response relationship.
Table 4-10. Summary of Tumor Data from Animal Studies
Tissue
Liver Male
Liver Male
Liver Male
Liver Female
Testes Male
Testes Male
Pancreas Male
Pancreas Male
Concentration in Diet (ppm)
Oa
7/50
0/80
2/80
0/50
0/50
0/80
1/80
0/80
30
2/50
NT
NT
0/50
2/50
NT
NT
NT
300
10/50
0/76
10/76
2/50
7/50
8/76
0/76
7/76
Tumor Type
Hepatocellular carcinoma
Hepatocellular carcinoma
Hepatocellular adenoma
Hepatocellular carcinoma
Leydig Cell adenomas
Leydig Cell adenomas
Acinar Cell carcinoma
Acinar Cell adenoma
Reference
Butenhoff etal. 2012
Biegel etal. 2001
Biegel etal. 2001
Butenhoff etal. 2012
Butenhoff etal. 2012
Biegel etal. 2001
Biegel etal. 2001
Biegel etal. 2001
Notes: a The value reported is for the ad libitum control
NT = Not tested
There are some data that provide support for the hypothesis that the PPARa agonism is the
MO A for the observed liver tumors in rats. PPARa is found in human livers and, when activated,
is linked through activation to a number of metabolic responses but not to the large-scale
peroxisome proliferation associated with tumors in rats and other rodent species. The data
support a PPARa MOA for the liver tumors and, thus, are indicative of lack of relevance to
humans.
PPARa activation might also play a role in the other tumor types observed. However for the
Leydig tumors the PPARa involvement is indirect. The favored hypothesis for the DNA
replication errors responsible for induction of Leydig tumors are postulated to be a consequence
of the following sequence of events:
Decreased testosterone synthesis.
Increased GnRH and increased levels of LH leading to chronic stimulation of Leydig
cells by growth-stimulating mediators including IGF-1, TGF-a, leukotrienes and various
free radicals (Clegg et al. 1997; Li et al. 2011).
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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 MO As are needed.
Current MOA theories for the PACT tumors are linked to the impact of either the mitogenic
effects of elevated testosterone levels or intestinal tissue hormones (CCK, gastrin, or both) in
promoting proliferation of acinar cell preneoplastic foci (Klaunig et al. 2003; Obourn et al.
1997). PACT tumors are most commonly found in rats but also occur in humans. Because PFOA
is associated with decreased rather than increased levels of testosterone, the mechanistic link
between PFOA exposure and PACT is more likely associated with gastric hormone changes,
possibly associated with alterations in bile composition. Some of the membrane transporters that
are impacted by PFOA function in transport of bile components from the liver to the gallbladder
and thereby to the intestines. Cholecystokinen and gastrin stimulate contraction of the
gallbladder and release of bile into the intestines. Data to support this hypothesis are not
available for PFOA. Obourn et al. (1997) studied the impact of PFOA on CCK using in vitro
assays and found that it was not an agonist for the CCKA receptor that activates CCK release.
The increase in hepatocellular tumors did not show a direct relationship to dose in male rats
and was not significantly elevated in either males or females at the high dose when compared to
controls.
There was a dose-related significant increase in LCTs in male rats in the Butenhoff et al.
study (2012), which was confirmed by the high dose in the single-dose mechanistic study by
Biegel et al. (2001). At the high dose (300 ppm in the diet; 14.2 mg/kg/day), tumors were found
in 14% of the male rats at the end of 2 years in the Butenhoff et al. study (2012) and 4% at the
low dose (1.3 mg/kg/day). In the Biegel et al. study (2001), 11% were affected at a dose of 300
ppm in the diet (13.6 mg/kg/day). In each case, there were no LCTs in the controls. The PACT
tumors, only detected in the single dose Biegel et al. study (2001), do not support quantification.
Under the EPA 2005 cancer guidelines, the evidence for the carcinogenicity of PFOA is
considered suggestive because only one species has been evaluated for lifetime exposures and
the tumor responses occurred primarily in males. Dose-response data are only available for the
LCTs in one study. However, two studies involving members of the C8 Health Project showed a
positive association between PFOA levels (mean at enrolment of 0.024 jig/mL) and kidney and
testicular cancers (Barry et al. 2013; Vieira et al. 2013). Therefore, the data on LCTs from
Butenhoff et al. (2012) were modeled to provide a perspective on the magnitude of the potential
cancer risk as it compares with the level of protection provided by the RfD.
The dose-response for the LCTs from Butenhoff et al. (2012) was modeled using EPA's
Benchmark Dose Software (BMDS) Version 2.3.1. The multistage cancer model predicted the
dose at which a 4% increase in tumor incidence would occur. The 4% was chosen as the low-end
of the observed response range within the Butenhoff et al. (2012) results. Both the first and
second degree polynomials gave identical goodness-of-fit criteria (p value and Akaike's
Information Criterion [AIC]). Results are shown in Table 4-11 and Figure 4-1 and details of the
model run are given in Appendix A.
Perfluorooctanoic acid (PFOA) - May 2016 4-20
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Table 4-11. Multistage Cancer Model Dose Prediction Results for a 4% Increase
in LCT Incidence
First Degree Polynomial Fit
Second Degree Polynomial Fit
AIC = 62.6936
BMD (mg/kg/day)
3.51
3.51
P = 0.2245
BMDL (mg/kg/day)
1.99
1.99
Source: Butenhoff etal. (2012)
Multistage Cancer Model with 0.95 Confidence Level
0.25
0.2
0.15
0.1
0.05
0
X^^-
--"
BMDL
^
0 2
IVIUILISLdyC V^dMOCI
^<
^^^
^^"
BMD :
4 6 8 10 12 14
dose
11:5905/092013
Figure 4-1. BMD Model Results for LCTs (Butenhoff et al. 2012)
The CSF for PFOS is derived from the BMDLo4 of 1.99 mg/kg/day after converting the
animal BMDL to a HED using body weights to the 3/4 power. The HED is calculated as follows2:
HED = Animal BMDL x (animal body weight)1/4 H- (human body weight)1/4
HED = 1.99 mg/kg/day x [(0.523 kg)1/4 H- (70 kg)1/4] = 1.99 mg/kg/day x 0.29 =
0.58 mg/kg/day
Where:
1.99 mg/kg/day = BMDL04 for LCTs
0.29 = The dosimetric adjustment factor
The CSF is calculated from the BMDLo4 HED as follows
CSF = response -H BMDLo4 HED
CSF = 0.04 H- 0.58 mg/kg/day = 0.07 (mg/kg/day)-1
2 Body weight for male Sprague-Dawley rats (chronic Exposures) USEPA 1988
Perfluorooctanoic acid (PFOA) - May 2016
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The CSF should not be used at doses > 0.58 mg/kg/day, the HED corresponding to the POD
for the 4% incidence of LCTs following lifetime exposure to PFOA. The observed dose-response
relationships do not continue linearly above this level, and the fitted dose-response models better
characterize the dose-response for the higher exposures. The calculated concentration in drinking
water with one-in-a-million risk for an increase in testicular tumors at levels greater than
background is 0.0005 mg/L.
10-6 Cancer Risk Concentration = 0.000001 x 80 kg + [0.07 (mg/kg/day)-1 x 2.5 L/day] =
0.000457 mg/L (rounded to 0.0005 mg/L).
The equivalent concentration derived from the RfD for noncancer effects using default adult
body weight of 80 kg and a default DWI rate of 2.5 L/day (USEPA 2011) and a 20% relative
source contribution (RSC) to account for the contribution of drinking water to the total exposure
is 0.0001 mg/L. The 0.0001 mg/L concentration derived for an adult based on the RfD for
noncancer effects with a 20% RSC for drinking water is lower than the concentration of
0.0005 mg/L associated with a one-in-a-million risk for testicular cancer also derived with adult
default values indicating that a guideline derived from the developmental endpoint will be
protective for the cancer endpoint.
Perfluorooctanoic acid (PFOA) - May 2016 4-22
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2010. Effects of perfluorooctanoic acid (PFOA) exposure to pregnant mice on
reproduction. The Journal of Toxicological Sciences 35:527-533.
Yan, S., H. Zhang, F. Zheng, N. Sheng, X. Guo, and J. Dai. 2015. Perfluorooctanoic acid
exposure for 28 days affects glucose homeostasis and induces insulin hypersensitivity in
mice. Scientific reports 12(5): 11029.
Yang, Q., Y. Xie, and W. Depierre. 2000. Effects of peroxisome proliferators in the thymus
and spleen of mice. Clinical & Experimental Immunology 122:219-226.
Yang, Q., Y. Xie, A.M. Ericksson, B.D. Nelson, and J.W. DePierre. 2001. Further evidence for
the involvement of inhibition of cell proliferation and development in thymic and splenic
atrophy induced by the peroxisome proliferator perfluoroctanoic acid in mice.
Biochemical Pharmacology 62:1133-1140.
Yang, Q., Y. Xie, S.H.E. Alexson, B.D. Nelson, and J.W. DePierre. 2002a. Involvement of the
peroxisome proliferator-activated receptor alpha in the immunomodilation caused by
peroxisome proliferators in mice. Biochemical Pharmacology 63:1893-1900.
Yang, Q., M. Abedi-Valugerdi, Y. Xie, X. Zhao, G. Molle, B.D. Nelson, and J.W. DePierre.
2002b. Potent suppression of the adaptive immune response in mice upon dietary
exposure to the potent peroxisome proliferator, perfluorooctanoic acid. International
Immunopharmacology 2:389-397.
Yang, C., Y.S. Tan, J.R. Harkema, and S.Z. Haslam. 2009. Differential effects of peripubertal
exposure to perfluorooctanoic acid on mammary gland development in C57B1/6 and
Balb/c mouse strains. Reproductive Toxicology 27:299-306.
Yang, C.-H., K.P. Glover, and X. Han. 2009. Organic anion transporting polypeptide
(Oatp)lal-mediated perfluorooctanoate transport and evidence for a renal reabsorption
mechanism of Oatplal in renal elimination of perfluorocarboxylates in rats. Toxicology
Letters 190:163-171.
Yang, C-H., K.P. Glover, and X. Han. 2010. Characterization of cellular uptake of
perfluorooctanoate via organic-anion transporting polypeptide 1A2, organic anion
transporter 4, and urate transporter 1 for their potential roles in mediating human renal
reabsorbtion of perfluorocarboxylates. Toxicological Sciences 117:294-302.
Yeung, L.W.Y., K.S. Guruge, S. Taniyasu, N. Yamashita, P.W. Angus, and C.B. Herath. 2013.
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urinary excretion of perfluorooctanoic acid in the male rat. Pharmacological &
ToxicologicalMethods 65:274-277.
Ylinen, M., A. Kojo, H. Hanhijarvi, and P. Peura. 1990. Disposition of perfluorooctanoic acid
in single and subchronic administration. Bulletin of Environmental. Contamination and
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York, R.G., G.L. Kennedy, G.W. Olsen, and J.L. Butenhoff 2010. Male reproductive system
parameters in a two-generation reproduction study of ammonium perfluorooctanoate in
rats and human relevance. Toxicology 271:64-72.
Yoshikane, M., Y. Shibata, and N. Shimizu. 2010. BranchedIsomer Profiles ofPerfluoroalkyl
Carboxylates in Japanese Environment. Poster 63. PFAA Days III: Recent advances in
perfluoroalkyl acid (PFAA) research. U.S. Environmental Protection Agency, Office of
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Yu, N., X. Wang, B. Zhang, J. Yang, J. Li, W. Shi, S. Wei, and Y. Yu. 2015. Distribution of
perfluorooctane sulfonate isomers and predicted risk of thyroid hormonal perturbation in
drinking water. Water Research 76:171-180.
Zai'r, Z.M., JJ. Eloranta, B. Stieger, and G.A. Kullak-Ublick. 2008. Pharmacogenetics of
OATP (SLC21/SLCO), OAT and OCT (SLC22) and PRPT (SLC15) transporters in the
intestine, liver, and kidney. Pharmacogenomics 9:597-624.
Zeng, H.-C., Q.-Z. He, Y.-Y. Li, C.-Q. Wu, Y.-M. Wu, and S.-Q. Xu. 2014. Prenatal exposure
to PFOS caused mitochondria-mediated apoptosis in heart of weaned rat. Environmental
Toxicology 30:1082-1090.
Zhang, L., X.-M. Ren, and L.-H. Guo. 2013. Structure-based investigation on the interaction of
perfluorinated compounds with human liver fatty acid binding protein. Environmental
Science & Technology 47:11293-11301.
Zhang, T., H. Sun, Y. Lin, Y. Qin, X. Geng, and L. Kannan. 2013. Distribution of poly- and
perfluoroalkyl substances in matched samples from pregnant women and carbon chain
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Zhang, Y., S. Beesoon, L. Zhu, and J.W. Martin. 2013. Biomonitoring of perfluoroalkyl acids
in human urine and estimates of biological half-life. Environmental Science &
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Zhang, L., X.-M. Ren, B. Wan, and L.-H. Guo. 2014. Structure-dependent binding and
activation of perfluorinated compounds on human peroxisome proliferator-actovated
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Zhang, T., H. Sun, X. Qin, Z. Gan, and K. Kannan. 2014. PFOS and PFOA in paired urine and
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Zhang, W., F. Wang, P. Xu, C. Miao, X. Zeng, X. Cui, C. Lu, H. Xie, H. Yin, F. Chen, and J.
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Zhang, C., R. Sundaram, J. Maisog, A.M. Calafat, D. Boyd Barr, and G.M. Buck Louis. 2015.
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Zhao, G., J. Wang, S. Chen, Y. Zhao, F. Gu, A. Xu, and L. Wu. 2010. Mutagenicity of PFOA
in mammalian cells: role of mitochondrial-dependent reactive oxygen species.
Environmental Science & Technology 45(4): 1638-1644.
Zhao, Y., Y.S. Tan, S.Z. Haslam, and C. Yang. 2010. Perfluorooctanoic acid effects on steroid
hormones and growth factor levels mediate stimulation of peripubertal mammary gland
development in C57BL/6 mice. Toxicological Sciences 33(4):563-576.
Zhao, B., L. Li, J. Liu, H. Li, C. Zhang, P. Han, Y. Zhang, X. Yuan, R.S. Ge, and Y. Chu.
2014. Exposure to perfluorooctane sulfonate in utero reduces testosterone production in
rat fetal Leydig cells. PLOS ONE 9:e78888.
Perfluorooctanoic acid (PFOA) - May 2016 5-39
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Appendix A: Literature Search Strategy Developing the Search
The literature search strategy was planned with input from EPA library services staff. Chemical
Abstracts Service (CAS) numbers served as the basis for identification of relevant search terms.
Trial searches were conducted, and results were evaluated to refine the search strategy (e.g., to
pevent retrieval of citations unrelated to health and occurrence). The search string was refined to
improve the relevancy of the results. All searches were conducted in the PubMed database,
which contains peer-reviewed journal abstracts and articles on various biological, medical,
public health, and chemical topics. The first search string (as well as future iterations) is
presented below.
Every 2 weeks, a search was run in PubMed and a bibliography of the search results was
compiled.
In 2012, the State of New Jersey Department of Environmental Protection (NJDEP) initiated a
monthly search in PubMed for emerging literature on perfluorinated chemicals, primarily from
the carboxylic acid and sulfonate families. These searches were provided to the EPA on a
monthly basis. There was a high degree of overlap with the results from the EPA search, thus
increasing the confidence in the search strategy.
In 2013, the EPA search strategy was expanded to cover other members of the
perfluorocarboxylic acids (C-4 to C-12) and sulfonate families (C-4, C-6, C-8). The search string
was altered in June of 2013 to rely more on the search features offered by PubMed.
A change in the PubMed database structure in 2015 required additional modifications to the
search strategy.
The NJDEP search terms did not change from 2012 to 2015. All search iterations are noted
below. However, the reports shared with EPA were streamlined to remove information on
analytical methods and other nonhealth related citations that were not consistent with the criteria
presented previously in the backgorund section of this document.
Search Strategy Examples: (Arranged from most recent to oldest).
2015
Search: perfluorooctanoate OR "perfluorooctanoic acid" OR "perfluoroctanoic acid" OR pfoa
OR "perfluorinated chemicals" OR "perfluorinated compounds" OR "perfluorinated homologue
groups" OR "perfluorinated contaminants" OR "perfluorinated surfactants" OR perfluoroalkyl
acids OR "perfluorinated alkylated substances" OR "perfluoroalkylated substances" OR pfba OR
"perfluorobutanoic acid" OR perfluorochemicals OR "telomer alcohol" OR "telomer alcohols"
OR "fluorotelomer alcohols" OR "polyfluoroalkyl compounds" OR "perfluorooctane sulfonate"
OR pfos OR "perfluorooctanesulfonic acid" OR "perfluorooctane sulfonic acid" OR
"perfluorooctane sulphonate" OR perfluorooctane sulfonate OR "perfluorooctanyl sulfonate" OR
"Heptadecafluorooctane-1-sulphonic" OR"Heptadecafluoro-l-octanesulfonic acid" OR
perfluorononanoate OR pfhxa OR "perfluorohexanoic acid" OR "fluorinated surfactants"
Perfluorooctanoic acid (PFOA) - May 2016 A-1
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Filters: English.
Frequency: Every 2 weeks
September 2013
Search: perfluorooctanoate OR "perfluorooctanoic acid" OR "perfluoroctanoic acid" OR pfoa
OR "perfluorinated chemicals" OR "perfluorinated compounds" OR "perfluorinated homologue
groups" OR "perfluorinated contaminants" OR "perfluorinated surfactants" OR perfluoroalkyl
acids OR "perfluorinated alkylated substances" OR "perfluoroalkylated substances" OR pfba OR
"perfluorobutanoic acid" OR perfluorochemicals OR "telomer alcohol" OR "telomer alcohols"
OR "fluorotelomer alcohols" OR "polyfluoroalkyl compounds" OR "perfluorooctane sulfonate"
OR pfos OR "perfluorooctanesulfonic acid" OR "perfluorooctane sulfonic acid" OR
"perfluorooctane sulphonate" OR perfluorooctane sulfonate OR "perfluorooctanyl sulfonate" OR
"Heptadecafluorooctane-1-sulphonic acid" OR"Heptadecafluoro-l-octanesulfonic acid" OR
perfluorononanoate OR pfhxa OR "perfluorohexanoic acid" OR "fluorinated surfactants"
Filters: English.
Frequency: Every 2 weeks
June 2013
Search: (PFOA[tw] OR perfluorooctanoic acid[tw] OR335-67-l[tw] ORPFBA[tw] OR
perfluorobutanoate[tw] OR 3794-64-7[tw] ORPFDA[tw] OR perflurordecanoic acid[tw] OR
335-76-2[tw] OR PFHpA[tw] OR perfluoroheptanoic acid[tw] OR 375-85-9[tw] OR PFHxA[tw]
OR perfluorohexanoic acid[tw] OR 307-24-4[tw] OR PFNA[tw] OR perfluorononanoic acid[tw]
OR 375-95-1 [tw] OR PFPtA[tw] OR perfluoropentanoic acid[tw] OR 2706-90-3[tw] OR
PFPA[tw] OR pentafluoropropionic acid[tw] OR 422-64-0[tw]) AND (human* [tw] OR
mammal*[tw]) NOT (environment* OR ecolog*)
Filters: English.
Frequency: Every 2 weeks
February, 2013
Search: perfluorooctanoate OR "perfluorooctanoic acid" OR "perfluoroctanoic acid" OR pfoa
OR "perfluorinated chemicals" OR "perfluorinated compounds" OR "perfluorinated homologue
groups" OR "perfluorinated contaminants" OR "perfluorinated surfactants" OR perfluoroalkyl
acids OR "perfluorinated alkylated substances" OR "perfluoroalkylated substances" OR pfba OR
"perfluorobutanoic acid" OR perfluorochemicals OR "telomer alcohol" OR "telomer alcohols"
OR "fluorotelomer alcohols" OR "polyfluoroalkyl compounds" OR "perfluorooctane sulfonate"
OR pfos OR "perfluorooctanesulfonic acid" OR "perfluorooctane sulfonic acid" OR
"perfluorooctane sulphonate" OR perfluorooctane sulfonate OR "perfluorooctanyl sulfonate" OR
"Heptadecafluorooctane-1-sulphonic acid" OR"Heptadecafluoro-l-octanesulfonic acid" OR
perfluorononanoate OR pfhxa OR "perfluorohexanoic acid" OR "fluorinated surfactants"
Perfluorooctanoic acid (PFOA) - May 2016 A-2
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Filters: English.
Frequency: Every 2 weeks
June 2011
Search (perfluorooctanoate OR "perfluorooctanoic acid" OR "perfluoroctanoic acid" OR pfoa
OR "perfluorinated chemicals" OR "perfluorinated compounds" OR "perfluorinated homologue
groups" OR "perfluorinated contaminants" OR "perfluorinated surfactants" OR
perfluoroalkylacids OR "perfluorinated alkylated substances" OR "perfluoroalkylated
substances" OR pfba OR "perfluorobutanoic acid" OR perfluorochemicals OR "telomer alcohol"
OR "telomer alcohols" OR "fluorotelomer alcohols" OR "polyfluoroalkyl compounds" OR
"perfluorooctane sulfonate" OR pfos OR "perfluorooctanesulfonic acid" OR "perfluorooctane
sulfonic acid" OR "perfluorooctane sulphonate" OR perfluorooctanesulfonate OR
"perfluorooctanyl sulfonate" OR"Heptadecafluorooctane-l-sulphonic acid" OR
"Heptadecafluoro-1-octanesulfonic acid" OR perfluorononanoate OR pfhxa OR
"perfluorohexanoic acid" OR "fluorinated surfactants" OR 335-67-1 [rn])
Limits: Publication DateDates will change for each search, English Language only
June 2009
Search (perfluorooctanoate OR "perfluorooctanoic acid" OR "perfluoroctanoic acid" OR pfoa
OR "perfluorinated chemicals" OR "perfluorinated compounds" OR "perfluorinated homologue
groups" OR "perfluorinated contaminants" OR "perfluorinated surfactants" OR
perfluoroalkylacids OR "perfluorinated alkylated substances" OR "perfluoroalkylated
substances" OR pfba OR "perfluorobutanoic acid" OR perfluorochemicals OR "telomer alcohol"
OR "telomer alcohols" OR "fluorotelomer alcohols" OR "polyfluoroalkyl compounds" OR
"perfluorooctane sulfonate" OR pfos OR "perfluorooctanesulfonic acid" OR "perfluorooctane
sulfonic acid" OR "perfluorooctane sulphonate" OR perfluorooctanesulfonate OR
perfluorononanoate OR pfhxa OR "perfluorohexanoic acid" OR "fluorinated surfactants" OR
335-67-1 [rn] OR 1763-23-1 [rn])
Limits: Entrez Date from 2009/04/07 to 2009/04/12
New Jersey Search Terms
Search: perfluorinated OR perfluorooctanoate OR perfluorononanoate OR
perfluorooctanesulfonate OR perfluorooctanesulphonate OR perfluoroalkylated OR
perfluoroalkyl OR polyfluoroalkyl OR polyfluorinated OR PFBA OR PFBS OR PFDA OR
PFHA OR PFFffA OR PFHXA OR PFHXS OR PFNA OR PFOA OR PFOAs OR PFOS OR
PFUNDA OR "perfluorooctanoic acid" OR "perfluoro octanoic acid" OR "perfluorooctane
sulfonate" OR "perfluorooctane sulfonic acid" OR "perfluorooctanesulfonic acid" OR
"perfluorooctane sulphonate" OR "perfluorooctanyl sulfonate" OR "perfluorobutanoic acid" OR
"perfluoroalkyl acids" OR "perfluorononanoic acid" OR "perfluorohexanoic acid" OR
"perfluorohexane sulfonate" OR "perfluorohexane sulphonate" OR perfluorobutanoate OR
"perfluoro butanoate" OR perfluorohexanoate OR "perfluoro hexanoate"
Filters: 1
Perfluorooctanoic acid (PFOA) - May 2016 A-3
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Appendix B: Studies Evaluated Since August 2014
The tables that follow identify the papers that were retrieved and reviewed for inclusion
following the August 2014 peer review for the draft PFOS Health Effects Support Document.
The papers listed include those recommended by the peer reviewers or public commenters, as
well as those identified from the literature searches between the completion of the peer review
draft and December 2015. The review of papers recommended by the commenters and their
potential impact on the updates to the draft assessments was facilitated by publications such as
the critical review of the recent literature by Post et al. (2012). Post et al. (2012) provides an in-
depth analysis of the available health effects literature for PFOA. Papers included in the final
FtESD are noted and reasons provided for those that were not included in the final document.
The tables for document retrieval and review are followed by updated versions of the
summaries of the epidemiology summary tables from the peer reviewed draft as recommended
by the peer reviewers. They are a useful tool to facilitate a high level comparison of the study
outcomes for each of the epidemiological study groupings.
The criteria utilized in determining the papers that were included in the FtESD after the peer
review and presented in the Background were the following:
1. The study examines a toxicity endpoint or population that had not been examined by
studies already present in the draft assessment.
2. Aspects of the study design, such as the size of the population exposed or quantification
approach, make it superior to key studies already included in the draft document.
3. The data contribute substantially to the weight of evidence for any of the toxicity
endpoints covered by the draft document.
4. There are elements of the study design that merit its inclusion in the draft assessment
based on its contribution to the mode of action or the quantification approach.
5. The study elucidates the mode of action for any toxicity endpoint or toxicokinetic
property associated with PFOA exposure.
6. The effects observed differ from those in other studies with comparable protocols.
Table B-l. PFOS Epi PapersPost Peer Review (Retrieved and Reviewed)
Authors and Year
Andersen etal. 2013
Back etal. 2015
Barrett etal. 2015
Berg etal. 2015
Bonefeld-Jergenson et al. 2014
Bonefeld-Jergenson et al. 201 1
Brieger etal. 2011
Buck Louis etal. 2015
Chang etal. 2014
Chen etal. 2015
Dankers etal. 2013
Topic Keywords
Postnatal growth
Time to pregnancy
Ovarian hormone
Thyroid
Breast cancer
Breast cancer
Immune effects
Semen quality
Analysis of human cancer
studies
Birth weight
Blood-testis barrier
Status/Notes
Added PFOA/PFOS
Added PFOA
Not Added No association observed for
PFOA; PFOS was not included in the
assessment
Added PFOA/PFOS
Added PFOA/PFOS
Added PFOA/PFOS
Already presented in PFOS/PFOA
Added PFOA/PFOS
Added PFOA in the cancer weight of
evidence section
Added PFOS
Reviewed, not added; Study of an assay
that used PFOA as one chemical in the test
battery
Perfluorooctane sulfonate (PFOS) - May 2016
B-l
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Authors and Year
Darrowetal. 2013
Darrowetal. 2014
Donaueretal. 2015
Eriksenetal. 2013
Fitz-Simonetal. 2013
Fisher etal. 2013
Fletcher etal. 2013
Fu etal. 2014
Geigeretal. 2014a
Geigeretal. 2014b
Ghisari etal. 2014
Governini et al. 2015
Grandjean and Clapp 2015
Granumetal. 2013
Hardell etal. 2014
Hey er etal. 201 5a
Hey er etal. 201 5b
Humblet etal. 2014
Jain 20 14
Innesetal. 2014
Joensenetal. 2013
Kerger etal. 2011
Kjeldsen and Bonefeld-Jergensen
2013
Kristensen et al. 2013
Liew etal. 2014
Looker etal. 2014
Lopez-Doval et al. 2014
Maisonet etal. 2015
Maisonet etal. 2012
Merck etal. 2015
Okada etal. 2014
Osuna etal. 2014
RothandWilks2014
Shrestha etal. 2015
Topic Keywords
Reproductive outcome
Miscarriage
Infant Neurobehavior
Total cholesterol Danish
Serum lipids
Plasma lipids
Cholesterol-genes
Serum lipids in Chinese
subjects
Lipids/children
Hypertension/children
Breast cancer Inuit
DNA effects in sperm
Health Risks
Immune children
Prostate cancer
Human weight
Behavior motor development
Asthma
NHANES
Colorectal cancer
Sperm
Cholesterol C8
Sex hormones
Prenatal female repro
Cerebral palsy children
Immune
Male repro
Gestational diabetes
Birth weight
PFAS levels in children
Allergy children
Antibodies PFOS PFOA
Neurodevelopmental
Thyroid
Status/Notes
Added PFOA/PFOS
Added PFOA/PFOS
Not added negative for PFOS; No
statistical differences in PFOA levels
during pregnancy and any neuro endpoint.
Better studies.
Added PFOA/PFOS
Added PFOA/PFOS
Added PFOA/PFOS
Added PFOA/PFOS
Not added: Chinese population, dataset
available on U.S. population. More
branched chain isomers found among the
people in China.
Added PFOA/PFOS
Added PFOA/PFOS
Not added; same population as Bonefeld-
Jergensen et al. 2014; this study focuses on
gene polymorphisms
Added PFOA/PFOS
Not added; the primary studies are already
included in the documents.
Added PFOA/PFOS
Added PFOA/PFOS
Added PFOA/PFOS
Added PFOA/PFOS
Added PFOA/PFOS
Added PFOA/PFOS
Added PFOA/PFOS
Added PFOA/PFOS
Added; demographics for cholesterol and
PFOS in summary section of epi studies
Covered multiple PFAS in vitro no impact
on weight of evidence
Added PFOA/PFOS
Added PFOA/PFOS
Added PFOA/PFOS
Added PFOS
Added PFOA/PFOS
Added PFOA/PFOS
Not added; No significant impact
Added PFOS
Not added; focus more on methylHg and
PCB than PFAS; only n = 38 as
preliminary study
. Not added; no significant impact
Added PFOA/PFOS
Perfluorooctane sulfonate (PFOS) - May 2016
B-2
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Authors and Year
Starling etal. 2014
Steenland et al. 2015
Stein et al. 2009
Taylor etal. 2014
VandenHeuvel2013
Vassiliadou etal. 2010
Velez etal. 2015
Verner etal. 2015
Verner and Longnecker 2015
Vested etal. 2013
Vesterinen et al. 2014
Wang etal. 2013
Watkins etal. 2013
Webster etal. 2014
Webster etal. 2015
Wen etal. 2013
Yeung etal. 2013
Zhang etal. 2015
Topic Keywords
Plasma lipids
Workers
Pregnancy
Menopause
Serum lipids
PFOS in cancer vs non-cancer
patients
Fertility
Fetal growth GFR
Menstruation/excretion
Semen quality and hormones
Fetal Growth GFR
Thyroid
Kidney function
Maternal thyroid
Thyroid iodine statue
Thyroid
Liver cancer
Gestational diabetes
Status/Notes
Added PFOA/PFOS
Added PFOA
Added PFOA
Added PFOA/PFOS
Not added; is a rebuttal of Fletcher et al.
2013 conclusions. No significant impact
Added PFOA/PFOS
Added PFOA/PFOS
Added PFOA/PFOS
Added PFOS
Added PFOS/PFOS
Added PFOA/PFOS
Added PFOA/PFOS
Added PFOA/PFOS
Added PFOA/PFOS
Added PFOS
Added PFOA/PFOS
Added PFOA/PFOS
Added PFOA/PFOS
Table B-2. PFOA Post Peer Review Animal Toxicity Studies
Authors and Year
Bjork etal. 2011
Corsinietal. 2014
Corsinietal. 2012
Dewitt etal. 2015
Fenton2015
Filgo etal. 2015
Hall etal. 2012
Koustas etal. 2014
Liu etal. 2015
Long etal. 2013
Lu etal. 2015
Ngo etal. 2014
Post etal. 2012
Quist etal. 2015
Rigden etal. 2015
Shabalina etal. 2015
Topic
Nuclear receptor activation
Immune data review
Immune in vitro data review
Immunotoxicity
Repro editorial
Liver tumors in females
developmentally exposed
PPARa and cancer
Fetal growth (animal studies)
navigation guide
Testes
Neurotoxicity adult PFOS
Testes
Tumors mice Min/+ PFOS
Review paper
Liver histopathology/high fat
diet post weaning exposure
Acute liver effects
Brown fat uncoupling protein 1
Action Notes
In vitro, mechanistic findings comparable to
studies already included
Not added; no significant impact
Not added; no significant impact
Added PFOA
Not added
Added PFOA
Cited in synthesis. Paper on adversity of liver
hypertrophy PFOA/ PFOS
Added PFOA
Added PFOA
Added PFOS
Added PFOA
Added PFOS
Not added. Key studies included in the document;
no significant impact
Added PFOA
Added PFOA
Not added. Mechanistic; no significant impact
Perfluorooctane sulfonate (PFOS) - May 2016
B-3
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Authors and Year
Shengetal. 2016
Tan etal. 2012
Tan etal. 2013
Tucker etal. 2015
Wallace etal. 2013
Wan etal. 20 14b
Wan etal. 2012
Wan etal. 20 14a
F.Wang etal. 2015
S.Wang etal. 2014
L.Wang etal. 2014
Y.Wang etal. 2015
Yan etal. 2015
Yu etal. 2015
Zeng etal. 2014
L.Zhang etal. 2013
Y.Zhang etal. 2013
W.Zhang etal. 2014
Zhao etal. 2014
Topic
Binding to liver fatty acid
binding protein
Gene activation
Gene activation dietary fat
Mammary gland
Mitochodrial respiration
Glucose metabolism
Hepatic steatosis
Sertoli cells
MiRNA liver PFOS early life
Lysine decarboxylase
Inhibition of LDL
Special learning and memory
Glucose homeostasis
Thyroid PFOS isomers
Mitochondria! mediated
apoptosis of the heart
Fatty acid binding protein
Biological half-life
Breast cancer cell invasion
mechanistic
Testosterone reduction in
Leydig cells PFOS
Action Notes
Not added; no significant impact,
other papers
topic covered by
Added PFOA/PFOS
Added PFOA
Added PFOA
Not added. No significant impact,
by other papers
topic covered
Added PFOS
Added PFOS
Added PFOS
Not added; no significant impact
Added PFOA/PFOS
Added PFOS
Added PFOS
Not added. Dose-response in Wan (2014b)
presented (more robust). Single dose for whole
animal
Added PFOS
Added PFOS
Added PFOA/PFOS
Added PFOA/PFOS
Not added; in vitro, no significant
impact
Added PFOS
Perfluorooctane sulfonate (PFOS) - May 2016
B-4
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Table B-3. Toxicokinetics: Post Peer Review
Authors and Year
D'Alessandro et al. 2013
Augustine et al. 2005
Beesoonetal. 2011
Beesoon and Martin 2015
Cuietal. 2010
Fabrega etal. 2014
Kerstner-Wood et al. 2003
Klaassen and Aleksunes
2010
Loccisano et al. 2013
Mondal et al. 2014
Ospinal-Jimenez and Pozzo
2012
Perez etal. 2013
Ren etal. 2015
Rigden etal. 2015
Shabalina etal. 2015
Slitt et al. 2007
Tucker etal. 2015
Verner and Longnecker
2015
Wambaugh etal. 2013
Wong etal. 2014
T.Zhang etal. 2014
L.Zhang etal. 2014
Y.Zhang etal. 2013
T.Zhang etal. 2013
Topic
Serum albumin
Transporter expression testes
Isomer profile
Albumin binding
Excretion subchronic
PK model
Plasma protein binding
Transporter paper Provided
diagram of kidney transporters
PK model Human
Breast milk
Protein denaturation
Human tissue levels
Thyroid hormone receptor
binding (in vitro)
Liver and excretion
Brown fat
Transporter expression PFOA
Menstruation-excretory route
Excretion PFOS
PK model
Menstrual blood as excretory
route
Excretion general population
and pregnancy
PPAR gamma
Excretion, half-life
Maternal transfer
Action Notes
Added PFOS
Not added background paper on testes transporters
-no relationship to PFOA PFOA or any PFAS
Added PFOA
Added PFOA
Added PFOA/PFOS
Added PFOA/PFOS
Added PFOA/PFOS
Added PFOA
Added PFOA/PFOS
Added PFOS/PFOA
Added PFOS
New PFOA/PFOS
Added PFOA/PFOS
Added PFOA
Not added; No information on MOA for body
weight effects in the animal or human studies
Not added. Reported on transporters during
extrahepatic cholestasis. No data on PFOA and
PFOS. No significant impact.
Added PFOA
Added PFOS
Added PFOA/PFOS
Added PFOA/PFOS
Added PFOA/PFOS
Added PFOS
Added PFOA/PFOS
Added PFOA/PFOS
Perfluorooctane sulfonate (PFOS) - May 2016
B-5
-------
Tables B-4 through B-8 provide updated versions of the epidemiology summary tables from the peer-reviewed draft, as
recommended by the reviewers. They are a useful tool to facilitate a high-level comparison of the study outcomes for each of the
epidemiology study groupings.
Table B-4. Association between Serum PFOA and Serum Lipids and Uric Acid
Reference
Study Type
n
Mean Serum
PFOA
TC
VLDL
LDL
HDL
Non-HDL
TG
UA
Occupational Populations
Olsen et al.
2000
Olsen et al.
200 Ib, 2003
Olsen et al.
200 Ic, 2003
Sakr et al.
2007a
Sakr et al.
2007b
Olsen and
Zobel 2007
Costa et al.
2009
Cross-sectional
Cross-sectional
Longitudinal; ~5
years
Cross-sectional
Longitudinal
Cross-sectional
Cross-sectional
111(1993)
80 (1995)
74 (1997)
206 (Antwerp)
215 (Decatur)
175
(Decatur and
Antwerp
combined for
analysis)
1,025
454 (23 -yr
follow-up)
506 (Antwerp,
Cottage Grove,
Decatur
combined)
34 workers
107 controls
0-80 ug/mL
0-114 ug/mL
0.1-81 ug/mL
1.03 ug/mL
1.90 ug/mL
1.36-1.41
ug/mL (1995
baseline)
1.49-1.77
ug/mL (2000
follow-up)
0.428 ug/mL
1.04 ug/mL
(first)
1.16 ug/mL
(last)
2.21 ug/mL
4.02 ug/mL
< >
< >
< >
t
t
t
t
< >
t
NM
NM
NM
t
NM
NM
NM
< >
< >
< >
NM
NM
t
< >
< >
NM
< >
< >
< >
< >
<>
<>
< >
I
< >
NM
NM
NM
NM
NM
NM
NM
< >
< >
< >
t
t
4-^>
t
< >
NM
NM
NM
t
NM
NM
t
General Populations
Emmett et al.
2006
Steenland et al.
2009
Steenland et al.
2010
Cross-sectional
Cross-sectional
(C8)
Cross-sectional
(C8)
371
46,294
53,458
0.354 ug/mL
0.08 ug/mL
0.086 ug/mL
4-^>
t
NM
NM
NM
NM
NM
t
NM
NM
4-^>
NM
NM
t
NM
NM
t
NM
NM
NM
t
Perfluorooctanoic acid (PFOA) - May 2016
B-6
-------
Reference
Winquist and
Steenland
2014a
Frisbee et al.
2010
Fitz-Simon et
al. 2013
Nelson et al.
2010
Eriksen et al.
2013
Starling et al.
2014
Fisher et al.
2013
Study Type
Cross-sectional
(C8)
Cross-sectional
(C8, children and
adolescents)
Longitudinal; 4.4
years (C8)
Cross-sectional
(NHANES)
Cross-sectional
Cross-sectional
(maternal at 14-
26 weeks
gestation)
Cross-sectional
n
32,254
6,536 children
5,934
adolescents
521
1,445
753
891
2,700
Mean Serum
PFOA
0.0261 ug/mL
0.0777 ug/mL
0.0618 ug/mL
0.140 ug/mL
(baseline)
0.068 ug/mL
(follow-up)
0.0046 ug/mL
0.0071 ug/mL
0.00225
ug/mL
0.0025 ug/mL
TC
t
t
4-^>
t
t
< >
< >
VLDL
NM
NM
NM
NM
NM
NM
NM
LDL
NM
t
4-^>
4-^>
NM
< >
< >
HDL
NM
< >
<>
<>
NM
t
< >
Non-HDL
NM
NM
NM
t
NM
NM
NM
TG
NM
t
4-^>
NM
NM
< >
NM
UA
NM
NM
NM
NM
NM
NM
NM
Notes: | = positive association; j = 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 = tiiglycerides; UA = uric acid; NM = not measured
Perfluorooctanoic acid (PFOA) - May 2016
B-7
-------
Table B-5. Association of Serum PFOA and Biochemical and Hematological Measures
Reference
Study Type
n
Mean Serum
PFOA
Liver
enzymes
Bilirubin
Renal
Enzymes/Function
Glucose
Hematology
Occupational Populations
Olsen et al.
2000
Olsen et al.
200 Ib, 2003
Olsen et al.
200 Ic, 2003
Sakr et al.
2007a
Sakr et al.
2007b
Olsen and
Zobel 2007
Costa et al.
2009
Cross-sectional
Cross-sectional
Longitudinal;
~5 years
Cross-sectional
Longitudinal
Cross-sectional
Cross-sectional
111(1993)
80 (1995)
74 (1997)
206
(Antwerp)
215 (Decatur)
175
(Decatur and
Antwerp
combined for
analysis)
1025
454
506
(Antwerp,
Cottage
Grove,
Decatur
combined)
56 workers
0-80 ug/mL
0-1 14 ug/mL
0.1-81 ug/mL
1.03 ug/mL
1.90 ug/mL
1.36-1.41 ug/mL
(1995 baseline)
1.49-1.77 ug/mL
(2000 follow-up)
0.428 ug/mL
1.04 ug/mL (first)
1.16 ug/mL (last)
2.21 ug/mL
4.02 ug/mL
< >
<>
<>
<>
<>
<>
<>
t (GOT only)
t (AST only)
t(ALP,
ALT, GOT
Decatur only)
t (GOT,
ALP, ALT)
*-^>
*-^>
*-^>
*-^>
*-^>
*-^>
*-^>
I
I
I
*-^>
*-^>
*-^>
*-^>
*-^>
*-^>
*-^>
NM
NM
NM
4-^>
< >
<>
<>
<>
<>
<>
<>
< >
NM
NM
^^
< >
<>
<>
<>
<>
<>
<>
< >
NM
NM
4-^>
General Populations
Emmett et al.
2006
Lin et al.
2010
Gallo et al.
2012
Shankar et
al. 2011
Watkins et
al. 2013
Cross-sectional
Cross-sectional
(NHANES)
Cross-sectional
(C8)
Cross-sectional
(NHANES)
Cross-sectional
(C8)
371
1076 men
1 140 women
47,092
4587
9,660
(children)
0.354 ug/mL
0.00505 ug/mL
0.00406 ug/mL
0.028 ug/mL
0.0059 ug/mL
0.0283 ug/mL
^^
t(ALT,
GOT)
t(ALT)
NM
NM
NM
NM
NM
4-^>
NM
NM
t (chronic kidney
disease)
t (decreased eGFR)
NM
NM
NM
NM
NM
4-^>
NM
NM
NM
NM
Notes: | = positive association; j = negative association;
transpeptidase; AST = aspartate aminotransferase; ALT :
>^f = no association; ALP = alkaline phosphatase; eGFR = estimated glumerular filtration rate; GOT = gamma-glutamyl
alanine tiansaminase; NM = not measured
Perfluorooctanoic acid (PFOA) - May 2016
B-8
-------
Table B-6. Association between PFOA level and prevalence of thyroid disease and thyroid hormone levels
Study
Study Type
Population (n)
Mean Serum
PFOA (jig/niL)
Thyroid Disease
TSH
T3
T4
Occupational Populations
Olsenetal. 1998
Olsenetal. 200 Ib, 2003
Sakr et al. 2007a
Costa et al. 2009
Olsen and Zobel 2007
Cross-sectional
Cross-sectional
Cross-sectional
Cross-sectional
Cross-sectional
111 and 80 Adult
workers
Adult workers
215 (Decatur)
206 (Antwerp)
1,025 Adult workers
56 Adult workers
506 Adult workers
10-30
>30
1.9
1.03
0.428
4.02
2.21
NM
NM
NM
NM
NM
NM
t
4-^>
4-^>
4-^>
4-^>
4-^>
NM
4-^>
4-^>
4-^>
t
NM
< >
<>
<>
*-> serum
| free
General Populations
Emmett et al. 2006
Pirali et al. 2009
Bloom etal. 2010
Shrestha et al. 2015
Winquist and Steenland
2014b
Lopez-Espinosa et al. 2012
Melzer etal. 2010
Wen etal. 2013
de Cock etal. 2014
Lin etal. 2013
Chan etal. 2011
Wang etal. 2013
Cross-sectional
Cross-sectional
Cross-sectional
Cross-sectional
Cross-sectional
Cross-sectional
Cross-sectional
Cross-sectional
Cross-sectional
Cross-sectional
Cross-sectional
Cross-sectional
40 (thyroid disease)
331 (no thyroid
disease)
28 Adults
31 Adults
51 men
36 women
32,254 (C8)
10,725 children (C8)
3,966 Adults
(NHANES)
1,181 (NHANES)
83 newborns
545
271 Pregnant
women
903 women at
gestation week 18
0.387
0.451
2.0 ng/g thyroid
tissue
0.00133
0.0104
0.0261
0.0293
0.025 (men)
0.019 (women)
0.00415
0.000943 (cord)
0.00267
0.00135
0.0022
< >
< >
NM
< >
<-> (men)
t (women)
t
<-> (men)
t (women)
NM
NM
NM
NM
NM
NM
NM
4-^>
4-^>
NM
4-^>
NM
< >
NM
< >
< >
< >
NM
NM
NM
t
NM
NM
NM
< >
NM
NM
NM
NM
NM
NM
4-^>
t
NM
4-^>
NM
< >
<->boys
t girls
< >
< >
NM
Perfluorooctanoic acid (PFOA) - May 2016
B-9
-------
Study
Berg etal. 2015
Webster etal. 2014
Study Type
Cross-sectional
Cross-sectional
Population (n)
375 women at
gestation week 18,
day 3 and week 6
after delivery
(Norwegian
Mother/Child
Cohort)
152 women at
gestation week 15-
18
Mean Serum
PFOA (jig/niL)
0.00153
0.0017
Thyroid Disease
NM
NM
TSH
"
T3
"
T4
"
Notes: | = positive association; J, = negative association; <-» = no association; NM = Not Measured
Perfluorooctanoic acid (PFOA) - May 2016
B-10
-------
Table B-7. Association between Serum PFOA and Markers of Immunotoxicity
Study
Steenland et al.
2015
Okadaetal. 2012
Feietal. 2010b
Grandjean et al.
2012
Grandjean et al.
2012
Granum et al.
2013
Dong etal. 2013
Humblet et al.
2014
Looker etal. 2014
Study Type
Cross-sectional
Prospective
cohort
Cross-sectional
Prospective
cohort
Prospective
cohort
Prospective
cohort
Cross-sectional
Cross-sectional
Cross-sectional
Population (n)
Workers (6,027)
Maternal, third
trimester (343)
Maternal, first
trimester
(1,400)
Maternal at
gestation week
32 (587)
Children age 5
years (587)
Women at
delivery (56)
Children age
10-15 years
(23 1 asthmatics
and 225
controls)
Children age
12-19 years
(1,877)
Adults (4 11)
Mean or Median
Serum PFOA
(jig/mL)
0.113
0.0014
0.0056
0.0032
0.00406
0.0011
0.0015 (asthmatics)
0.0010
(nonasthmatics)
0.0043 (asthmatics)
0.0040
(nonasthmatics)
0.0337
Disease
Prevalence
t ulcerative
colitis
t rheumatoid
arthritis
I asthma
<->upto 18
months old
<-> early
childhood
NM
NM
< >
1 for asthma
t for asthma
NM
Vaccine
Response
NM
NM
NM
1 (antibody liter)
1 (antibody liter)
< >
NM
NM
I (antibody tiler)
Notes: | = positive association; J, = negative association; <-» = no association; NM = Not Measured
Perfluorooctanoic acid (PFOA) - May 2016
B-ll
-------
Table B-8. Association between Serum PFOA and Reproductive and Developmental Outcomes
Study
Study Type
n
Mean Serum PFOA
Outcome
Measures at
Birth
Growth/
Development
Fecundity/
Fertility
Reproductive Outcome, Anthropometric Measures at Birth
Fei et al. 2007,
2008a, 2009,
2010a
Velezetal. 2015
Nolan et al. 2009,
2010
Stein et al. 2009
Darrow et al.
2013,2014
Apelberg et al.
2007
Monroy et al.
2008
Washino et al.
2009
Hammetal. 2010
Whitworth et al.
2012
Maisonet et al.
2012
Chen etal. 2012
Cross-sectional
Cross-sectional
Cross-sectional
Cross-sectional
(C8)
Cross-sectional
(C8)
Cross-sectional
Cross-sectional
Cross-sectional
Cross-sectional
Cross-sectional
Cross-sectional
Cross-sectional
1,400
1,743
1,555
1,505
1,330 and
1,129
293
101
428
252
849
395
429
0.0056 ug/mL
0.00166 ug/mL
0.00678 ug/mL
0.0488 ug/mL
0.03 1-0.0337 ug/mL
0.0016 ug/mL (cord
blood)
0.00254 ug/mL
(maternal at 24-28
weeks)
0.00224 ug/mL
(maternal at delivery)
0.0019 ug/mL
(umbilical cord
blood)
0.0014 ug/mL
0.0021 ug/mL
0.0021 ug/mL
0.0037 ug/mL
0.0018 ug/mL
<-> (gestation length)
I (length of
breastfeeding)
NM
<-> (preterm birth,
congenital anomalies,
labor/delivery
complications,
maternal risk)
<-> (miscarriage)
<-> (preterm,
miscarriage)
<-> (gestational age)
NM
NM
<-> (gestation length)
NM
NM
NM
| (weight)
| (size)
<-> (Apgar score)
NM
<-> (weight)
<-> (low weight)
<-> (low weight,
birth weight)
I (weight, head
circumference,
ponderal index)
<-> (weight)
<-> (weight and
size)
<-> (weight)
<-> (birth weight)
I (birth weight)
<-> (birth weight)
NM
NM
NM
NM
NM
NM
NM
NM
NM
NM
NM
NM
t (TTP)
t (infertility)
t (TTP)
t (infertility)
NM
NM
t (hypertension)
NM
NM
NM
NM
NM
NM
NM
Perfluorooctanoic acid (PFOA) - May 2016
B-12
-------
Study
Study Type
n
Mean Serum PFOA
Outcome
Measures at
Birth
Growth/
Development
Fecundity/
Fertility
Male Fertility
Joensen et al.
2009
(PFOA/PFOS
combined)
Joensen et al.
2013
Buck Louis et al.
2015
Cross-sectional
Cross-sectional
Cross-sectional
105
247
462
0.0049 ug/mL
0.0035 ug/mL
0.00429-0.00509
ug/mL
NM
NM
NM
NM
NM
NM
NM
NM
NM
I (normal sperm)
<-> (testosterone)
<-> (semen
parameters)
<-> (testosterone,
hormones)
t (lower % sperm
with coiled tail)
(total of six PFAS
associated with
changes in sperm
quality)
Neurodevelopmental Endpoints
Fei et al. 2008b
Lieu etal. 2014
Fei and Olsen
2011
Hey er etal. 201 5a
Stein etal. 2013
Hoffman et al.
2010
Cross-sectional
Cross-sectional
Cross-sectional
Cross-sectional
Cross-sectional
(C8)
Cross-sectional
(NHANES)
1,400
156 cases
550 controls
787
(behavior)
537
(coordina-
tion)
1,106
321
571 children
0.0056 ug/mL
0.00456 ug/mL
0.0057 ug/mL
0.0014 ug/mL
0.0351 ug/mL (child)
0.0044 ug/mL
NM
NM
NM
NM
NM
NM
NM
NM
NM
NM
NM
NM
<-> (motor skills
and mental
develop, at 6 and
18 months)
t (cerebral palsy in
boys)
<-> (behavior and
coordination at 7
years)
<-> (motor skills)
t (hyperactivity)
<-> (behavioural
problems)
t (executive
function; ADHD
from mother, not
teacher)
t (ADHD)
NM
NM
NM
NM
NM
NM
Perfluorooctanoic acid (PFOA) - May 2016
B-13
-------
Study
Study Type
n
Mean Serum PFOA
Outcome
Measures at
Birth
Growth/
Development
Fecundity/
Fertility
Postnatal Development
Andersen et al.
2010
Andersen et al.
2013
Heyeretal. 2015b
Lopez-Espinosa et
al. 2011
Christensen et al.
2011
Kristensen et al.
2013
Vested etal. 2013
Halldorsson et al.
2012
Cross-sectional
Cross-sectional
Cross-sectional
Cross-sectional
(C8)
Cross-sectional
Cross-sectional
Cross-sectional
Cross-sectional
1,010
811 (children
at age 7
years)
1,022
3, 076 boys
2,931 girls
448 girls
343 women
169 men
665
0.0052 ug/mL
0.0053 ug/mL
0.001-0.0018 ug/mL
0.02-0.026 ug/mL
0.0036-0.0039
ug/mL (maternal)
0.0036 ug/mL
(maternal)
0.0038 ug/mL
(maternal)
0.0037 ug/mL
NM
NM
NM
NM
NM
NM
NM
NM
NM
NM
NM
NM
NM
NM
NM
NM
I (weight and BMI
in boys at 5 and 12
months)
<-> (height, weight,
waist
measurement, risk
of overweight)
<-> (overweight)
t (waist-to-height
ratio)
t (delayed puberty
in girls)
<-> (age at
menarche)
t (delayed
puberty)
t (lower sperm
cone and total
count)
t (overweight in
females at 20
years)
NM
NM
NM
NM
NM
NM
NM
NM
Notes: | = positive association; J, = negative association; <-» = no association; NM = Not Measured
Perfluorooctanoic acid (PFOA) - May 2016
B-14
-------
Appendix C: Multistage Model for Leydig Cell Tumors
Multistage Cancer Model. (Version: 1.9; Date: 05/26/2010)
Input Data File: C:/IData/MyFiles/PFOA-PFOS/PFOA Docs/msc_Leydig_Opt.(d)
Gnuplot Plotting File: C:/IData/MyFiles/PFOA-PFOS/PFOA Docs/msc_Leydig_Opt.pit
Thu May 09 11:59:27 2013
BMDS_Model_Run
The form of the probability function is:
P [response] = background + (1-background) * [1-EXP (-betal*dose/xl-beta2*dose/x2) ]
The parameter betas are restricted to be positive
Dependent variable = Col2
Independent variable = Coll
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: le-008
Parameter Convergence has been set to: le-008
Default Initial Parameter Values
Background = 0.0132945
Beta(l) = 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(l)
Background 1 -0.64
Beta(l) -0.64 1
Parameter Estimates
95.0% Wald Confidence Interval
Variable Estimate Std. Err. Lower Conf. Limit Upper Conf. Limit
Background 0.00409839 * * *
Beta(l) 0.0116288 * * *
Beta(2) 0 * * *
Indicates that this value is not calculated.
Perfluorooctanoic acid (PFOA) - May 2016 C-1
-------
Model
Full model
Fitted model
Reduced model
AIC:
Log(likelihood)
-28.6454
Analysis of Deviance Table
# Param's Deviance
3
-29.3468
-34.0451
62.6936
1.40286
10.7995
Test d.f.
1
2
P-value
0.2362
0.004518
ChiA2 = I.-
Goodness of Fit
Dose
0.0000
1.3000
14.2000
Est. Prob.
0.0041
0.0190
0.1557
Expected
0.205
0.952
7.784
Observed
0.000
2.000
7.000
Size
50
50
50
Scaled
Residual
-0.454
1.084
-0.306
d.f. = 1
P-value = 0.2245
Benchmark Dose Computation
Specified effect
Risk Type
Confidence level
BMD
BMDL
BMDU
0.04
Extra risk
0.95
3.51044
1.99346
10.7788
Taken together, (1.99346, 10.7788) is a 90 % two-sided confidence interval for the BMD
Multistage Cancer Slope Factor = 0.0200656
Perfluorooctanoic acid (PFOA) - May 2016
C-2
-------
Multistage Cancer Model with 0.95 Confidence Level
5
I
C
o
13
ro
0.3
0.25
0.2
0.15
0.1
0.05
Multistage Cancer
Linear extrapolation
6 8
dose
10
12
14
11:5905/092013
Multistage Cancer Model. (Version: 1.9; Date: 05/26/2010)
Input Data File: C:/IData/MyFiles/PFOA-PFOS/PFOA Docs/msc_Leydig_Opt.(d)
Gnuplot Plotting File: C:/IData/MyFiles/PFOA-PFOS/PFOA Docs/msc_Leydig_Opt.pit
Thu May 09 12:05:42 2013
BMDS_Model_Run
The form of the probability function is:
P[response] = background + (1-background)*[1-EXP(-betal*dose/xl)]
The parameter betas are restricted to be positive
Dependent variable = Col2
Independent variable = Coll
Total number of observations = 3
Total number of records with missing values = 0
Total number of parameters in model = 2
Total number of specified parameters = 0
Degree of polynomial = 1
Maximum number of iterations = 250
Relative Function Convergence has been set to: le-008
Parameter Convergence has been set to: le-008
Perfluorooctanoic acid (PFOA) - May 2016
C-3
-------
Default Initial Parameter Values
Background = 0.0132945
Beta(l) = 0.0097738
Asymptotic Correlation Matrix of Parameter Estimates
Background
Background
Beta(l)
1
-0.64
Beta(l)
-0.64
1
Variable
Background
Beta(l)
Estimate
0.00409839
0.0116288
Parameter Estimates
95.0% Wald Confidence Interval
Std. Err.
Lower Conf. Limit Upper Conf. Limit
Indicates that this value is not calculated.
Analysis of Deviance Table
Model
Full model
Fitted model
Reduced model
AIC:
Log(likelihood)
-28.6454
-29.3468
-34.0451
62.6936
# Param' s
3
2
1
Deviance
1.40286
10.7995
Test d.f.
1
2
P-value
0.2362
0.004518
ChiA2 = I.-
Goodness of Fit
Dose
0.0000
1.3000
14.2000
Est. Prob.
0.0041
0.0190
0.1557
Expected
0.205
0.952
7.784
Observed
0.000
2.000
7.000
Size
50
50
50
Scaled
Residual
-0.454
1.084
-0.306
d.f. = 1
P-value = 0.2245
Benchmark Dose Computation
Specified effect
Risk Type
Confidence level
BMD
BMDL
BMDU
0.04
Extra risk
0.95
3.51044
1.99346
8.7003
Taken together, (1.99346, 8.7003) is a 90% two-sided confidence interval for the BMD
Multistage Cancer Slope Factor = 0.0200657
Perfluorooctanoic acid (PFOA) - May 2016
C-4
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