Environmental Protection
Agency
Health Effects Document
for Perfluorooctanoic
Acid (PFOA)
Perfluorooctanoic Acid - February 2014
Draft - Do Not Cite or Quote
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Health Effects 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: 822R14001
Date: February 2014
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ACKNOWLEDGMENT
This document was prepared under the U.S. EPA Contract No. DW-8992342701, Work
Assignment No. 2011-001 with Oak Ridge National Laboratory. The Lead U.S. EPA Scientist is
Joyce Morrissey Donohue, Ph.D., Health and Ecological Criteria Division, Office of Science and
Technology, Office of Water.
The Oak Ridge National Laboratory is managed and operated by UT-Battelle, LLC., for the U.S.
Department of Energy under Contract No. DE-AC05-OOOR22725.
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Authors, Contributors, and Reviewers
Chemical manager
Joyce Morrissey Donohue, Ph.D.
Office of Water, Office of Science and Technology
Health and Ecological Criteria Division
U.S. Environmental Protection Agency, Washington D.C.
Authors (EPA)
Joyce Morrissey Donohue, Ph.D.
Office of Water, Office of Science and Technology
Health and Ecological Criteria Division
U.S. Environmental Protection Agency, Washington D.C.
Tina Moore Duke, M.S.
Office of Water, Office of Science and Technology
Health and Ecological Criteria Division
U.S. Environmental Protection Agency, Washington D.C.
Authors (Oak Ridge National Laboratory)
Jennifer Rayner, Ph.D., D.A.B.T.
Environmental Sciences Division
Oak Ridge National Laboratory, Oak Ridge, TN
Carol S. Wood, Ph.D., D.A.B.T.
Environmental Sciences Division
Oak Ridge National Laboratory, Oak Ridge, TN
Peer Reviewers
Internal
Christopher Lau, Ph.D.
National Health and Environmental Effects Research Laboratory, Office of Research and
Development
Reproductive Toxicology Division
U.S. Environmental Protection Agency, Research Triangle Park, NC
John Wambaugh, Ph.D.
National Center for Computational Toxicology, Office of Research and Development
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Systems Models for Chemical Toxicity and Exposure
U.S. Environmental Protection Agency, Research Triangle Park, NC
National Center for Environmental Assessment
Office of Research and Development
U.S. Environmental Protection Agency, Research Triangle Park, NC
External
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TABLE OF CONTENTS
ACKNOWLEDGMENT iii
Authors, Contributors, and Reviewers iv
TABLE OF CONTENTS vi
LIST OF TABLES viii
LIST OF FIGURES x
ABBREVIATIONS AND ACRONYMS xi
1.0 EXECUTIVE SUMMARY 1-1
2.0 IDENTITY: CHEMICAL AND PHYSICAL PROPERTIES 2-1
3.0 TOXICOKINETICS 3-1
3.1 Absorption 3-1
3.2 Distribution 3-3
3.3 Metabolism 3-22
3.4 Excretion 3-22
3.4.1 Mechanistic Studies of Renal Excretion 3-26
3.5 Toxicokinetic Considerations 3-33
3.5.1 PK Models 3-33
3.5.2 Half-Life Data 3-41
3.5.3 Volume of Distribution Data 3-45
4.0 HAZARD IDENTIFICATION 4-1
4.1 Human Effects 4-1
4.1.1 Long-Term and Epidemiological Studies 4-1
4.1.1.1 Non Cancer Systemic Toxicity Studies 4-2
4.1.1.2 Thyroid Effects 4-14
4.1.1.3 Steroid Hormones 4-17
4.1.1.4 Reproductive & Developmental Endpoints 4-20
4.1.2 Cancer 4-27
4.2 Animal Studies 4-30
4.2.1 Acute Toxicity 4-30
4.2.2 Short-Term Studies 4-30
4.2.3 Subchronic Studies 4-41
4.2.4 Neurotoxicity 4-44
4.2.5 Developmental/Reproductive Toxicity 4-46
4.2.6 Chronic Toxicity 4-65
4.2.7 Carcinogenicity 4-68
4.3 Other Key Data 4-71
4.3.1 Mutagenicity and Genotoxicity 4-71
4.3.2 Immunotoxicity 4-73
4.3.3 Hormone Disruption 4-83
4.3.4 Physiological or Mechanistic Studies 4-89
4.3.5 Structure-Activity Relationship 4-100
4.4 Hazard Characterization 4-101
4.4.1 Synthesis and Evaluation of Major Noncancer Effects 4-101
4.4.2 Synthesis and Evaluation of Carcinogenic Effects 4-111
4.4.3 Mode of Action and Implications in Cancer Assessment 4-113
4.4.4 Weight of Evidence Evaluation for Carcinogenicity 4-119
4.4.5 Potentially Sensitive Populations 4-119
5.0 DOSE-RESPONSE ASSESSMENT 5-1
5.1 Dose-Response for Noncancer Effects 5-1
5.1.1 RfD Determination 5-1
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5.1.1.1 Benchmark Dose Approach 5-6
5.1.1.2 Pharmacokinetic Model approach 5-9
5.1.1.3 RfD Quantification 5-18
5.1.2 RfC Determination 5-23
5.2 Dose-Response for Cancer Effects 5-24
6.0 REFERENCES 6-1
APPENDIX A 1
APPENDIX B 1
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LIST OF TABLES
TABLE 2- 1 Chemical and Physical Properties of PFOA and APFO 2-3
TABLE 3-1. Protein Binding in Rat, Human, and Monkey Plasma 3-3
TABLE 3-2. Dissociation Constants (Kd) of Binding Between PFOA and Albumin 3-4
TABLE 3-3.Tissue Distribution of PFOA in Wistar Rats After 28 Days of Treatment 3-7
TABLE 3-4. Distribution of PFOA in Male Sprague-Dawley Rats After Oral Exposure 3-9
TABLE 3-5. Distribution of PFOA in Female Sprague-Dawley Rats After Oral Exposure 3-10
TABLE 3-6. PFOA Concentration in Wild-type and PPARa-null Mice 3-12
TABLE 3-7. Plasma PFOA Concentrations ((ig/ml) in Postweaning Sprague-Dawley Rats 3-13
TABLE 3-8. Plasma PFOA Concentrations in Male Rats 3-13
TABLE 3-9. Plasma PFOA Concentrations in Female Rats 3-14
TABLE 3-10. Maternal Plasma PFOA Levels ((ig/ml) During Gestation and Lactation 3-15
TABLE 3-11. Placenta, Amniotic Fluid, and Embryo/Fetus PFOA Concentrations ((ig/ml) 3-16
TABLE 3-12. Fetus/Pup PFOA Concentration (fig/ml) During Gestation and Lactation 3-16
TABLE 3-13. PFOA Levels ((ig/ml) in Sprague-Dawley Maternal Milk During Lactation 3-16
TABLE 3-14. PFOA Levels (ng/ml) During Gestation and Lactation in Selected Fluids and Tissues...3-18
TABLE 3-15. Female Offspring PFOA Levels (ng/ml) After GD1-17 Exposure 3-19
TABLE 3-16. Female Offspring Serum PFOA Levels (ng/ml) After GD 10-17 Exposure 3-20
TABLE 3-17. Serum PFOA Levels (ng/ml) Over Three Generations 3-20
TABLE 3-18. Urine PFOA Concentrations in Male and Female Rats 3-24
TABLE 3-19. Cumulative Percent 14C-PFOA Excreted in Urine and Feces by Rats 3-24
TABLE 3-20. Cumulative Percent 14C-PFOA Excreted in Urine and Feces 3-25
TABLE 3-21. Kinetic Parameters of PFC Transport by OAT1, OAT3 and OATplal 3-31
TABLE 3-22. Plasma and Urine PFOA Concentration 24 hr After Treatment with 30 mg/kg PFOA...3-32
TABLE 3-23. Model Parameters for 1 and 10 mg/kg Single Doses of PFOA to GDI Mice 3-35
TABLE 3-24. Pharmacokinetic Parameters in Male Rats Following Administration of PFOA 3-44
TABLE 3-25. Pharmacokinetic Parameters in Female Rats Following Administration of PFOA 3-45
TABLE 4-1. Association of Serum PFOA with Serum Lipids and Uric Acid in Studies of Occupational
Populations 4-7
TABLE 4-2. Association of Serum PFOA with Serum Lipids and Uric Acid in Studies of General
Populations 4-9
TABLE 4-3. Associations of Serum PFOA with Serum Clinical Biochemistry and Hematology Measures
4-11
TABLE 4-4. Association of serum PFOA with the prevalence of thyroid disease and thyroid hormone
levels in studies of general and worker populations 4-17
TABLE 4-5. Hepatic Effects of Rats Exposed to PFOA 4-33
TABLE 4-6. Hepatic Effects in PFOA-treated Mice 4-36
TABLE 4-7. Mouse Hepatocyte Ultrastructure After PFOA or Wythe 14,643 Treatment 4-37
TABLE 4-8. Relative Response of hPPARa, mPPARa, and PPARa-null mice to PFOA 4-38
TABLE 4-9. Liver Effects In Male Rats 4-44
TABLE 4-10. Organ weight data from FO males 4-47
TABLE 4-11. Organ weight data from Fl males 4-49
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TABLE 4-12. Overview of studies in which pregnant CD-I mice were administered PFOA 4-56
TABLE 4-13. Liver weight data in monkeys administered PFOA for 6 months 4-66
TABLE 4-14. Incidence of Ovarian Stromal Hyperplasia and Adenoma in Rats 4-69
TABLE 4-15. Mammary Gland Tumor Incidence Comparison 4-70
TABLE 4-16. Genotoxicity of PFOA/« Vitro 4-72
TABLE 4-17. Selected Clinical Chemistry Parameters in Mice Treated with PFOA for 5 Days 4-78
TABLE 4-18. Selected Clinical Chemistry Parameters in Mice Treated with PFOA for 10 Days 4-79
TABLE 4-19. Impact of PFOA on Splenic and Thymic Lymphocyte Populations 4-80
TABLE 4-20. Estimated EC50 Values 4-86
TABLE 4-21. Data Collection for Female Mice Gestationally-exposed to PFOA 4-87
TABLE 4-22. mRNA Expression of Hepatic PPARa and Related Genes 4-92
TABLE 4-23. Activation of Mouse and Human PPAR by PFOA 4-95
TABLE 4-24. Impact of PFOA exposure on PC12 cells 4-97
TABLE 5 -1. NOAEL/LOAEL data for Oral Subchronic and Chronic Studies of PFOA 5-3
TABLE 5-2. Shorter term and Developmental Oral Exposure Studies 5-5
TABLE 5-3. Benchmark Dose Modeling fora 10% Increase in Liver Weight 5-7
TABLE 5-4. Description of prior distributions used to analyze all but the C57BL/6N mice 5-11
TABLE 5-5. Estimated and assumed pharmacokinetic parameters used in model development 5-12
TABLE 5-6. Predicted Final Serum Concentration and Time-integrated Serum Concentration (AUC) for
Studies in Rats 5-13
TABLE 5-7. Predicted Final Serum Concentration and Time-integrated Serum Concentration (AUC) for
Studies in Mice 5-14
TABLE 5-8. Predicted Final Serum Concentration and Time-integrated Serum (AUC) in Studies of
Monkeys 5-15
TABLE 5-9. Average Serum concentrations Derived from the AUC and the duration of Dosing 5-16
TABLE 5-10. Human Equivalent Doses Derived from the Modeled Animal Average Serum Values ..5-18
TABLE 5-11. RfD Point of Departure Options (mg/kg/day) from the PFOA Animal Studies 5-19
TABLE 5-12. The Impact of Quantification Approach on the RfD outcome for the PODs from the
Palazzolo et al (1993) Study 5-20
TABLE 5-13. The Impact of Quantification Approach on the RfD Outcome for PODs from the BMDLs
5-21
TABLE 5-14. The Impact of Quantification Approach on the RfD Outcomes for the HEDs from the
Pharmacokinetic Model Average Serum Values 5-22
TABLE 5-15. Summary of Tumor Data from Animal Studies 5-25
TABLE 5-16. Model Results for Leydig Cell Tumors (Buttenhoff et al. (2012) 5-27
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LIST OF FIGURES
FIGURE 2-1 Chemical Structures of PFOA and APFO 2-1
FIGURE 2-2 PFOA Anti-Conformer 2-1
FIGURE 2-3 PFOA Lowest Energy Conformer 2-2
FIGURE 3-1. PFOA binding sites on human serum albumin 3-5
FIGURE 3-2. Renal Organic Acid Transporters 3-27
FIGURE 3-3. Schematic for a physiologically-motivated renal resorptions pharmacokinetic model 3-33
FIGURE 3-4. Physiologically-motivated pharmacokinetic model schematic for PFOA-exposed rats ...3-34
FIGURE 3-5. Schematic for one compartment model 3-35
FIGURE 3-6. Pharmacokinetic model of gestation and lactation in mice 3-37
FIGURE 3-7. Structure of the PFOA PBPK model in monkeys and humans 3-38
FIGURE 3 -8. Structure of the PBPK Model for PFOA in the Adult Sprague Dawley Rat 3-40
FIGURE 4-1. PPARa-Agonism Mode of Action for Liver Tumors 4-114
FIGURE 5-1. BMDS graphic output from selected model runs 5-8
FIGURE 5-2. BMD Model Results for Leydig Cell Tumors (Buttenhoff etal., 2012) 5-27
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ABBREVIATIONS AND ACRONYMS
ADAF age-Dependent Agjustment Factor
ADHD attention deficit hyperactivity disorder
ADX adrenalectomized
AIC Akaike's Information Criterion
ALP alkaline phosphatase
ALT alanine transaminase
ANOVA analysis of variance
Areg amphiregulin
AP alkaline phosphatase
AST aspartate aninotransferase
ATSDR Agency of Toxic Substances and Disease Registry
AUC area under the plasma concentration time curve
AUCinf area under the plasma concentration time curve, extrapolated to infinity
AWWA American Water Works Association
AwwaRF American Water Works Association Research Foundation
BMD benchmark dose
BMDL benchmark dose - lower 95th percentile confidence bound
BMDS benchmark dose software
BMI body mass index
BMR benchmark response
BQL below quantifiable limit
BSA bovine serum albumin
BSEP bile salt export pump
BUN blood urea nitrogen
bw body weight
C Celsius
Cal-EPA California Environmental Protection Agency
CaMKII calcium/calmodulin-dependent protein kinase II
CAR constitutive androstane receptor
CAS Chemical Abstracts Service
CAT carnitine aceyltransferase
CCK cholecystokinin
CCL Contaminant Candidate List
CCL 3 Contaminant Candidate List 3
CDC Centers for Disease Control and Prevention
CFSE 6-carboxyfluorescein succinimidyl ester
ChAT choline acetyltransferase
CHO Chinese hamster ovary
CI confidence interval
Clp plasma clearance
CL clearance
CLR renal clearance
Cmax peak plasma concentration at the first intestinal absorption loci
CoA coenzyme A
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ConA concanavalinA
CORT corticosterone
Cox II cytochrome c oxidase subunit II
Cox IV cytochrome c oxidase subunit IV
CPT carnitine palmitoyltransferase
CRL Charles River Laboratory
CSF Cancer Slope Factor
CWS community water system
d day
DCDQ Developmental Coordination Disorder Questionnaire
DCFH-DA 2',7'-dichlorofluorescein diacetate
DBS diethylstilbestrol
DHT 5a-dihydroxy-testosterone
DMSO dimethyl sulfoxide
DNA Deoxyribonucleic acid
DTH delayed-type hypersensitivity
DWI drinking water intake
E2 17-p estradiol
E3 S estrone-3 -sulfate
ECso half maximal effective concentration
ECF Electro-Chemical Fluorination
EFSA European Food Safety Authority
EGFR epidermal growth factor receptor
ELISA enzyme-linked immunosorbent assay
ER endpplasmic reticulum
ERa estrogen receptor a
Erra estrogen-related receptor a
FID flame ionization detector
FSH follicle-stimulating hormone
FXR farnesoid receptor
g gram
GAP-43 growth-associated protein-43
GD gestation day
GGT gamma-glutamyl transpeptidase
GJIC gap junction intercellular communication
GlyT glycogen tropholblast cell
GnRH gonadotropin releasing hormone
GSD geometric standard deviation
hCG human chorionic gonadotropin
HDL high density lipoprotein
HED human equivalent dose
HEK human embryonic kidney
HET heterozygous
HGFa hepatocyte growth factor
HL-60 human promyelocytic leukemia cell line
HMG-CoA 3-hydroxy-3-methylglutaryl coenzyme A
HRBC horse red blood cells
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HRL health reference level
HSA human Serum Albumin
HSDl?pl hydroxysteroid 1?P dehydrogenase 1
HSD3P1 hydroxy steroid 3p dehydrogenase 1
HSDB Hazardous Substances Database
ICso half-maximal Inhibiting Concentration
ICR imprinting control region
IGF-I insulin like growth factor I
IHD ischemic heart disease
IL-6 interleukin 6
IRR incidence rate ratio
IRIS Integrated Risk Information System
IU international unit
IV intravenous
Ka adsorption rate constant
Kd dissociation constant
Ke elimination rate constant
kg kilogram
Km substrate concentration at which the initial reaction rate is half maximal
Koc organic carbon water partitioning coefficient
Kt affinity constant
L liter
LCso lethal concentration for 50% of animals
LC-MS liquid chromatography - mass spectrometry
LCT Leydig cell tumors
LD lactation day
LD50 lethal dose for 50% of animals
LDH lactic dehydrogenase
LDL low density lipoprotein
L-FABP liver fatty acid binding protein
LH luteinizing hormone
LHWA Little Hocking Water Association
LLOQ lower limit of quantification
LOAEL lowest observed adverse effect level
LOQ Limit of Quantitation
LPS lipopolysaccharide
m meter
MCAD medium chain acyl-CoA dehydrogenase
MCLG Maximum Contaminant Level Goal
Mdr2 mulidrug resistance protein 2
jig microgram
mg milligram
min minute
mL milliliter
|im micrometer
MMAD mass median aerodynamic diameter
MOA mechanism of action
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mol mole
mPL mouse placental lactogen
mPLP mouse prolactin like protein
MRL minimum reporting level
MRP multidrug restistance-associated protein
MTBE methyl tertiary-butyl ether
MTT 3-(4,5-dimethylthiazol-2-yl)-2,5-diphenyltetrazolium bromide
NAWQA National Water Quality Assessment
NDWAC National Drinking Water Advisory Council
ng nanogram
NAS National Academy of Sciences
ND not detected, no differences, or no data
Nd2 NADH dehydrogenase 2
NdufsS NADH dehydrogenase iron-sulfur protein 8
NHANES The National Health and Nutrition Examination Survey
NIRS National Inorganics and Radionuclides Survey
NK natural killer
NM not monitored
NMR Nuclear Magnetic Resonance
NMRI Naval Medical Research Institute
NOAEL no observed adverse effect level
NPDWR National Primary Drinking Water Regulation
NRC National Research Council
Nrfl nuclear respiratory factor 1
Nrf2 nuclear respiratory factor 2
NTCP sodium-taurocholate contransporting polypeptide
OAT organic anion transporter
OATp organic anion transporting peptide
OGWDW Office of Ground Water and Drinking Water
OPP-RED Office of Pesticide Programs - Reregi strati on Eligibility Decision
OR odds ratio
OVA ovalbumin
OVX ovariectomized
OW Office of Water
P progesterone
PACT pancreatic acinar cell tumors
PAH polycyclic aromatic hydrocarbon
PB phenobarbital
PBMC peripheral blood mononuclear cells
PBPK pharmacologically based pharmacokinetic
PCNA proliferating cell nuclear antigen
PenH enhanced pause airway respiration
PFAA perfluoroalkyl acid
PFC perfluorinated carboxylate acids
PFDA perfluorodecanoic acid
PFHxS perfluorohexanesulfonic acid
PFOA perfluorooctanoic acid
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PFOS perfluoroocatane sulfonate
Pgc-la peroxisome proliferator-activated receptor gamma coactivator la
PH peroxisomal bifunctional protein
PHA phytohemagglutinin
PK pharmacokinetic
PND postnatal day
POD point of departure
pKa acid dissociation constant
PPAR peroxisome proliferator activated receptor
Ppb parts per billion
ppm parts per million
PT peroxisomal thiolase
PWG Pathology Working Group
PWS public water system
PXR pregnane X receptor
Q flow in and out of tissues
Qjiic median fraction of blood flow to the filtrate
Reg Det 2 Regulatory Determinations on the Second CCL
RfC reference concentration
RfD reference dose
RNA ribonucleic acid
ROS reactive oxygen species
RR relative risk
RSA rodent serum albumin
RSC relative source contribution
RTECS The Registry of Toxic Effects of Chemical Substances
RT-PCR reverse transcription polymerase chain reaction
RXRa retinoid X receptor alpha
SD standard deviation
SDH sorbitol dehydrogenase
SDQ Strengths and Difficulties Questionnaire
SDWA Safe Drinking Water Act
SHBG sex hormone binding globulin
STAR SIDS Initial Assessment Report
SIR standardized incidence ratio
SMR standardized mortality ratio
SPI Society of the Plastics Industry
SRBC sheep red blood cells
S-TGC sinusoidal trophoblast giant cells
Tmax time of maximum plasma concentration
T3 triiodothyronine
T4 thyroxine
Ti/2 elimination half-time
Tm transporter maximum
Tmax time of maximum plasma concentration
TC total cholesterol
TCPOBOP l,4-bis[2-(3,5-dichloropyridyloxy)] benzene
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Tfam transcription factor A
TG triglycerides
TH tyrosine hydroxylase
TNFa tumor necrosis factor a
TRR total reactive residues
TSH thyroid stimulating hormone
TIP time to pregnancy
TTR thyroid hormone transport protein, transthyretin
UA uric acid
UCMR Unregulated Contaminant Monitoring Rule
UCMR 1 Unregulated Contaminant Monitoring Rule 1
UCMR 2 Unregulated Contaminant Monitoring Rule 2
UF uncertainty factor
URAT urate trasnporter
U.S. EPA U.S. Environmental Protection Agency
USGS U.S. Geological Service
Vd volume of distribution
Vmax maximum initial rate of an enzyme catalysed reaction
VLCAD very long chain acyl-CoA dehydrogenase
VLDL very low density lipoprotein
VOC volatile organic compound
WHO World Health Organization
WRF Water Research Foundation
WT wild type
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1.0 EXECUTIVE SUMMARY PFOA
Perfluorooctanoic acid (PFOA) is a completely fluorinated organic synthetic acid used to
produce fluoropolymers. Because of strong carbon-fluorine bonds, PFOA is stable to metabolic
and environmental degradation and is resistant to biotransformation. Extensive data in humans
and animals demonstrate ready absorption of PFOA and distribution of the chemical throughout
the body by noncovalent binding to plasma proteins. The liver is an important binding site with
increased liver weight in laboratory animals one of the early, low-dose manifestations of
exposure. At least three transport families seem to play a role in PFOA absorption, distribution,
and excretion: organic anion transporters (OATs), organic anion transporting peptides (OATps)
and multidrug resistance-associated proteins (MRPs). PFOA is not readily eliminated from
humans as evidenced by the half-life of 2.3 years. In contrast, half-life values for the monkey,
rat, and mouse are 20.8 days, 11.5 days, and 15.6 days, respectively. Differences in transporters
may explain species differences in elimination.
PFOA is known to activate the peroxisome proliferator activated receptor (PPAR)
pathway by increasing transcription of mitochondrial and peroxisomal lipid metabolism, sterol,
and bile acid biosynthesis and retinol metabolism genes. However, based on transcriptional
activation of many genes in PPARa null mice, the effects of PFOA involve more than activation
of PPAR. Also activated are the constitutive androstane receptor (CAR), farnesoid receptor
(FXR), and pregnane X receptor (PXR).
Epidemiology studies have examined occupational and residential populations at or near
large-scale PFOA production plants in the United States in an attempt to determine the
relationship between serum PFOA concentration and various health outcomes suggested by the
standard animal toxicological studies. Exposures were mainly through contaminated drinking
water and to multiple PFCs. These studies have found a positive association between serum
PFOA concentration and higher cholesterol in the general population and in worker populations
but no consistent trends for the low-and high-density protein lipids. A positive association has
been shown between serum PFOA concentrations and increased liver enzymes and/or decreased
bilirubin in both worker and general populations, chronic kidney disease in the general
population, and the odds of experiencing early menopause. Maternal or child plasma levels of
PFOA were positively associated with decreased antibody liters in children after vaccination,
obesogenic effects in female children at 20 years of age, and parent reported Attention Deficit
Hyperactivity Disorders.
No consistent associations were identified between serum PFOA and hyperglycemia,
type II diabetes, thyroid homeostasis, fertility, fecundity, and birth outcome. Maternal or child
plasma PFOA levels were not associated with changes in children reaching motor or mental
developmental milestones, child risk of hospitalization for infectious diseases, or attainment of
puberty.
In most animal studies, short-term and chronic exposure to PFOA resulted in an increase
in liver weight as at least one of the critical effects. Co-occurring effects in these studies
included changes in spleen, thymus, liver and/or developmental endpoints. In repeated dosing
studies, rats given 0.64 mg/kg/day for 13 weeks and monkeys given 3 mg/kg/day for 26 weeks
had increased liver weight accompanied by hepatocellular hypertrophy. As part of a two-
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generation study, male rats had increased liver and kidney weights at 1 mg/kg/day. In shorter-
term studies, slightly higher doses to rats also resulted in increased liver weight and liver lesions.
In mice, developmental toxicity and increased spleen weight at a dose of 1 mg/kg/day
accompanied increased liver weight. Slightly higher doses resulted in decreased
immunoglobulin levels.
U.S. EPA has selected 0.00002 mg/kg/day as the RfD for PFOA based on the consistency
of the response and with recognition of the use of liver weight as a common denominator for loss
of homeostasis and protection against co-occurring adverse effects. This value is the outcome
for modeled serum values from three rat studies and one mouse study. In two of the rat studies,
liver effects were accompanied by developmental effects and kidney weight increases. A
pharmacokinetic model was used to predict a serum area under the curve (AUC) concentration
for each LOAEL and NOAEL for liver weight effects. From each AUC, the average serum
concentration was calculated which was then used to calculate a human equivalent dose (HED).
The total uncertainty factor (UF) applied to the HEDNOAEL from rat study was 30 which included
a UF of 10 for intrahuman variability and a UF of 3 to account for toxicodynamic differences
between animals and humans. For the critical studies that lacked an HEDNOAEL, and additional
10-fold UF was added to adjust for the use of a HED based on a LOAEL. Four of six candidate
studies resulted in the same RfD. All of the studies supporting the RfD, except one, are studies
with exposures to PFOA for > 84 days meeting the duration requirements for determination of a
lifetime exposure value; the exception is a 17-day developmental toxicity study in the mouse.
Under the U.S. EPA 2005 cancer guidelines, the evidence for the carcinogenicity of
PFOA is considered suggestive because only one species has been evaluated and studies for gene
mutations were negative. Epidemiology studies demonstrate an association of serum PFOA with
kidney and testicular tumors among highly exposed members of the general population. The
dose-response data for Ley dig cell tumors in rats in was analyzed using the multistage cancer
model for a dichotomous dataset to predict the dose at which a 4% increase in tumor incidence
would occur. The 4% was chosen as the low-end of the observed response range within the
study results. The resulting BMDL04 was 1.99 mg/kg/day which yields a HED of 0.58
mg/kg/day and a slope factor of 0.07 (mg/kg/day)"1. The HED of 0.58 mg/kg/day is nearly
30,000-fold greater than the RfD. Thus, the proposed RfD, based on increased liver weight as a
common denominator for loss of homeostasis and protection of co-occurring effects, will also be
protective of Ley dig cell tumors.
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2.0 IDENTITY: CHEMICAL AND PHYSICAL PROPERTIES
Perfluorooctanoic acid (PFOA) is a completely fluorinated organic synthetic acid used to
produce fluoropolymers. It is manufactured by the Simons Electro-Chemical Fluorination (ECF)
process or by telomerization. In the ECF process, the carbon-hydrogen bonds on molecules of
organic feedstock are replaced with carbon-fluorine bonds when an electric current is passed
through a solution of hydrogen fluoride and the organic feedstock. In the telomerization process,
fluorine-bearing chemicals and tetrafluoroethylene react to produce fluorinated intermediates
that are converted into PFOA (HSDB, 2006). Ammonium perfluorooctanoate (APFO) is the salt
of PFOA (Figure 2-1) and is a processing aid in the manufacture of certain fluoropolymers,
especially as an emulsifier in aqueous solution during the emulsion polymerisation of
tetrafluoroethylene. APFO is not consumed during the polymerization process (SPI, 2005).
Some sources of PFOA in the atmosphere result from the atmospheric degradation or
transformation or surface deposition of precursors, including related fluorinated chemicals
(fluorotelomer alcohols, olefins, and perfluoroalkyl sulfonamido substances) (Wallington et al.,
2006).
PFOA APFO
FFFFFFF F F F F F
x x x x °~ NH:
FFFFFFF FFFFFFFF
Sources: SIAR (2006); http://commons.wikimedia.Org/wiki/File:Ammonium perfluorooctanoate.png
FIGURE 2-1 Chemical Structures of PFOA and APFO
Although PFOA is not a polar molecule, each of the carbon fluoride bonds is a dipole as a
result of the electronegativity difference between fluoride (4.1) and carbon (2.5) placing a partial
negative charge on each of the covalently bound fluorines and a partial positive charge on each
of the fluorinated carbons. Charge repulsion of the partially negative fluorines and steric factors
favor a PFOA conformation in which carbons 2 through 7 adopt an anti arrangement of
substituents (Figure 2-2) resulting in a fairly linear molecular shape (Figure 2-3) as the lowest
energy conformer.
(CF:
(CF2)X
FIGURE 2-2 PFOA Anti-Conformer
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FIGURE 2-3 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.
Some commercial PFOA is a combination of linear and branched chain isomers (80%
linear, 20% branched; Loveless et al., 2006). The samples studied by Loveless et al. (2006) had
the following mole percents of branched chain isomers: 12.6% internal monomethyl (non-alpha),
9% isopropyl, 0.2%, tert-butyl, 0.1% gem-dimethyl, and 0.1% alpha monomethyl. A study by
Yoshikane et al. (2010) reported finding perfluoro-6-methylheptanoic acid (the isopropyl isomer)
using mass spectroscopy analysis of environmental fluorosurfactants in Japan.
The physical and chemical properties and other reference information for PFOA and its
salt APFO are provided in Table 2-1. These properties help to define the behavior of PFOA in
living systems and the environment. PFOA and its salt are highly stable compounds. They are
solids at room temperature with low vapor pressures. The melting point for PFOA is identified
as 40-50°C and vapor pressures increase at temperatures near the melting point.
PFOA is moderately soluble in water and APFO is even more soluble. Both compounds
are considered insoluble in nonpolar solvents which results in their being described as olephobic.
Water solubility is increased by the presence of other ions and is an important factor governing
solubility in body fluids. As the concentration of PFOA in aqueous solution increases, it forms
colloidal micelles with the carboxyl functional groups on the exterior and the fluorocarbon chain
on the interior. The critical micelle concentration has been identified as 3.6-3.7 g/L.
The pKa (acid dissociation constant) for PFOA has been reported as 1.5-2.8. As a result,
it will be present in most biological fluids (gastric secretions excluded) primarily as the
perfluorooctanoate anion. This is an important feature in governing absorption and membrane
transport.
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TABLE 2- 1 Chemical and Physical Properties of PFOA and APFO
Property
Chemical Abstracts Registry
(CAS) No.
Synonyms
Chemical Formula
Molecular Weight
Color/Physical State
Boiling Point
Melting Point
Density (at 20°C)
Vapor Pressure:
pKa
pH value
Koc
Solubility in water (g-L"1)
Solubility in organic solvents
Conversion Factors for vapor
phase
Perfluorooctanoic Acid
335-67-1
PFOA; Hexanoyl fluoride,
3,3,4,4,5,5,6,6,6-nonafluoro-2-oxo-;
Pentadecafluoro-1-octanoic acid;
Pentadecafluoro-n-octanoic acid;
Octanoic acid, pentadecafluoro-;
Perfluorocaprylic acid;
Pentadecafluoroocanoic acid;
Perfluoroheptanecarboxylic acid;
C8HF1502
414.09
White powder
189°C
45-50 °C
1.7921 g/cm3
4.2 (25°C)
2.3 (20°C)
128 (59.3°C)
2.5
2.8
1.5-2.8
2.6, 1 g/L (20°C)
27,000 estimated
9.5 (25°C)
4.1 (22°C)
-
1 ppm= 17.21 mg/m3
Ammonium Perfluorooctanoate
3825-26-1
APFO; C-8; Ammonium
pentadecafluoroctanoate; Ammonium
perfluorocarpylate; DS 101; FC 1015;
FC 143; FX 1006; Fluorad® FC 143;
Perfluorooctanoic acid ammonium
salt; Unidyne® DS 101-20
C8HF15O2NH3
431.10
White powder
Decomposition
Decomposition starts above 105°C
0.6-0.7 g/cm3
3.7(90.1°C)
0.0081 (20°C)
~5
>500
Heptane, Toluene: 0
Methanol, Acetone: > 500
1 ppm = 17.21 mg/m3
Sources: HSDB (2006); SIAR (2006), EFSA (2008); RTECS (2008)
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3.0 TOXICOKINETICS
Because of strong carbon-fluorine bonds, PFOA is stable to metabolic and environmental
degradation. PFOA, is resistant to biotransformation, making the toxicity of the parent
compound the concern, not that of a metabolite. However, because of their impact on cellular
receptors and proteins, they possess the ability to impact the biotransformation of dietary
molecules, intermediate metabolites, and other xenobiotic chemicals and may alter enzyme
activities and transport kinetics. PFOA is known to activate the peroxisome proliferator
activated receptor (PPAR) pathway by increasing transcription of mitochondrial and peroxisomal
lipid metabolism, sterol, and bile acid biosynthesis and retinol metabolism genes. However,
based on transcriptional activation of many genes in PPARa null mice, the effects of PFOA
involve more than activation of PPAR. They may also activate the constitutive androstane
receptor (CAR), farnesoid receptor (FXR), and pregnane X receptor (PXR).
PFOA is not readily eliminated from humans and other primates. Toxicokinetic profiles
and the underlying mechanism for half-life differences are not completely understood although
many of the differences appear to be related to elimination kinetics and factors that control
membrane transport. Thus far, three transport families seem to play a role in PFOA absorption,
distribution, and excretion: organic anion transporters (OATs), organic anion transporting
peptides (OATps) and multidrug resistance-associated proteins (MRPs) (Launay-Vacher et al.,
2006). The transporters play a critical role in gastrointestinal absorption, uptake by the tissues,
and excretion via the kidney. These transport systems are located at the blood brain barrier,
blood placental barrier, blood testes barrier, and mammary gland where they function to protect
the brain, fetus, and male reproduction organs (Ito and Alcorn, 2003; Zai'r et al., 2008). There
are differences in transporters across species, genders, and individuals. There is more PFOA-
specific information regarding the OAT and OATp families than the MRPs. Nomenclature
abbreviations for the various transporters are not totally standardized, and there are
inconsistencies across individual publications.
3.1 Absorption
Absorption data are available for oral, inhalation, and dermal exposure in laboratory
animals. There are extensive data from humans demonstrating the presence of PFOA in serum.
These data demonstrate absorption by one or more routes but do not quantify the amounts
absorbed relative to dose. The human serum data are included in Section 3.5 of this report.
The absorption processes requires transport across the interface of the gut, lung, and skin
with the external environment. Since PFOA is moderately soluble in aqueous solutions and
oleophobic (minimally soluble in body lipids), movement across the apical and basal membrane
surfaces of the lung, gastrointestinal tract, and skin involves transporters or mechanisms other
than simple diffusion across the lipid bilayer. As discussed above, there are data that identify
involvement of OATs, OATps, and MRPs in entrocytes (Zair et al., 2008). OAT2, OAT3,
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). No primary data sources relative to the specificity of these transporters for
PFOA in humans or laboratory animals were identified.
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Oral Exposure
PFOA is well absorbed following oral exposure. Gibson and Johnson (1979)
administered a single dose of 14C-PFOA averaging 11.4 mg/kg by gavage to groups of 3 male
10-week old CD rats. Twenty-four hours after administration, at least 93% of total carbon-14
was absorbed. In another study, Cui et al. (2010) exposed male Sprague-Dawley rats (10/group)
to PFOA (96% a.i.) at 0, 5, and 20 mg/kg/day once daily by gavage for 28 days. The percent of
the dose absorbed was 92.8% and 92.3% for the low and high dose, respectively under the
assumption that fecal excretion over the first 24-hours after dosing was estimated to be
unabsorbed material.
Steady state serum PFOA levels were reached within 4-6 weeks in cynomolgus monkeys
dosed with oral capsules containing 3, 10, or 20 mg/kg PFOA for 6 months (Butenhoff et al.,
2004b). Urine steady state levels were reached after 4 weeks. Serum PFOA concentration in
male rats fed diets containing 0.06, 0.64, 1.94, or 6.5 mg PFOA/kg for 90 days was 7.1, 41.2,
70.3, and 137.6 ug/mL, respectively (Palazzolo, 1993). Peak blood levels of PFOA were
attained 1-2 hours following a 25 mg/kg dose to male and female rats (Kennedy et al., 2004).
Blood levels of PFOA over time were similar in female rats given a single dose of 25 mg
PFOA/kg when compared to a female rat given 10 daily doses of 25 mg PFOA/kg (Kennedy et
al., 2004). Plasma PFOA concentrations in male Sprague-Dawley rats fed a diet containing 300
ppm PFOA for 1, 7, or 28 days were 259, 234, and 252 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. Correspondingly, female rats show much
greater urinary excretion of PFOA than do male rats with serum half-life values in hours for
females compared with days for males. These differences are discussed in more detail in the
following sections.
Inhalation Exposure
Hinderliter (2003) measured the serum concentrations of PFOA following single and
repeated inhalation exposures in Sprague-Dawley rats. For the single exposure study, male and
female rats (3/sex/group) were exposed to a single nose-only exposure of an aerosol of 0, 1, 10,
or 25 mg/m3 PFOA. Preliminary range-finding studies demonstrated that aerosol sizes were 1.8 -
2.0 |im mass median aerodynamic diameter (MMAD) with geometric standard deviations
ranging from 1.9-2.1 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 was analyzed
by liquid chromatography-mass spectrometry (LC-MS). PFOA plasma concentrations were
proportional to the inhalation exposure concentrations.
The male Cmax values were approximately 2-3 times higher than the female Cmax. The
female Cmax occurred approximately one hour after the exposure period with plasma
concentrations then declining. In males, Cmax was observed immediately after the exposure
period ended and persisted for up to six hours. The data are illustrative of absorption of PFOA
via inhalation.
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Dermal Exposure
There is evidence that PFOA is absorbed following dermal exposure. Kennedy (1985)
treated rabbits and rats dermally with a total of 10 applications of PFOA at doses of 0, 20, 200 or
2,000 mg/kg. Treatment resulted in elevated blood organofluorine levels that increased in a dose-
related manner. O'Malley and Ebbens (1981) treated groups of 2 male and 2 female New
Zealand White rabbits dermally with doses of 100, 1,000 and 2,000 mg/kg PFOA for 14 days.
Mortality among the exposed animals demonstrated dermal uptake. All of the animals died at
the highest dose, 3 of 4 died in the mid-dose group and none from the low-dose group. Although
these data demonstrate dermal absorption they do not provide quantitative data.
The results of in vitro percutaneous absorption studies of PFOA through rat and human
skin have been reported (Fasano et al, 2005). The permeability coefficient for PFOA was
calculated to be 3.25 ± 1.51 x 10"5 cm/h and 9.49 ± 2.86 x 10"7 cm/h in rat and human skin,
respectively.
3.2 Distribution
Distribution of absorbed material requires vascular transport from the portal of entry to
receiving tissues. It has been suggested that PFOA circulates in the body by noncovalently
binding to plasma proteins. Several studies have investigated the binding of PFOA to plasma
proteins in rats, humans or monkeys to gain an understanding of its absorption, distribution and
elimination, plus information on species and gender differences.
Protein Binding. Protein binding in plasma from cynomolgus monkeys, rats, and humans was
tested with PFOA via in vitro methods (Kerstner-Wood et al., 2003). The results are summarized
in Table 3-1. Rat, human, and monkey plasma proteins were able to bind 97 to 100% of the
PFOA added at concentrations ranging from 1-500 ppm. Human serum albumin carried the
largest portion of the PFOA among the protein components of human plasma. Serum albumin is
a common carrier of hydrophobic materials in the blood including short and medium chain fatty
acids, inorganic ions, and some pharmaceuticals. Approximately 60% of the serum protein in
humans and rats is albumin (Saladin, 2003; Harkness and Wagner, 1983). At 68%, the
percentage bound to albumin in mice is slightly higher than in humans and rats (Harkness and
Wagner, 1983).
TABLE 3-1. Protein Binding in Rat, Human, and Monkey Plasma
PFOA Concentration
(ppm)
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
From Kerstner-Wood et al., 2003
%Binding values reported as "~100" reflect a nonquantifiable amount of test article in the plasma water below the quantifiable
limit (BQL)<6.25 ng/mL
Han et al. (2003) investigated the binding of PFOA to rat and human plasma proteins in
vitro. The authors concluded that there was no correlation between the PFOA persistence and
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binding of the PFOA to rat serum. The primary PFOA binding protein in plasma was serum
albumin. However, the method used (ligand blotting) would not theoretically allow the
identification of low abundance proteins with high affinity for PFOA. Further investigation of
purified rodent and human serum albumin binding using labeled 19F nuclear magnetic resonance
(NMR) allowed the calculation of disassociation constants for PFOA binding to rodent and
human serum albumin. No significant difference in binding to the serum albumin of rat versus
human was detected (Table 3-2).
Male and female rats treated in vivo showed no gender difference in the binding of PFOA
to serum proteins though the persistence of PFOA in vivo is much greater in male than female
rats.
TABLE 3-2. Dissociation Constants (Kd) of Binding Between PFOA and Albumin
Parameter
£d(mM)
£d(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
From Han etal, 2003
RSA= Rodent Serum Albumin; HSA= Human Serum Albumin
a Average of the two Kd values (0.31 ± 0.15 and 0.27 ± 0.05 mM) obtained by NMR.
b Values were obtained from three independent experiments and their standard deviations are shown.
c On the basis of the result of unpaired t-test at 95% confidence interval, the difference of Kd values determined by NMR and micro-
SEC is statistically insignificant.
Wu et al. (2009) examined the interaction of PFOA and human serum albumin (HSA).
The authors tested their hypothesis that PFOA, after absorption, was transported bound to
albumin by dialyzing PFOA solutions in the presence and absence of HSA. In the absence of
HSA, 98% of the dissolved PFOA crossed the dialysis membrane into the dialysate within 4
hours. In the presence of HSA, the amount of PFOA found in the dialysate after 4 hours
decreased in direct proportion to the albumin concentration, demonstrating binding to the
protein. No albumin was identified in the dialysate.
Using the dialysis data and thermodynamic considerations, the authors concluded that
albumin could bind up to 12 PFOA molecules on its surface via chemical monolayer absorption
with a 13th molecule bound noncovalently in the more hydrophobic interior of the protein. The
surface nature of the binding may well indicate potential binding to other serum proteins as well.
Circular dichroism measurements of the albumin/PFOA complex suggested a conformational
change in the protein as a result of the PFOA binding. The beta-pleated sheet content of the
albumin decreased, and the alpha-helix content increased by 15%. These conformational
changes could interfere with the functional properties of serum albumin or other serum proteins
impacted by surface monolayers of PFOA. For example, albumin's ability to transport its natural
ligands could be decreased by the presence of PFOA on the protein surface. The interaction of
albumin with target cellular receptors could also be altered.
MacManus-Spencer et al. (2010) used a variety of approaches to quantify the binding of
PFOA to serum albumin (surface tension measurements, 19FNMR spectroscopy, and
fluorescence spectroscopy). When taken together, the results from these analyses suggest the
presence of primary and secondary binding sites on albumin. The PFOA-albumin association
constants for the primary site (Kia) was about 1.5 ± 0.2 x 105/mol bovine serum albumin (BSA)
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while the association constant for the secondary site (K2a) was 0.8 ± 0.1 x 102/mol BSA at a
concentration of IjiM. When the concentration of BSA increased to 10 |iM, the binding per
mole of BSA decreased: Kia = 0.33 ± 0.004; 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 the most relevant to the potential physiological impact of
environmental levels of PFOA.
A modeling study by Salvalaglio et al. (2010) was conducted to determine the binding
sites of PFOA on HSA and classify them by their interaction energy using molecular modeling;
this study builds on the binding studies of Wu et al. (2009) and MacManus-Spencer et al. (2010).
It was estimated that maximum number of PFOA binding sites on HSA was 9. The site locations
were common to the natural binding sites for fatty acids, thyroxine, warfarin, indol, and
benzodiazepine (Figure 3-1; Salvalaglio et al., 2010). The binding site closest to tryptophan
residue #214 had the highest binding affinity (-8.0 kcal/mol) compared to other sites (—6—1
kcal/mol).
Indolebenzcdiazepine ^ *
site 8 stte 7
L Warfarin
6
FIGURE 3-1. PFOA binding sites on human serum albumin
Weiss et al. (2009) screened several perfluorinated compounds (n=30), differing by
carbon chain length C4-18, fluorination degree, and functional groups for potential binding to the
serum thyroid hormone transport protein, transthyretin (TTR) using a radioligand-binding assay.
The natural ligand of TTR is thyroxine (T4). PFOA was one of the chemicals evaluated.
Human TTR was incubated overnight with 125I-labeled T4, unlabled T4, and 10-10,000
nM PFOA as a competitor for the T4 binding sites. The unlabeled T4 was used as a reference
compound. The levels of T4 in the assay were close to the lower range for total T4 measured in
healthy adults. The authors concluded that binding affinity for TTR was highest for the fully
fluorinated compounds tested and those having at least an eight carbon length chain,
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characteristics that apply to PFOA. PFOA demonstrated a high binding affinity for TTR with
949 nM causing a 50% inhibition of T4 binding to the TTR.
It is also possible that PFOA will display nonspecific binding to proteins within the
cellular matrix as well as in the serum but little work has been done to investigate this
probability. Luebker et al. (2002) conducted in vitro studies of the ability of a variety of
perfluorinated chemicals to displace a fluorescent substrate (1 l-(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 to 5% of the
cytosolic protein in hepatocytes. Luebker et al. (2002) reported that PFOA (IC5o>10|iM)
exhibited some binding to L-FABP, but the binding potential was only about 50% of that for
perfluorooctane sulfonate (PFOS) (ICso = 4.9 jiM) and far less than that of oleic acid (ICso = 0.01
Oral Exposure
Tissue distribution
Human. No clinical studies are available in humans on administration of a controlled dose of
PFOA or its distribution. However, samples collected in biomonitoring and epidemiological
studies provide data showing distribution of PFOA. Olsen et al. (2001c) analyzed human sera
and postmortem liver samples and found that more than 90% of the liver samples (n=30) were
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Serum concentrations reached steady-state levels within four to six weeks in all dose
groups. Steady-state concentrations of PFOA in serum were 77 ± 39, 86 ± 33, and 158 ± 100
|ig/ml after six weeks (Butenhoff et al., 2002) and 81 ± 40, 99 ± 50, and 156 ± 103 |ig/ml
(Butenhoff et al., 2004b) after six months for the 3, 10, and 20 mg/kg dose groups, respectively.
The mean serum concentration of PFOA in control monkeys was 0.134 - 0.203 |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 jig/g (Butenhoff et al.,
2004b).
Rat. Ylinen et al. (1990) administered newly weaned Wistar rats (18/sex/group) doses of 3, 10,
and 30 mg/kg-day PFOA by gavage for 28 days. At necropsy, serum was collected as well as the
brain, liver, kidney, lung, spleen, ovary, testis, and adipose tissue. The concentration of PFOA in
the serum and tissues was determined with capillary gas chromatography equipped with a flame
ionization detector (FID). A mass spectrometer was used in the selected ion monitoring mode
when the PFOA concentration was below the quantitation limit of the FID (1 jig/ml).
The concentration of PFOA in the serum and tissues following 28 days of administration
is presented in Table 3-3. PFOA was not detected in the adipose tissue. The concentrations of
PFOA in the serum and tissues were much higher in males than in females. In the males, the
levels of PFOA in the serum were generally lower in the 30 mg/kg-day dose group than in the 10
mg/kg-day dose group due to increased urinary elimination in the 30 mg/kg-day group. The
tissue levels were similar for the 10 and 30 mg/kg/day doses. In females there was a dose-related
increase in tissue levels while the serum levels were comparable for the 10 and 30 mg/kg/day
dose-groups. Among solid tissues, the liver had the highest tissue concentration followed by the
testes, spleen, lung, kidney and brain, respectively. In females the concentration in the kidneys
exceeded that in the liver for the 10 and 30 mg/kg-day doses but not at the lowest dose. Ovary
and spleen had similar concentrations followed by lower levels in the lung and brain.
TABLE 3-3.Tissue Distribution of PFOA in Wistar Rats After 28 Days of Treatment
Tissue
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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Kemper (2003) examined the distribution of PFOA in tissues of male and female
Sprague-Dawley rats following administration by gavage. Rats were administered 1, 5, and 25
mg/kg 14C-PFOA by oral gavage. Tissue concentrations were determined at the time of
maximum plasma concentration (Tmax) and at the time that plasma concentration had fallen to
one half the maximum (Tmax/2). Values for Tmax and Tmax/2 for male and female rats were
determined from pharmacokinetic experiments. In those experiments, plasma was collected over
the course of several days and PFOA concentration was analyzed. Non-compartmental
pharmacokinetic models were applied to identify Tmax and elimination half-time (Ti/2) from the
data. The Tmax/2 was calculated as the time (hr) for the maximum plasma concentration plus the
elimination half-time (hr) (Tmax + Ti/2). In some cases, elimination may occur in a rapid phase
followed by a slower elimination phase. For cases in which biphasic elimination was evident,
the rapid phase Ti/2 was used for calculation of Tmax/2.
Tissues from male rats were collected at 10.5 hours (Tmax) and 171 hours (Tmax/2) after
dosing. Tissues from female rats were collected at 1.25 hours (Tmax) and 4 hours (Tmax/2) after
dosing. The results are summarized in Table 3-4 and Table 3-5 for males and females,
respectively. Liver, kidney and blood were the primary tissues for distribution of 14C-PFOA. In
males the fraction of the dose found in liver increased between Tmax and Tmax/2, but remained
constant or decreased in other tissues. In females, the fraction of the dose present in all tissues
remained constant or decreased between Tmax and Tmax/2. Liver-to-blood concentration ratios for
14C at Tmax in males were greater than 1, and increased between Tmax and Tmax/2. Blood to kidney
concentration ratios in females were between 1 and 2 at all dose levels and remained relatively
constant between Tmax and Tmax/2. In males the blood to kidney ratios were 10 or higher at Tmax
and declined slightly at Tmax/2.
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TABLE 3-4. Distribution of PFOA in Male Sprague-Dawley Rats After Oral Exposure
Dose
Tissue
Prostate
skin3
blood3
Brain
fat3
Heart
Lungs
Spleen
Liver
Kidney
G.I. tract
G.I.
contents
Thyroid
Thymus
Testes
adrenals
muscle3
bone3
Total"
1 m
% 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
5/kg
% 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
5 m
% 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
2/kg
% 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
56.03 1±1.025
25mg/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
From Kemper, 2003
Percent of dose recovered at Tmaxand Tmax/2 in tissues
a Percent recovery scaled to whole animal assuming the following: skin= 19%, whole blood=7.4%, fat=7%, muscle=40.4%,
bone=7.3% of body weight.
b Totals are calculated from individual animal data.
Examination of the residuals from the administered PFOA in the male tissues at Tmax/2
(171 hours) indicate that the 40 to 60% percent of the dosed PFOA retained was present in the
liver, blood, skin, and muscle tissues in decreasing amounts (Table 3-4). In males, about 1% of
the label was present in the gastrointestinal tissues and contents at Tmax/2, while the value for
females was about 10%. However, the samples were collected at 1.25 and 4 hours in females
and 10.5 and 171 hours in males providing more time for absorption in the males.
In the female tissues at Tmax/2 (4 hours), approximately 30% percent of the dosed PFOA
retained was present in the liver, blood, kidney, muscle, and skin tissues in decreasing amounts
(Table 3-5). About 14% of the administered dose remained in the gastrointestinal tissues and
contents. Based on the timing of the measurements and the results, females appear to absorb and
excrete PFOA more rapidly than males.
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TABLE 3-5. Distribution of PFOA in Female Sprague-Dawley Rats After Oral Exposure
Dose
Tissue
skin3
blood3
Brain
fat3
Heart
Lungs
Spleen
Liver
Kidney
G.I. tract
G.I.
contents
Thyroid
Thymus
Ovaries
adrenals
muscle3
Uterus
bone3
Total"
1 m
% 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
?/kg
% 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
5 m
% 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
g/kg
% 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
25mg/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
From Kemper, 2003
Percent of dose recovered at Tmax and Tmax/2 in tissues
a Percent recovery scaled to whole animal assuming the following: skin= 19%, whole blood=7.4%, fat=7%, muscle=40.4%,
bone=7.3% of body weight.
b Totals are calculated from individual animal data.
Lau et al. (2006) studied PFOA's toxicokinetic properties in rats as part of a larger study.
The authors gavage dosed adult male and female Sprague-Dawley rats (n=8) with 10 mg/kg for
20 days, and sacrificed them 24 hours after the last treatment. After 20 days of treatment, male
rats had serum PFOA levels of 111 |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, or 5 days by oral gavage and determined the liver and serum levels of PFOA.
Blood was collected via cardiac puncture and PFOA concentration was determined by high-
performance liquid chromatography-electrospray tandem mass spectrometry. The mean liver
PFOA concentration was 92 ± 6, 250 ± 32, and 243 ± 23 ug/g after 1, 3, or 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.
Mouse. Lau et al. (2006) gavage dosed adult male and female CD-I mice (5-7/group) with 20
mg/kg for 7 or 17 days. The animals were sacrificed 24 hours after the last treatment. After
seven days of treatment male mice had serum PFOA levels of 181 |ig/mL and females had 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.
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As part of a physiologically based pharmacokinetic (PBPK) modelling exercise, Lou et
al. (2009) administered single doses of 1 or 10 mg/kg to groups of 3 male and 3 female CD-I
mice. The mice were sacrificed for analysis of plasma, liver, and kidney tissues after 4, 8, or 12
hours and at 1, 3, 6, 9, 13, 20, 27, 34, or 48 days after dosing. This study was repeated for a
second analysis which extended the sacrifice times to 55, 62, 70, and 80 days.
Measures of PFOA in serum were presented graphically and indicate that the order of
magnitude difference between the doses was generally reflected in an order of magnitude
difference between serum concentration for males and females across the 80 day observation
period. [The study procedures indicated that serum was collected and analyzed, but the graphic
presentation described the values as plasma values. Contact with one of the authors confirmed
that the values should have been listed as serum rather than plasma.] The peak concentrations
were 10 and 100 mg/L for the 1 and 10 mg/kg/day, respectively. Declines in serum
concentrations for females were roughly parallel reaching concentrations of about 2 mg/L and
<0.2 mg/L for the high and low doses, respectively, at the end of 80 days. Peak serum
concentrations were slightly lower in the males (~8 and 80 mg/L) than in the females, and final
serum concentrations were higher in the males (-0.5 and 8 mg/L). Liver and kidney
concentrations were also higher in males than females for each of the two doses.
Lou et al. (2009) also collected serum data for up to 28 days after administration of a 60
mg/kg dose to groups of 3 female mice. Based on the graphic presentation of the data, the 60
mg/kg dose was cleared from the serum much more rapidly than the 1 and 10 mg/kg doses. For
example, a serum concentration of about 0.4 mg/L was reached in about 28 days for the 60
mg/kg dose, 61 days for 10 mg/kg dose, and 70 days for the 1 mg/kg dose (values estimated from
the Figure 3 of the published paper). No measurements were made for liver or kidney in the high
dose animals.
In the final experimental portion of the study, Lou et al. (2009) exposed groups of 5
female CD-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 ug/mL reported by Lau et al. (2006).
Minata et al. (2010) orally administered 0, 12.5, 25, or 50 umol/kg PFOA (~0, 5.4, 10.8,
or 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 3-6. 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 38 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 the existence of a capacity-limited and PPARa independent PFOA
transport from the liver to the bile duct.
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TABLE 3-6. PFOA Concentration in Wild-type and PPARa-null Mice
PFOA Concentration (ng/mL)
Dose
umol/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
12.9 ±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
From Minata et al., 2010
Mean ± SD; ND= not detected (<0.001 ng/mL)
Tissue transporters. The liver is an important binding site for PFOA. Increased liver weight is
one of the early, low-dose manifestations of exposure. OATps and MRPs, at least one OAT and
the sodium-taurocholate cotransporting polypeptide (NTCP) have been identified at the
interphace of the liver with the portal blood and the canalicular membranes within the liver
(Kim, 2003; Kusuhara and Sugiyama, 2009; Zalr et al., 2008). OATs and MRPs have been
identified along the hepatocyte interface with systemic circulation.
The impact of PFOA on several membrane transporter systems linked to biliary transport
was studied by Maher et al. (2008) as part of a more detailed study of perfluorodecanoic acid
(PFDA). A dose of 80 mg/kg by i.p. injection (propylene glycol: water vehicle) was found to
significantly increase (p<0.05) the expression of MRPS and MRP4 in the livers of C57BL/6 mice
two days after treatment as reflected in quantification of their DNA transcripts. MRPS and
MRP4 are believed to protect the liver from accumulation of bile acids, bilirubin, and potentially
toxic exogenous substances by promoting their excretion in bile. Conversely, there were
significant decreases (p<0.05) in the protein levels for OATplal, OATpla4, and OATplb2
(Cheng and Klaassen, 2008) as determined by Western Blot analysis and mRNA measurements.
There was no significant impact on NTCP protein or the serum levels of bile acids. The OATps
are transporters responsible for the uptake of bile acids and other hydrophobic substances such as
steroid conjugates, ecosinoids, and thyroid hormones into the liver.
These studies, by the same laboratory, were carried out at high, single dose exposures
which limit their value in extrapolating to low and repeat dose scenarios. The results suggest a
decrease in the uptake of favored substrates into the liver and an increase in removal of favored
substrates from the liver via bile. Based on the results with the more extensive evaluation of
PFDA including mouse strains null for several receptors (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 four to eight week old
Sprague-Dawley rats (10/sex/age) a single dose of 10 mg/kg-day PFOA by oral gavage. Blood
samples were collected 24 hours after dosing and the plasma concentration of PFOA was
measured by HPLC-MS. In the 4 week old rats the concentration of plasma PFOA was
approximately 2.7 times higher in males than in the females (Table 3-7). In the 5 and 6 week old
female rats, the plasma PFOA concentrations were about 2 fold lower than in the 4-week old
rats. However, in the 5 week old males, the concentration of plasma PFOA was about 5-fold
higher than in the 4 week group suggesting a developmental change in excretion rate. Plasma
concentrations did not differ appreciably among 5-, 6-, 7- and 8-week old rats within each sex
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but did differ between sexes. In fact, PFOA plasma concentrations were 35-65-fold higher in
males than in females at every age except at 4 weeks. Thus, it appears that maturation of the
factor(s) responsible for the gender difference in elimination occurs between the ages of 4 and 5
weeks in the rat.
TABLE 3-7. Plasma PFOA Concentrations (jig/ml) in Postweaning Sprague-Dawley Rats
Age (weeks)
4
5
6
7
8
Males
7.32±1.01a
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
From Han, 2003
a mean ± SD; samples from 10 animals/sex/group
Hinderliter (2004, 2006a) continued the investigation of the relationship between age and
plasma PFOA in male and female Sprague-Dawley rats. Immature rats 3, 4, or 5 weeks of age
were administered PFOA via oral gavage as a single dose of 10 or 30 mg/kg. Rats were not
fasted prior to dosing. Two hours after dosing, 5 rats/sex/ per age group and dose group were
sacrificed and blood samples were collected. The remaining 5 rats/sex/per age and dose group
were placed in metabolism cages for 24-hour urine collection. These rats were sacrificed at 24
hours and blood samples were collected.
In the male rats, plasma PFOA concentrations for either the 10 or 30 mg/kg dosage
groups did not differ significantly by sample time (at 2 and 24 hours) or by animal age (3, 4 or 5
weeks), except at 2 hours for the five week age group (p<0.01) which showed the lowest PFOA
level (Table 3-8). PFOA plasma concentrations following a 30 mg/kg dose were 2-3 times higher
than those following a 10 mg/kg dose. These data do not demonstrate a difference between the 5-
week old rats and the younger 3- and 4-week age groups at 24 hours after dosing and thus do not
support the observations from the Han (2003) study.
TABLE 3-8. Plasma PFOA Concentrations in Male Rats
Age
(weeks)
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
From Hinderliter 2004
* Statistically significantly different by sample time and animal age (p<0.01).
In the female rats, plasma PFOA concentrations were significantly lower in the 5-week
age group than in the 3- or 4-week age groups at the 24-hour time period for both doses and at
the 30 mg/kg dose group at 2-hours (Table 3-9). Plasma PFOA concentrations following a 30
mg/kg dose were approximately 1.5 to 4 times higher than those observed following a 10 mg/kg
dose. At 24 hours post dose, plasma PFOA levels in the female rats was significantly lower than
the plasma PFOA levels in male rats, especially at 5 weeks of age. The data for the 5-week old
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female rats compared to the 3- and 4-week age groups at 24 hours are consistent with the Han
(2003) data in that they demonstrate a decline in plasma levels compared to their earlier
measurements. Thus, the developmental change is one that appears to be unique to the female
rat.
TABLE 3-9. Plasma PFOA Concentrations in Female Rats
Age
(weeks)
3
4
5
3
4
5
Dose
(mg/kg)
10
10
10
30
30
30
Plasma PFOA (M-g/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
18.98 b
1.36 a'b
51.43 b
28.01 b
3.42a'b
SD
3.83
7.01
0.87
13.61
9.90
1.95
From Hinder-liter, 2004
a Statistically significantly different from the 3 and 4 week values (p<0.01).
b Statistically significantly different from 2 hour values (p<0.01).
In a supplemental study to determine the effect of fasting (Hinderliter, 2004, 2006a), 4
week old rats, 4 rats/sex, were administered 10 mg/kg PFOA via oral gavage. Animals (2/sex)
were fasted overnight for 12 hours before dosing with PFOA. All the rats were sacrificed at 24
hours post dosing and blood was collected for analysis of PFOA in plasma. Plasma PFOA
concentrations in male rats were 64.95 and 30.00 (ig/ml for the fasted and non-fasted animals,
respectively. Plasma PFOA concentrations in the female rats were 68.16 and 26.54 |j,g/ml for the
fasted and non-fasted animals, respectively. Given the consistency in the 4-week old rat plasma
PFOA concentrations, the authors concluded that age dependent changes in female PFOA
elimination are observable between 3 and 5 weeks of age.
Distribution during pregnancy and lactation
Rat. An oral two-generation reproductive toxicity study of PFOA in rats was conducted (York,
2002; Butenhoff et al., 2004a). Five groups of rats (30 sex/group) were administered PFOA by
gavage at doses of 0, 1, 3, 10, and 30 mg/kg-day. At scheduled sacrifice, after completion of the
cohabitation period in FO male rats and on lactation day (LD) 22 in FO female rats, blood
samples (3/sex/group-control; 10/sex/group-treated) were collected from animals dosed with 0,
10, or 30 mg/kg for analysis of PFOA. Serum analysis for the 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 the limits of quantitation (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 or 30
mg/kg-day PFOA during days 4-10, 4-15, or 4-21 of gestation, or from gestation day 4 to
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lactation day 21. Maternal blood samples were collected at 2 hours ± 30 minutes post-dose on a
daily basis. Clinical observations and body weights were recorded daily. Five animals per dose
group were sacrificed at specific time periods to harvest the conceptus and/or placenta and
amniotic fluid. On gestation day (GD) 10 only embryos were recovered and on GD 15 and 21
the placentas, amniotic fluid, and embryos/fetuses were collected.
The remaining 5 rats per group were allowed to deliver their pups. On lactation days 0, 3,
7, 14, and 21, the pups were counted, weighed (sexes separate), and examined for abnormal
appearance and behavior. Randomly selected pups were sacrificed and blood samples were
collected. On lactation days 3, 7, 14, and 21, the dams were anesthetized and milk and blood
samples were collected; dams were removed from their litters 1-2 hours prior to collection.
Plasma, milk, amniotic fluid extract, and tissue homogenate (placenta, embryo, and fetus)
supernatants were analyzed for PFOA concentrations by HPLC-MS. Maternal PFOA plasma
levels during gestation and lactation are presented in Table 3-10. Maternal plasma levels at 2 hrs
post-dosing (approximately the time of peak blood levels following a gavage dose) were fairly
similar during the course of the study with a mean level of 11.2, 26.8, and 66.6 |ig/ml in the 3,
10, and 30 mg/kg-day groups, respectively; PFOA levels in the control group were below the
limit of quantitation (0.05 jig/ml).
TABLE 3-10. Maternal Plasma PFOA Levels (jig/ml) 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
From Hinderliter et al., 2005 and Mylchreest, 2003
Mean ± SD; samples were from 5 dams/group/time point and were collected 2 hrs post-dosing
PFOA levels in the placenta, amniotic fluid, and embryo/fetus are presented in Table 3-
11. The levels of PFOA in the placenta on gestation day 21 were approximately twice the levels
observed on gestation day 15, and the levels of PFOA in the amniotic fluid were approximately
four times higher on day 21 than on day 15. The concentration of PFOA in the embryo/fetus was
highest in the day 10 embryo and lowest in the day 15 embryo; PFOA levels in the day 21 fetus
were intermediate.
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TABLE 3-11. Placenta, Amniotic Fluid, and Embryo/Fetus PFOA Concentrations (fig/ml)
Exposure Period
GD 4 - GD 10
GD 4 - GD 15
GD4-GD21
Tissue
GD 10 - embryo
GD 15 - placenta
- amniotic fluid
- embryo
GD21 -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
From Hinderliter et al., 2005 and Mylchreest, 2003
Mean ± SD; samples were pooled by litter and were collected 2 hrs post-dosing
The concentrations of PFOA in the plasma of the day 21 fetus were approximately half
the levels observed in the maternal plasma (Table 3-10). The values were about twice as high in
the dams as in the pups with mean values of 14.04, 34.20 and 76.36 |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 in the,
respectively for the pups. Pup plasma levels decreased between birth and lactation day 7 (Table
3-12), and were thereafter similar to the levels observed in the milk (see Table 3-13 below). The
pups were not separated by gender.
TABLE 3-12. Fetus/Pup PFOA Concentration (fig/ml) During Gestation and Lactation
Exposure Period
GD4-GD21
GD 4 - LD 3
GD 4 - LD 7
GD 4 - LD 14
GD4-LD21
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
From Hinderliter et al., 2005 and Mylchreest, 2003
Mean ± SD; samples were pooled by litter and were collected 2 hrs post-dosing
The concentration of PFOA in the milk was also fairly similar throughout lactation and
was approximately l/10th of the PFOA levels in the maternal plasma (see Table 3-11 above); the
mean values for maternal milk were 1.1, 2.8, and 6.2 |ig/ml in the 3, 10, and 30 mg/kg-day
groups, respectively (Table 3-13).
TABLE 3-13. PFOA Levels (jig/ml) in Sprague-Dawley Maternal Milk During Lactation
Exposure Period
GD 4 - LD 3
GD 4 - LD 7
GD 4 - LD 14
GD4-LD21
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
From Hinderliter et al., 2005 and Mylchreest, 2003
Mean ± SD; samples were from 5 dams/group/time point and were collected 2 hrs post-dosing
Mouse. Fenton et al. (2009) orally dosed pregnant CD-I mice (n=25/group) with 0, 0.1, 1, or 5
mg PFOA/kg on GDI7. On GDIS, 5 dams/group were sacrificed and trunk blood, urine,
amniotic fluid, and the 4th and 5th mammary gland were collected. One fetus/dam was
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euthanized and retained for whole pup analysis. The remaining dams were allowed to litter.
Biological samples as described above excluding amniotic fluid were also collected on postnatal
days 1, 4, 8, and 18. As before, at each time-point, a single pup was euthanized and retained for
whole pup analysis. Blood from the remaining pups was collected and pooled. Milk was
collected from dams on postnatal days 2, 8, 11, and 18 following a 2 hour separation of the pups
from the dam.
The concentration of PFOA in dam serum was approximately twice that detected in
amniotic fluid (Table 3-14). Compared to the amniotic fluid, the concentration of PFOA in the
fetuses was increased by 2.3-, 3.1-, and 2.7-fold at 0.1, 1, and 5 mg/kg, respectively. The highest
concentration of PFOA was detected in the serum of nursing dams. In the dams, the
concentration of PFOA in the serum exhibited a U-shaped response curve; the lowest serum
concentration was observed at the time of peak lactation. Dam mammary tissue and milk PFOA
concentrations showed a U-shaped response that mirrored to that found in the dam's serum. The
concentration of PFOA in pup's serum was significantly higher than PFOA concentration in
dam's serum, and appeared to decrease as the time for weaning approached. When pup PFOA
concentration was calculated with consideration for pup body weight gain, PFOA body burden
increased through the peak of lactation, and began to decrease by PND18 showing an inverse U-
shaped response curve.
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TABLE 3-14. PFOA Levels (ng/ml) During Gestation and Lactation in Selected Fluids and Tissues
Tissue
Dam Serum
Amniotic Fluid
Dam Urine
Mammary Gland
Milk
Whole Pup
Pup Serum
Day
GDIS
PND1
PND4
PND8
PND18
GDIS
GDIS
PND1
PND4
PND8
PND18
GDIS
PND1
PND4
PND8
PND18
PND2
PND8
PND11
PND18
GDIS
PND1
PND4
PND8
PND18
PND1
PND4
PND8
PND18
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
146 1.8 ±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
3118.5 ±424
1391.5 ± 118
16,286.4 ± 1372
11,925.2 ±1077
9215.8 ±594
5894.3 ± 743
From Fenton et al., 2009
Pregnant C57BL/6/Bkl mice were fed diets containing 0.3 mg PFOA/kg/day from GDI
through the end of pregnancy. At birth, the PFOA concentrations in the offspring were 0.7 ±0.1
ug/g in the brain and 16.3 ± 4.1 ug/g in the liver (Onishchenko et al., 2011).
Macon et al. (2011) gavage dosed CD-I mice with 0, 0.3, 1.0, or 3.0 mg PFOA/kg from
GDI to GDI? or with 0, 0.01, 0.1, or 1.0 mg PFOA/kg from GD10 to GD17. In the full
gestation experiment (GD1-17), offspring were were sacrificed on postnatal day (PND) 7, 14, 21,
28, 42, 63, and 84, and in the half gestation experiment (GDI0-17), female offspring were
sacrificed on PND1, 4, 7, 14 and 21. Serum, liver, and brain from the offspring were analyzed
for PFOA by HPLC-MS/MS.
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At the lowest dose, PFOA concentration in the serum peaked around PND7, but the two
higher doses peaked around PND 14 (Table 3-15). Calculated blood burdens which take into
account the increasing blood volumes and body weights for females showed an inverted U-
shaped curve peaking at PND14 for all doses. In the liver, PFOA concentration decreased over
time with the highest concentration observed at PND7. Lower concentrations of PFOA were
detected in the brain of the offspring on PND 7 and 14.
TABLE 3-15. Female Offspring PFOA Levels (ng/ml) After GD1-17 Exposure
Tissue
Female Serum
Female Liver
Female Brain
Day
PND7
PND 14
PND21
PND28
PND42
PND63
PND84
PND7
PND 14
PND21
PND28
PND42
PND63
PND84
PND7
PND 14
PND21
PND28
PND42
PND63
PND84
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 3-16. Female Offspring Serum PFOA Levels (ng/ml) After GD10-17 Exposure
Tissue
Female Serum
Female 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
From Macon et al., 2011
- = not measured, blood burden determined by (body weight x (58.5/1000) x serum x 0.55)
White et al. (2011) measured serum PFOA concentrations in three generation of CD-I mice
(Table 3-17). Pregnant mice (FO, n=10-12 dams/group) were gavage dosed with 0, 1, or 5 mg
PFOA/kg from GD1-17. A separate group of pregnant mice (n=7-10 dams/group) were gavage
dosed with either 0 or 1 mg PFOA/kg from GDI-17 and received drinking water containing 5
ppb PFOA beginning on GD7 and continuing until the end of the study for their offspring, except
during breeding and early gestation, to simulate a chronic low-dose exposure. An increase in
serum PFOA concentration was observed in the control + 5 ppb PFOA groups in the Fl and F2
generations and in the 1 mg/kg + 5ppb PFOA group of the F2 generation. A decrease was
observed for the remaining groups.
TABLE 3-17. Serum PFOA Levels (ng/ml) Over Three Generations
Dams at weaning
Offspring
Generation/
Day
FO/ PND22
F1/-PND91
F1/PND22
F1/PND42
F1/PND63
F2/PND22
F2/PND42
F2/PND63
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
From 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 Crl:CD (SD)IGS BR rats were
gavage dosed with 25 mg/kg [14C]PFOA and sacrificed 2 hours after dosing. Blood was
collected and the liver and kidneys were removed. Five subcellular fractions (nuclei and cell
debris, lysosome and mitochondria, microsome, light microsome and ribosome, and membrane-
free cytosol) were obtained by differential centrifugation. The radioactivity per gram of each
fraction and the total radioactivity were measured.
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In the male liver, the highest proportion of total reactive residues (TRR) of PFOA was
located in the nuclei and cell debris (40%). The TRR for the other subcellular fractions were as
follows: membrane-free cytosol 26% TRR, lysosome and mitochondria -14% TRR, microsome
-16% TRR. The level of PFOA in the light microsome and ribosome was -1% TRR. In the
female liver, the highest proportion of PFOA was found in the membrane-free cytosol, 48%
TRR. The TRR were nuclei and cell debris -31% TRR, lysosome and mitochondria -12% TRR,
and microsome -8% TRR. As observed in the males, the level of PFOA in the light microsome
and ribosome was -1% TRR.
In the male kidney, the level of PFOA was 79% TRR in the membrane-free cytosol, 15%
TRR in the nuclei and cell debris, and 4% TRR in the lysosome and mitochondria/microsome/
light microsome and ribosome (combined). In the female kidney, the level of PFOA was 71%
TRR in the cytosol, 21% TRR in the nuclei and cell debris, and 8% TRR in the lysosome and
mitochondria/ microsome/light microsome and ribosome (combined). Further examination
showed that in both sexes 98% of PFOA in the plasma was protein bound and the protein bound
fraction of PFOA in the liver cytosol was 56% TRR. In the kidney, the protein bound fraction of
PFOA in males was 42% TRR and 17% TRR in the females.
Based on the results, the authors concluded that subcellular distribution of PFOA in the
rat liver was sex-dependent because the proportion of PFOA in the liver cytosol of female rats
was almost twice that in the male. They hypothesized that the female may have a greater amount
than males of an unknown liver cytosolic binding protein with an affinity for perfluorinated
acids. They also hypothesized that the unknown protein or protein complex might normally aid
in transport of fatty acids from the liver. In the kidney, the subcellular distribution did not show
the gender difference seen with the liver, however the protein bound fraction for the males (42%)
was about twice that for thefemales (17%).
Inhalation Exposure
In a repeated exposure study, Hinderliter (2003, 2006b) exposed 6-8 week old male and
female rats (5/sex/group) to 0, 1, 10, or 25 mg/m3 aerosol concentrations of PFOA for 6
hours/day, 5 days/week for 3 weeks. Blood was collected immediately before and after the daily
exposure period three days per week. The aerosols had mass median aerodynamic diameters
(MMAD) of 1.3 - 1.9 |im with geometric standard deviations of 1.5 - 2.1. PFOA plasma
concentrations were proportional to the inhalation exposure concentrations, and repeated
exposures produced little plasma carryover in females, but significant day-to-day carryover in
males. Male rats reached steady state plasma levels of 8, 21, and 36 |ig/ml for the 1, 10 and 25
mg/m3 groups, respectively by three weeks. In females, the post-exposure plasma levels were 1,
2, and 4 |ig/ml for the 1,10, and 25 mg/m3 groups, respectively. When measured immediately
before the daily exposure, plasma levels had returned to baseline in females demonstrating
clearance within 24 hours of each daily dose.
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Dermal Exposure
No data were identified on tissue distribution following dermal exposures.
3.3 Metabolism
Several studies have examined metabolism of PFOA. However, no studies show clear
evidence of metabolism. Ophaug and Singer (1980) found no change in fluoride ion level in the
serum or urine following oral administration of PFOA to female Holtzman rats. Ylinen et al.
(1989) found no evidence of phase II metabolism of PFOA following a single intraperitoneal
PFOA dose (50 mg/kg) in male and female Wistar rats. The free anionic and possible
conjugated forms of PFOA in the urine were separated using BondElut tubes. The tubes contain
NH2 which is a weaker anion exchange sorbent and a good choice for retaining strong anions.
The samples were aspirated through the tube, washed with water, and eluted with sodium
bicarbonate/ carbonate-buffer. The aspirate and eluate from the separation method were
analyzed by gas chromatography. PFOA was not detected in the aspirate, but was retained with
the cationic animo phase found in the eluate. This also occurred in control blanks spiked with
PFOA. The authors concluded that because the PFOA anion was completely bound to the weak
cationic amino phase in both the spiked controls and urine samples, PFOA is secreted in the
urine as such and has a neglible rate of phase II metabolism (Ylinen et al., 1989). The potential
for reduction of the carboxyl group does not appear to have been evaluated but is not likely to
occur.
3.4 Excretion
Excretion data are available for oral exposure in humans and laboratory animals.
Several studies have investigated the elimination of PFOA in humans, cynomolgus monkeys, and
rats. In human females, elimination pathways include pregnancy (cord blood) and lactation
(breastmilk) (Apelberg et al., 2007; Tao et al., 2008; Thomsen et al., 2010; Volkel et al., 2008;
von Ehrenstein et al., 2009).
Urinary elimination of PFOA was reported in a case history of of a single human male
with high serum levels of perfluorinated chemicals (Genuis et al., 2010). Treatment with
cholestyramine, a bile acid sequestrant for 20 weeks (4g/day, three times a day) lowered his
serum PFOA concentration from 5.9 ng/g serum to 4.1 ng/g serum. More dramatic decreases
were observed with serum PFOS (23-14.4 ng/g serum) and perfluorohexanesulfonic acid
(PFHxS) (58-46.8 ng/g serum) which were present at higher levels in the serum. This
observation that suggests excretion with bile and possible enterohepatic resorption via intestinal
transporters limiting the loss of absorbed PFOA via feces in the absence of a binding agent such
as cholestyramine.
Elimination half-lives differ among species. There are also significant gender differences
in the elimination of PFOA in some laboratory animal species. Information from humans does
not at present provide sufficient data to determin the magnitude of interindividual and gender
differences in excretory half lives. The organic acid transport system appears to play an
important role in renal excretion of PFOA and possibly biliary elimination.
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Oral Exposure
Monkey. Butenhoff et al. (2004b) studied the fate of PFOA in cynomolgus monkeys in a six-
month oral exposure study. Groups of four to six male monkeys each were administered PFOA
daily via oral capsule at dose rates of 0, 3, 10, or 30/20 mg/kg for six months. Two monkeys
exposed to 10 mg/kg and three monkeys exposed to 20 mg/kg were monitored for 21 weeks
(recovery period) following dosing. Urine and fecal samples were collected at two-week
intervals and were analyzed for PFOA concentrations.
Urine PFOA concentrations over the duration of the study were 53 ± 25, 166 ± 83, and 181 ±
100 |ig/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, 2006a) investigated the relationship between age and
urine PFOA concentrations in male and female Sprague-Dawley rats. Immature rats 3, 4, or 5
weeks of age were administered PFOA via oral gavage as a single dose of 10 or 30 mg/kg. Two
hours after dosing, 5 rats/sex/age group and dose group were sacrificed and blood samples were
collected (See section 3.2). The remaining 5 rats/sex/ age and dose group were placed in
metabolism cages for 24-hour urine collection. Urinary output (volume) was not quantified or
standardized for creatinine levels.
Urine PFOA concentrations differed significantly with age, dose and sex (p<0.01, Table 3-
18). Female rats had higher urine PFOA concentrations than males, and the female urine PFOA
concentrations increased with age. In male rats, 24-hour urine PFOA concentrations decreased
with age up to five weeks. In both sexes, urine PFOA was higher (2.5 to 6.5 times) at the 30
mg/kg dose as compared to the 10 mg/kg dose.
There was a difference in urinary excretion between the 3 week and 4/5 week male rats with
the older rats excreting -50% less PFOA in the urine compared to the younger rats at 10 mg/kg
and 30 mg/kg. If the data from urine are integrated with the plasma data in the same study
(Table 3-8), the male plasma levels increased from the 3-week value and were relatively stable
for weeks 4 and 5. In the females, urine excretion increased gradually with age (Table 3-18) and
plasma concentrations decreased (Table 3-9).
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TABLE 3-18. Urine PFOA Concentrations in Male and Female Rats
Age
(weeks)
3
4
5
3
4
5
Dose
(mg/kg)
10
10
10
30
30
30
Urine PFOA (|J,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
From Hinderliter, 2004
Hundley et al. (2006) examined excretion of PFOA in one male and one female CD rat
(sexually mature). Each was given a single dose of 10 mg/kg 14C-PFOA and housed in a
metabolism cage. Urine and feces were collected at 12, 24, 48, 72, 96, and 120 hours post-dose.
The female rat excreted more PFOA over the 120 hour collection period than the male rat. In the
male rat, 25.6% and 9.2% 14C-PFOA were excreted in the urine and feces, respectively. In the
female rat, 73.9% and 27.8% 14C-PFOA were excreted in the urine and feces, respectively. The
female rat excreted almost all of the PFOA by 48 hours compared with only 19% of the dose
excreted by the male rat over the same amount of time. The cumulative percent of the dose
excreted is shown in Table 3-19.
TABLE 3-19. Cumulative Percent "C-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
From Hundley et al., 2006
Adult male Sprague-Dawley rats (n=7) were given a single gavage dose of 0.5 mg
PFOA/kg and monitored for 38 days (Benskin et al, 2009). Over the course of the study, the rats
were held in metabolic cages and urine and feces were collected. The mean blood PFOA
concentration was 1.1 |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/group) to PFOA
(96% a.i.) at 0, 5, and 20 mg/kg/day once daily by gavage for 28 days. Urine and fecal samples
were collected through use of metabolism cages at 24 hour intervals immediately following
dosing on days 1, 2, 5, 7, 10, 14, 18, 21, 24, and 28 of the study. Daily urine volume and fecal
weight were comparable across all groups throughout the study. As measured by excretion 24-
hours after the first dose, 17.9% of the applied dose was excreted in the urine of the low-dose
group and 22% for the high dose group. The percent of the absorbed dose was 92.8% and 92.3%
for the low and high doses, respectively when the fecal excretion over the 24-hours following
dosing was estimated to be unabsorbed material. During week 1, a sharp increase in urinary and
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fecal excretion expressed as percent of dose/day was observed in rats of both groups. The
excretion rate leveled off at about 50% for the low-dose animals for the remainder of the 28
days. In the case of the high-dose animals, the urinary excretion remained level at about 80% for
the second and third weeks and then increased sharply to about 140% at 28 days. The fecal
excretion rates were 7.2% and 7.7% for rats in the 5 and 20 mg/kg groups, respectively during
the first 24 hours post dosing and continued an upward trend throughout the 28 days with the
terminal percent/day about 25% for the low dose group and 40% for the high dose group.
Hormonal changes during pregnancy do not appear to cause a change in the rate of
elimination of 14C after oral administration of a single dose of 14C-PFOA (Gibson and Johnson,
1983). At 8 or 9 days after conception, four pregnant CD rats and two nonpregnant female CD
rats were given a mean dose of 15 mg/kg 14C-PFOA. Individual urine samples were collected at
14,
12, 24, 36, and 48 hours post dose and analyzed for C content. Essentially all of the
eliminated via the urine within 24 hours for both groups of rats
14,
C was
Other Species. Hundley et al. (2006) examined excretion of PFOA in CD mice, BIO-15.16
hamsters, and New Zealand White rabbits. One male and female of each species was given a
single dose of 10 mg/kg 14C-PFOA and housed in metabolism cages. Urine and feces were
collected at 12, 24, 48, 72, 96, and 120 hours post-dose. Additional samples were collected from
rabbits at 144 and 168 hours post-dose.
Over 120 hours, the male mouse excreted 3.4% and 8.3% 14C-PFOA, and the female
mouse excreted 6.7% and 5.7% 14C-PFOA in urine and feces, respectively. The mice were
similar in amount excreted. The male hamster excreted 90.3% and 8.2% 14C-PFOA, and the
female hamster excreted 45.3% and 9.3%14C-PFOA in urine and feces, respectively. The male
14,
hamster excreted a greater amount of C-PFOA than the female hamster. Over 84% of C-
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, and the female rabbit excreted 87.9% and 4.6% 14C-
PFOA in the urine and feces, respectively. Both rabbits excreted most of the dose by 24 hours.
The cumulative percentage of 14C-PFOA excreted is shown in Table 3-20.
TABLE 3-20. Cumulative Percent "C-PFOA Excreted in Urine and Feces
Species
Mouse
Hamster
Rabbit
Sex
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
From Hundley et al., 2006
Inhalation Exposure
Although no data were identified on urine or fecal excretion of PFOA following
inhalation exposures, the Hinderliter (2003) study provides evidence of clearance following
single and repeated inhalation exposures in Sprague-Dawley rats. Plasma PFOA concentrations
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following a single exposure to 1, 10, or 25 mg/m3 PFOA declined 1 hour after exposure in
females and 6 hours after exposure in males. In females, the elimination of PFOA was rapid at
all exposure levels, and by 12 hours after exposure the plasma levels had dropped below the
analytical limit of quantitation (0.1 jig/ml). In males, the plasma elimination was much slower
and at 24 hours after exposure, the plasma concentrations were approximately 90% of the peak
concentrations at all exposure levels. In the repeated exposure study, male and female rats were
exposed to the same concentrations for 6 hr/day, 5 days/week for 3 weeks. Steady state plasma
levels were reached in males by 3 weeks, but plasma PFOA levels in females returned to
baseline with 24 hours of each dose. The data are illustrative of distinct toxicokinetic differences
between male and female rats in their response to PFOA exposure (Hinderliter, 2003).
Dermal Exposure
No data were identified on excretion following dermal exposures. Little to no fecal
excretion is anticipated for the dermal route of exposure.
3.4.1 Mechanistic Studies of Renal Excretion
Several studies have been conducted to elucidate the cause of the gender difference in the
elimination of PFOA for rats. Many of the studies have focused on the role of transporters in the
kidney tubules. Most studies have examined the OAT transporters located in the proximal
portion of the descending tubule. They are located in other tissues as well and were discussed
earlier for their role in absorption and distribution. In the kidney they are responsible for
delivery of organic anions including a large number of medications from the serum into the
kidney tubule for excretion as well as reabsorption of anions from the glomerular filtrate. The
transporters are particularly important in excretion of PFOA because its binding to surfaces of
serum proteins (particularly albumin) make much of it unavailable for removal during
glomerular filtration. Other transporter families believed to be involved in renal excretion are
the OATps and the MRPs. However, they have not been evaluated as extensively as the OATs
for their role in the kidney.
OATs are located on both the basolateral (serum interphase) and apical surfaces of the
brush boarder of the proximal tubule inner surface. At the basolateral surface the OATs
transport the perfluorooctanoate anion from the serum to the tubular cells (Anzai et al, 2006;
Cheng and Klaassen 2008; Nakagawa et al. 2007, 2009; Klaassen and Lu 2008). OATs 1, 2, and
3 are located on the basolateral membrane surface. The OAT 4 and 5 transporters are located on
the apical surface of the tubular cells where they function to reabsorb the PFOA anions from the
glomerular filtrate. Figure 3.2 diagrams the flow of organic anions such as PFOA anion from
serum to the glomerular filtrate for excretion and resorption of organic acids from the glomerular
filtrate with transport back to serum. OATs can function for uptake into the cell across both the
basolateral and apical surfaces.
There are several MRP transporters that also appear to function in the kidney and move
organic anions in and out of cells at both the basolateral (e.g. MRP2/4) and apical (e.g. MRP1)
surfaces (Launay-Vacher et al., 2006; Klaassen and Lu, 2008; Cheng and Klaassen, 2009;
Kusuhara and Sugiyama, 2009) and one or more OATps on each surface (Cheng and Klaassen,
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2009; Kusuhara and Sugiyama, 2009; Yang et al., 2009). Bidirectional movement of PFOA
across both the basolateral and apical surfaces can occur as driven by concentration gradients
and/or active transport. There are far more data on the PFOA and OATs in the kidneys than on
OATps and MRPs. Abreviations for individual transporters on the basolateral and apical
surfaces differ across publications.
Proximal Tubule Renal Tubular Cell Capillary
FIGURE 3-2. Renal Organic Acid Transporters
From Anzai et al., 2006; Luanay-Vacher et al., 2006; Cheng and Klaassen, 2006; Klaassen and Lu, 2007; Cheng et al., 2008; Zair
et al., 2008; Kusuhara and Sugiyama, 2009
Knowledge about specific OAT, OATp and MRP transporters in the kidneys is rapidly
evolving. A low membrane density or blockage of basolateral OATs will decrease PFOA
excretion while low membrane densities or blockage of apical OATs will increase excretion
because they decrease resorption of anions from the glomerular filtrate.
The earliest studies of the impact of gender on urinary excretion were conducted by
Hanhijarvi et al (1982) using probenecid, an inhibitor of renal excretion of organic acids on
PFOA excretion in male and female Holtzman rats. The female rats that had not received the
probenecid excreted 76% of the administered dose of PFOA, while males excreted only 7.8% of
the administered dose over a 7-hr period. Probenecid administration modified the cumulative
excretion curve for males only slightly. However, in females probenecid markedly reduced
PFOA elimination to 11.8%. The authors concluded that the female rat possesses an active
secretory mechanism which rapidly eliminates PFOA from the body while male rats do not.
Kudo et al. (2002) examined the role of sex hormones and organic anion transporters on
the renal clearance (CLR) of PFOA. Renal mRNA levels of specific organic anion transporters in
castrated male and ovariectomized female Wistar rats were also determined. Castration of male
rats caused a 14-fold increase in CLR of PFOA. The elevated PFOA CLR in castrated males was
reduced by treating them with testosterone. Treatment of male rats with estradiol increased the
CLR of PFOA. In female rats, ovariectomy caused a significant increase in CLR of PFOA (2-fold
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increase), but the administration of estradiol to ovariectomized female rats returned CLR of
PFOA to normal values. Treatments of female rats with testosterone reduced the CLR of PFOA.
Treatment with probenecid, a known inhibitor of OATs 1-6 and 8, markedly reduced the
CLR of PFOA in male rats, castrated male rats, and female rats (Kudo et al., 2002).
Accordingly, the male sex hormones appear to decrease the presence of OATs in the renal
basolateral membranes while the female sex hormones appear to increase the receptors.
To identify the transporter molecules that are responsible for PFOA transport in the rat
kidney, renal mRNA levels of specific organic anion transporters were determined in male and
female rats under various hormonal states and compared with the CLR of PFOA (Kudo et al.
2002). The level of OAT2 mRNA in male rats was only 13% that in female rats. Castration or
estradiol treatment increased the level of OAT2 mRNA whereas treatment of castrated male rats
with testosterone reduced it. Ovariectomy of female rats significantly increased the level of
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 the organic anion-transporting polypeptide (OATpl) was mentioned as a
possible factor in male rat retention of PFOA. OAT2 and OATS are located on the basolateral
cell surface. OATpl is located on the apical surface of the renal tubule cells.
Cheng et al. (2006) examined whether sex hormones influenced gender specific OATp
expression in the kidneys of adult male and female C57BL/6 mice. Gonadectomized mice were
used for these studies in conjunction with hormone replacement measures (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 that in the females.
Treatment with E2 almost abolished the expression of OATplal in the kidneys and caused no
significant change in OATpSal. In the intact control animals there was almost no expression of
OATplal in the kidneys of females and a significantly lower (p<0.05) expression of OATpSal.
In the gonadectomized control animals, there was litle or no expression of OATplal in either
sex, and expression of OATpSal was equivalent in both sexes.
Nakagawa et al. (2007) investigated the role of OATs in the renal excretion of PFOA
using in vitro methods. HEK293 transformed cells, derived from human embryonic kidney,
were transfected with human or rat OAT1, OAT2, or 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 only served as control cells. The transfected HEK293
cells were incubated for one minute with or without 0, 10, or 100 jiM [14C]PFOA and/or varying
concentrations of favored OAT substrates as follows: 5 jiM [14C]para-aminohippuric acid
(OAT1), 20 nM [14C]estrone sulfate (OATS), or 10 nM [14C]prostaglandin F2a (OAT2) to
determine inhibitory effects of PFOA.
PFOA significantly inhibited para-aminohippuric acid and estrone sulfate uptake
mediated by OAT1 and OATS, respectively. At 10 |iM PFOA, uptake of 5 |iM [14C]para-
aminohippuric acid was inhibited by 75-85% compared to control, and 100 jiM PFOA inhibited
it to 35-45% of control. Estrone sulfate uptake by human OATS was inhibited by 65% compared
to control at 10 |iM PFOA and 40% of control at 100 |iM PFOA. Estrone sulfate uptake by rat
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OATS was inhibited to 15% of control in the presence of 10 |iM PFOA and by 100% at 100 |iM
PFOA. Prostaglandin F2a uptake by OAT2 was inhibited moderately by PFOA, 75-85% of
control at 10 |iM PFOA and 65% of control at 100 |iM PFOA.
In the second part of their study Nakagawa et al. (2007) incubated HEK293 and S2
transfected cells with 10 jiM [14C]PFOA for 1 minute to determine uptake. Time dependent
uptake of 5 jiM [14C]PFOA from 0-30 minutes was conducted in the HEK293 cells transfected
with human or rat OAT1 OAT2 or OAT3. Experiments were conducted in triplicate. Uptake of
PFOA was stimulated (p<0.001) in cells transfected with human or rat OAT1 or OAT3, while no
uptake was stimulated in cells transfected with OAT2 in either cell line. In the time dependent
experiments, uptake by human or rat OAT1 or OAT3 increased linearly up to 2 minutes and
reached a plateau in about 15 minutes. Kinetic evaluations resulted in Km (substrate
concentration at which the initial reaction rate is half maximal) values of 48.0, 51.0, 49.1, and
80.2 jiM for human OAT1, rat OAT1, human OAT3, and rat OAT3, respectively. The authors
showed that both human and rat OAT1 and OAT3 transport PFOA in the kidney while human
and rat OAT2 does not.
Yang et al. (2009) investigated the role of organic anion transporter polypeptide lal
(OATplal) in the renal elimination of PFOA. The polypeptide is located on the apical side of
proximal tubule cells and may be the mechanism for renal reabsorption of PFOA in rats. The
level of mRNA of 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 (E3S). 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 cDNA. The transfected
CHO cells were incubated with 4 jiM [14C]PFOA for up to 10 minutes or with 0-1000 jiM
[14C]PFOA for 2 minutes 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 jiM
[14C]PFOA for 2 minutes in the absence or presence of inhibitors (sulfobromophthalein,
taurocholate, probenecid, />-aminohippurate, and naringin) for inhibition studies. The transfected
CHO cells were incubated with 2 jiM E3S and 0, 0.1, or 1 mM perfluorocarboxylates with
carbon chain lengths ranging from 4 to 12 including PFOA (C8) for 30 seconds for E3S
inhibition studies.
In time-dependent uptake experiments, uptake of PFOA by OATplal-transfected cells
increased proportionally to time during the first 2 minutes of incubation. Vector-transfected cells
had a significant level of uptake of PFOA attributed to non-specific passive diffusion. In the
concentration-dependent uptake experiments, uptake velocity of PFOA in 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 uptake, the difference between the
uptake of the OATplal cells and the vector-transfected cells, was described by the Michaelis-
Menton equation and exhibited saturable kinetics.
Inhibition experiments with substrates of OATs and OATps showed that
sulfobromophthalein, taurocholate, and naringin inhibited PFOA uptake to 10-30% of control
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and/>-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
may be an important route of PFOA distribution, and that renal reabsorption in the male rat may
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 re-absorption of organic anions. Transformed cells derived from
human embryonic kidney 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 jiM [14C]PFOA for 15 minutes to determine
uptake. Transfected cells were also incubated with 10 jiM [14C]PFOA for 15 minutes and then
washed with incubation medium containing 1%, 3%, and 5% BSA to investigate the contribution
of non-specific binding of PFOA on the cell membrane. Experiments were conducted in
triplicate.
Uptake of PFOA was significantly stimulated (p<0.01) in cells transfected with human
OAT1, OAT3, and OAT4. Uptake of PFOA in human OAT1 transfected cells was 1.6 fold
higher than in control cells. In human OAT3 cells, PFOA uptake was -2.4 fold higher than in
control cells. In human OAT4 transfected cells, PFOA uptake was 2.7 fold higher than in
control cells. Accumulation of PFOA in transfected human OAT4 cells was also significantly
greater than in human OAT1 cells (p<0.01). Washing the cells with BSA reduced PFOA uptake
by 30% at most, suggesting mediation by the transporters into the transfected cells. The
experiments showed that human OAT4 transports PFOA and that human OAT4 activity may
play a role in reabsorption of PFOA from the tubule resulting in poor urinary excretion.
Yang et al. (2010) examined cellular uptake of PFOA by OATpl A2, OAT4, and URAT1
to determine their roles in mediating human renal reabsorption. Chinese hamster ovary (CHO)
and HEK293 cells were transfected with OATpl A2, OAT4, and URAT1 plasmid DNA or vector
DNA (control). In uptake studies, PFOA incubation times were 10 seconds (OAT4) or 30
seconds (URAT1). Cells transfected with OAT4 were incubated with 5 uM PFOA for up to 1
minute in time-dependent uptake experiments. In inhibition studies, cells transfected with OAT4
were incubated with 5 uM C14-PFOA for 10 seconds in the presence and absence of 100 uM
BSP, probenecid, glutarate, or polycyclic aromatic hydrocarbon (PAH). Perfluorochemicals with
differing chain lengths (C4-C12) were used in chain-length dependent inhibition experiments.
Incubations with 3H-E3S (OAT4, OATplA2) or 6 uM C14-uric acid(URATl) in the presence and
absence of 100 uM perfluorinated carboxylate acids (PFCs) lasted 10 seconds (OAT4), 30
seconds (OATplA2), or 1 minute (URAT1).
PFOA uptake in OATpl A2 transfected HEK293 cells was not different than uptake in
control cells. At 100 uM PFCs, 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
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uptake was observed with approximately 90% uptake at pH 8 and 250% at pH 5.5. In
concentration-dependent uptake experiments, uptake increased with increasing PFOA
concentration (0-1000 uM) 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 ESS 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 4-
fold compared to uptake in control cells. Time-dependent PFOA (5 uM) uptake by URAT1
increased with time during the 5 minute incubation period, and a concentration-dependent
increase in PFOA uptake was observed (0-700 uM). 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, the authors concluded that PFOA was not a substrate
for OATpl A2, but that OAT4 and URAT1 were probably involved in the renal reabsorption of
PFOA.
Weaver et al. (2010) published in vitro studies on the transport activities of the rat renal
transporters OAT1, OAT2, OAT3, OATplal, and URAT1. The transporters were transfected
into one of several cell lines and exposed to a series of PFCs having chain lengths ranging from 2
to 18 carbons (C). The activity of the PFC on the transporters was quantified on the basis of its
ability to inhibit the transport of a favored radiolabeled substrate. The PFC inhibition of the
individual transporters varied with chain length. The PFCs with 6, 7, and 8 carbon chains caused
a significant decrease in OAT1 transport of tritiated p-aminohippurate with the C7 acid having
the strongest effect. The PFCs with 5- through 10- carbon chains caused a significant decrease
in transport of tritiated estrone-3-sulfate by OAT3 with C8 and C9 acids having the strongest
effect. The transport of tritiated estadiol-17p-glucuronide by OATplal was significantly
inhibited by PFCs with 6 through 11 carbon chains with CIO acid having the strongest effect.
The PFCs did not inhibit OAT2 or URAT1 transport of favored substrates.
The kinetic response of the OAT1, OAT3, and OATplal transporters to increasing
concentrations of selected PFCs was also evaluated by Weaver et al. (2010). The change in
transport velocity (ng/mg protein/min) with increasing concentrations of the PFCs exhibited a
Michaelis-Menton type response. The kinetic data were analyzed to determine the Km and Vmax,
and those data are summarized in Table 3-21 below.
TABLE 3-21. Kinetic Parameters of PFC Transport by OAT1, OAT3 and OATplal
Transporter
OAT1
OATS
OATplal
PFC
C7
C8
C8
C9
C8
C9
CIO
Km(nM)
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
From Weaver etal., 2010
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The Michaelis-Menton kinetic data [Km and Vmax (maximum initial rate of an enzyme
catalyzed reaction)] indicate that there are substantial differences in the affinity of the PFCs with
8 and 9 carbon chains for 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 of the C7 and C8 acids by OAT1, an
export transporter on the basolateral surface of the renal tubular cells.
Based on the Hinderliter (2004) study, a developmental change in renal transport occurs
in female rats between 3 and 5 weeks of age that allows for expedited excretion of PFOA. When
the transporters become active, there is a decrease in plasma PFOA levels and an increase in
urinary excretion (Table 3-22). The developmental change in male rats appears to have the
opposite effect. Sexual maturity appears to influence these events because castrated male rats
become more like females and ovariectomized females become more like males in their PFOA
excretion capabilities. The change in female rats seems to involve the OATs (Kudo et al., 2002)
while that in males seems to involve the OATps (Cheng et al., 2006).
TABLE 3-22. Plasma and Urine PFOA Concentration 24 hr After Treatment with 30 mg/kg PFOA
Age
(weeks)
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
From Hinderliter (2004)
When considered together, the studies of the transporters suggest that female rats are
efficient in transporting PFOA across the basolateral and apical membranes of the proximal
kidney tubules into the glomerular filtrate but male rats are not. Males, on the other hand, have a
higher rate of resorption for the smaller amount they can transport into the glomerular filtrate via
OATplal in the apical membrane than females. This scenario may explain the inverse
relationship between the levels of PFOA in female urine and plasma and the plateau of plasma
PFOA in male rats when compared to their losses via urine.
Unfortunately much work remains to be done in order to explain the gender differences
between male and female rats and to determine whether it is relevant to humans. Similarities are
possible because of the long half-life in humans suggests that they may be more like the male rat
rather than the female rat. There is a broad range of half-lives in human epidemiology studies
suggesting a variability in human transport capabilities resulting from genetic variations in
structures and consequently in function. Genetic variations in human OATs and OATps have
been identified as described in a review by Zai'r et al. (2008).
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3.5 Toxicokinetic Considerations
3.5.1
PK Models
One of the earliest PK models (Andersen et al., 2006) was done using the post-dosing
plasma data from the Butenhoff et al. (2004b) study in cynomolgus monkeys. In this study,
groups of 6 monkeys (3/sex/group) were dosed for 26 weeks with 0, 3, 10, or 20 mg/kg PFOA
(high-dose =30 mg/kg PFOA for the first 12 days) and followed for >160 days after dosing.
Metabolism cages were used for overnight urine collection. Since urine specimens could
account for only overnight PFOA excretion, the total volume and total PFOA were extrapolated
to 24-hour values based on the excretion rate (volume/hour) for the volume collected and the
hours of collection.
The Andersen et al. (2006) model was based on the hypothesis that saturable resorption
capacity in the kidney would possibly account for the unique half-life properties of PFOA across
species and genders. The model structure (Figure 3-3; Andersen et al.2006) was derived from a
published model for glucose resorption from the glomerular filtrate via transporters on the apical
surface of renal tubule epithelial cells.
input
(iv, oral)
Tissue
Compartment
A
ku
L
1
r k»
Central
Compartment
(V^; C,; (-„.„„)
QB ,
i
L
TUB
Filtrate
Compartment
(Vni, Cm
C,,,
FIGURE 3-3. Schematic for a physiologically-motivated renal resorptions pharmacokinetic model.
The renal-resorption model includes a central compartment that receives the chemical
from the oral dose and a filtrate compartment for the glomerular filtrate from which resorption
and transfer to the central compartment can occur. Transfer from the filtrate compartment to the
central compartment decreases the rate of excretion. The resorption in the model was saturable,
meaning that there was less resorption and greater excretion at high serum PFOA concentrations
than at low concentrations.
The model was parameterized using the body weight and urine output of cynomolgus
monkeys (Butenhoff et al., 2004) and a cardiac output of 15 L/h-kg from the literature (Corley et
al., 1990). A 20% blood flow rate to the kidney was assumed based on data from humans and
dogs. Other parameters were optimized to fit the data for plasma and urine at lower
concentrations and then applied for the 20 mg/kg/day dose which was assumed to represent a
concentration where renal resorption was saturated. Based on the data for the dose of 20
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mg/kg/day, the model was able to predict the decline in plasma levels after the cessation of
dosing. The predictions were fairly adequate for one of the three modeled monkeys; for the
other two monkeys, the model predicted higher levels than were observed. This could be
because the model did not allow for efflux of PFOA into the glomerular filtrate through
transporters on the basolateral surface of the tubular cells. The authors observed that three
monkeys had faster renal clearance of PFOA than the other three monkeys.
Tan et al. (2008) divided the second compartment in the Anderson et al. (2006) model
into a liver compartment and a tissue compartment. A storage compartment was added between
the filtrate compartment and urinary excretion (Figure 3-4; Tan et al., 2008).
Input (oral)
UVer Compartment
(Liver volume; Free fraction in liver; Saturable binding)
Input
PL
Fecol Elimination
Central Compartment
(Volume of distribution; Free fraction in serum; Plasma cone)
f<13
Tissue Compartment
(Amount in tissue)
Tm.
Filtrate Compartment
(Volume of renal filtrate; Penal filtration rate;
Saturable resorption)
Storage Compartment
Urinary elimination
FIGURE 3-4. Physiologically-motivated pharmacokinetic model schematic for PFOA-exposed rats
The models were parameterized and applied to the Kemper (2003) data for male and
female CD rats given doses of 1, 5, or 25 mg/kg/day. The model did not provide a satisfactory
fit between the predictions of plasma concentration or urine + fecal excretion and experimental
data for either sex.
Lou et al. (2009) utilized the data they collected on the serum, liver, and kidney PFOA
concentration (see section 3.2) in CD-I mice to examine if one (Figure 3-5) or two compartment
pharmacokinetic (PK) models would fit the experimental data for 1,10, and 60 mg/kg/day single
gavage doses. Both models assumed first order absorption and elimination. The two
compartment model included a central compartment that received PFOA after absorption and
transferred it to a second compartment for excretion. The excretion compartment was coupled
with bidirectional flow between the two compartments. The net loss from the central
compartment differed during and after distribution. The models were fit using a general
nonlinear squares approach. A likelihood ratio squared approach was applied to determine
which model achieved the best fit to the data.
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Oral Dose
Ka
Central Compartment
Vd, Cl, Free PFOA
Ke
Urinary Loss
FIGURE 3-5. Schematic for one compartment model.
The one compartment model performed well for serum, liver, and kidney in this analysis
and output was not significantly improved with use of a two compartment model. The input
parameters for the one compartment model included volume of distribution (Vd), serum half-life,
and absorption (Ka) and elimination (Ke) rate constants for serum, liver, and kidney. There were
slight differences in the fitted values between males and females for some parameters. The Ke
values were consistently higher in the female mice (Table 3-23). The quantitative measures for
liver and kidney were only available for the 1 and 10 mg/kg/day doses.
TABLE 3-23. Model Parameters for 1 and 10 mg/kg Single Doses of PFOA to CD1 Mice
Tissue
Serum
Liver
Kidney
Parameter (abbreviation)
Volume of distribution (Vd)
Adsorption rate constant (Ka)
Elimination rate constant (Ke)
Half life (T !/2)
Volume of distribution (Vd)
Adsorption rate constant (Ka)
Elimination rate constant (Ke)
Volume of distribution (Vd) - 1 mg/kg
Volume of distribution (Vd) - 10 mg/kg
Adsorption rate constant (Ka)
Elimination rate constant (Ke)
Females
0.135L/kg
0.537 L/hr
0.00185 L/hr
15.6 days
0.161L/kg
0.5 170 L/hr
0.00161 L/hr
0.822 L/kg
1.092L/kg
0.527 L/hr
0.00151 L/hr
Males
0.266 L/kg
0.00133 L/hr
2 1.7 days
0.120 L/kg
0.00 129 L/hr
1.280 L/kg
1.170 L/kg
0.00113 L/hr
From Lou etal., 2009
The one compartment model described above was not able to predict serum concentration
in female mice given a single 60 mg/kg dose. This suggests a change in kinetics with the 60
mg/kg dose compared to the 1 and 10 mg/kg doses. This conclusion is supported by comparison
of the serum measurements made during the 30 day post dosing period for all three doses. The
serum PFOA concentration at the 60 mg/kg dose declined more rapidly with time than serum
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PFOA concentrations at the 1 and 10 mg/kg doses. For example, a serum concentration of about
0.4 mg/L was reached in about 28 days at 60 mg/kg dose, 61 days at 10 mg/kg, and 70 days at 1
mg/kg (values estimated from Figure 3, Lou et al., 2009). The one compartment model also
produced a poor fit for the serum level measurements taken 24 hours after the cessation of a 17-
day exposure to 20 mg/kg/day. The two compartment model provided a better fit with
experimental serum concentration data for the single 60 mg/kg dose and the repeat 20 mg/kg/day
dose, but the fit was still unsatisfactory.
The authors also tried the Andersen et al. (2006) renal-resorption model to determine if it
provided an improved fit for the data. The Andersen et al. (2006) model fit to the data was
superior to that of the one compartment and two compartment models of Lou et al. (2009) for the
60 mg/kg single dose and the 20 mg/kg/day repeat dose scenarios.
The Andersen et al. (2006) model includes a second tissue compartment that articulates
with the central compartment but not the filtrate compartment. In addition to values for Vd, Ka,
and Ke, the model includes values for cardiac output, volume for the renal filtrate, renal blood
filtration rate, intercompartmental clearance, transport maximum, transport affinity constant, and
the proportion of free PFOA in serum. With the exception of body weight and cardiac output,
the input parameters for the model were either assumed (volume of renal filtrate and proportion
of free serum PFOA) or optimized for the model. The wide confidence bounds around the
optimized values are indicative of considerable parameter uncertainty.
The Lou et al. (2009) parameter estimates indicate that there may be several biological
limitations to the Anderson et al. (2006) PK model for adult mice including the fact that it
requires an unreasonably high portion of the cardiac output to pass through the kidneys in order
to optimize fit to the experimental data. It also does not include excretion via export transporters
in the renal tubular cells and does not consider that the bound fraction in the serum could vary
with the magnitude of the dose and duration of dosing. Much of the emerging data is consistent
with a variety of tubular transporters functioning in both efflux and resorption from the
glomerular filtrate. In addition, there may also be opportunities for protein binding within
organs, including the liver, which could function to retard distribution especially at low doses
with changes occurring when binding sites become saturated.
A model has also been developed that applied to female CD-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 (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 were connected by a milk compartment. Milk yield information was obtained from
the literature. Milk was assumed to be consumed as it was produced without any circadian
impact on consumption patterns. PFOA excreted in pup urine was routed back to the dam.
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Both adsorption and excretion were assumed to be first order processes as was lactation
transfer from the dam to the litter (Figure 3-6; Rodriguez et al., 2009). Resorption of a portion of
the PFOA urinary efflux was included in the model. The renal excretion/resorption was
parameterized for cardiac output, kidney blood flow, glomerular filtration rate, urine flow rate,
volume of renal plasma (fraction of body weight), and volume of renal filtrate (fraction of body
weight). The fraction of free PFOA in serum reaching the glomerulus was assumed to be 0.01
based on protein binding information. As was the case with the Lou et al. (2009) model,
Rodriguez et al. (2009) did not include parameters to adjust for transporter-mediated efflux from
the renal tubular cells into the glomerular filtrate.
2
«
sf
Ordg
do
(mgj
Lactation: Day 19-39
Concept!
Vcon Cam
Q«"v "'Qcon
Kidney
Dam , . , Qur
cvaqe -S^ Fmrate =H "
kad , ••"-;«
,, ^ Cdom Vdam ^ Renal — *-"
"9J Qr-T^F Plasm*
1 r — '
klac -
Mill, Qr-Qur
1 1 r
Pup excreta Cmj(k Vm
rccirculation .
[from birth tePN&H) klOC
Pjps
Cpup Vpup
kep
T
. Urine
FIGURE 3-6. Pharmacokinetic model of gestation and lactation in mice.
One of the limitations of the Rodriguez et al. (2009) modeling effort was the limited
amount of laboratory data against which to evaluate projections. Serum measures from the Lau
et al. (2006) and Abbott et al. (2007) studies were available for only a few time points.
Nevertheless, the authors reached several conclusions based on the model projections as follows:
• The model had a tendency to overestimate serum levels suggesting nonlinearity as doses
increased.
• Gestation and lactation as a source of exposure contributed about equally to the pups of
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 clearance pathway for the dam.
The authors caution that the model is accompanied by a number of uncertainties because
of the assumptions regarding the flow limitation on transport to the fetus and to maternal milk,
the first order elimination rate constant for the pups, the milk: maternal serum partition
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coefficient, and the limited knowledge regarding the renal tubular transporters. They caution
that the model should not be applied for cross-species or high to low dose extrapolation.
Loccisano et al. (2011) developed a PFOA physiologically-based pharmacokinetic
(PBPK) model for monkeys based on the Anderson et al (2006) and Tan et al. (2008) models,
and then extrapolated it for use in humans (Figure 3-7). The model reflects saturable renal
absorption of urinary PFOA by the proximal tubule of the kidney. This is represented in Figure
3-7 by the interactions between the plasma and kidney plus the interaction of the filtrate
compartment with both plasma and kidney.
The fraction of PFOA free in plasma and available for glomerular filtration was based on
data fit and estimated to be less than 10% because of binding to serum proteins, especially
albumin. Lacking the kinetic data on tubular resorption, the rate was based on the best fit to the
plasma/urine data. A storage compartment was added to the model between filtrate and urine.
Tissue plasma partition coefficients were derived from the data by Kudo et al. (2007) following
the disposition pattern of a single IV dose to male Wistar rats.
Plasma
Free
fraction
,,
QGut
3>
QLiv
-.,
QFat
^
QSkn
=>
QR
.
QKid
QFil
Gut
S t
Liver
Fat
Skin
Rest of body
Kidney
j\
Tm.Kt
Filtrate
\ f
Oral dose, drinking water
storage
kurine
urine
Tm = transporter maximum, Kt= affinity constant and Q= flow in and out of tissues
FIGURE 3-7. Structure of the PFOA PBPK model in monkeys and humans
Existing i.v. and oral data sets from Butanhoff et al. (2004b) for the cynomolgus monkey
were used to develop the monkey model. In the oral study (Section 3.2), animals were dosed for
six months and followed for 90 days after dosing. Plasma and urine samples were analyzed
periodically during dosing and recovery. The model projections for the oral study were in good
agreement with the Butanhoff et al. (2004b) data for the 10 mg/kg dose showing a rapid rise to
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plasma steady state and a slow terminal half-life. The model performance for the high dose
(30/20 mg/kg/day) did not fit as well, partially as consequence of the observed toxicity with the
initial 30 mg/kg/day dose that necessitated cessation of dosing on study day 12, followed by
resumption of dosing at 20 mg/kg/day on study day 22.
The structure of the human model was similar to that used for the monkeys. Human
serum data (means with standard deviations or medians) for PFOA are available for occupational
and general populations (Emmett et al., 2006; Calafat et al., 2007 a,b; Olsen et al., 2005; Holzer
et al., 2008; Steenland et al., 2009; Bartell et al., 2010). The fact that the serum data were the
results from measurements made following uncertain routes and uncertain exposure durations
presented a challenge in the assessment of model fit. The human half-lives used for the model
(3.8 and 2.3 years) came from an occupational study (Olsen et al., 2005) and a study of the Little
Hocking population after reduction of the PFOA in drinking water as a result of treatment
(Bartell et al., 2010); see section 3.5.2. In each case the major source of exposure was assumed
to be the drinking water. Both half-life values were used in evaluating the model's ability to
predict serum concentration at the time the serum samples were collected.
The model produced results that can be characterized as fair to good when compared to
the reported average serum measurements. For the Little Hocking Population studied by Emmett
et al. (2006), the model indicated the need for a 30 year exposure to reach steady state
concentrations. The model indicated that both half-life values provided reasonable results
compared to the measured serum values. The authors concluded that more data are needed on
the kinetics of renal transporters, intra-human variability, and more definitive information on
exposures in order to further refine the human model.
Loccisano et al. (2012a) also utilized the saturable resorption hypothesis when
developing a model for adult Sprague Dawley Rats (Figure 3-8). The structure of the model is
similar to that for the monkey/human model depicted in Figure 3-7 but lacks the fat and skin
compartments and includes a storage compartment to accommodate fecal loss of unabsorbed
dietary PFOA as well as that from biliary secretions. Oral and IV data used in model
development came from studies by Kemper (2003), Kudo et al. (2007) and Perkins et al. (2004).
Partition coefficients for 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. Most of the other kinetic parameters were based on values providing the best fit to the
experimental data. Because a number of the renal transporters involved with PFOA resorption
are known; available kinetic information was utilized where appropriate. Model performance
was evaluated primarily based on its ability to predict plasma and liver concentrations from the
studies identified above. As was the case for the monkey model, performance was generally
good given the limitations in the primary data sources.
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Oral, diet
faces
FIGURE 3-8. Structure of the PBPK Model for PFOA in the Adult Sprague Dawley Rat
Loccisano et al. (2012b) expanded the adult Sprague Dawley rat model described above
to cover gestational and lactational exposure to the fetus and pups through their dams. The data
from Hinderletter et al. (2005) were used in model development for both the gestation and
lactation periods. The gestational model structure for the dams is similar to Figure 3-8. The
model was expanded to include the fetuses linked to the dams by way of the placenta. Uptake
from the placenta was described by simple diffusion; the fetal plasma compartment was separate
from the dams as was distribution to fetal tissues and amniotic fluid. Based on the transporter
data for PFOA, elimination differed for male and female rats and was considered to be
developmentally regulated resulting in faster elimination for female fats than male rats after
sexual maturation. The lactation model linked the pups to their dams through the mammary
gland secretions. Pup compartments included the gut, liver, kidney, renal filtrate, plasma and
rest of the body.
Model performance was judged by its ability to predict concentrations in maternal and
fetal plasma, amniotic fluid and milk. The predictive capability of the model ranged from fair to
good depending on the medium. The fit of the projections to the data was weakest for the whole
embryo during gestation where measured levels were greater than projection for two of three
data points and for neonate plasma during lactation where all data points fell below the
predictions.
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3.5.2 Half-Life Data
Human. There have been several studies of half-lives in humans and all support a long residence
time for serum PFOA with estimates measured in years rather than months or weeks. Bartell et
al. (2010) determined an average half-life of 2.3 years based on a study of the decreases in
human serum levels after treatment of drinking water for PFOA removal was instituted by the
Lubeck Public Services District and the Little Hocking Water Association in West Virginia
(WV). Source waters for these systems had become contaminated with PFCs from the Dupont
Works Plant in Washington, WV between 1951 and 2000.
The Bartell et al. (2010) study was based on a series of serum measurements (8 over 4
years) from 200 individuals who agreed to participate in the study. Inclusion criteria for the
participants included: serum PFOA concentrations > 50 ng/mL, residential water service
provided by one of the two treatment plants, never employed at the Dupont plant, not growing
their own vegetables, and signed acceptance of the study consent form. The participants were
almost equally divided between males and females with an average age of about 50 years (range
18-89 years). Most of the participants consumed public tap water (172) as their primary cource,
but a small number (28) consumed bottled water as their source.
The participants were required to report primary use of home tap water for cooking,
bathing, and showering between 2005-2007. The tap water users had to report public water as
their primary source of residential water consumption, and bottled water users had to report the
use of bottled water as their primary source of residential water consumption. The initial blood
draw for serum occurred in June 2007, with subsequent samples at 1, 2, 3, 6, and 12 months after
the initial sample. Samples were analyzed by the Centers for Disease Control and Prevention
(CDC), and 19 samples from the 2-month blood draw were not analyzed due to mislabeling.
A linear mixed model was used to determine the decline in serum PFOA concentration
over time. With these models, the decline from baseline by the participants was essentially first
order. The serum PFOA concentration was the only time-varying measurement entered into the
model. Serum concentrations were log normally distributed, as described by the following
equation:
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/year after
adjustment for covariates and a half-life of 2.3 years [confidence interval (CI) = 2.1-2.4]. The
covariates considered included the water treatment system, the time exposed before and after
filtration, public vs. bottled water, sex, age, consumption of local or home grown vegetables, and
exposure to the public water supply at work. The only potential confounders determined to be
significant were the treatment plant (p=0.03) and home grown vegetable consumption (p<0.001).
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Identification of consumption of home grown vegetables as a significant confounder
revealed a weakness in the study design because it had been an exclusion factor, yet was
identified as an exposure source at the 12-month interview of the study participants. The
researchers concluded that this problem was a result of the way the exclusion question was
phrased for the original interview, "Do you grow your own vegetables?". When the question
was asked later in the study, it was phrased, "Do you eat any fruits and vegetables grown at your
own home?" Some people who answered no to the original questions andwered yes to the
second question.
Changes in the source of drinking water during the study could have also impacted the
results. When baseline interview data were compared with the results from the 12 month
interview, 39% of the bottled water group reported using public water at home. Some of the
public water drinkers (10%) reported using primarily bottled water at the 6 month interview.
In another study, the drinking water supply was contaminated with a mixture of
perfluorinated chemicals when a soil-improver mixed with industrial waste was applied upriver
to agricultural lands in Arnsberg, Germany (Brede et al., 2011). The PFOA levels in the finished
drinking water were measured as 500-640 ng/L in 2006. PFOS and PFHxS were also present.
The plasma PFOA levels in the Arnesberg population were 4.5 to 8.3 times higher than those in a
reference community at the time the problem was discovered. Charcoal filtration was added to
the potable treatment train and succeeded in reducing the PFCs in the drinking water.
The authors used the differences in plasma 2008 PFOA measurements from a subset of
the participants (children and adults) initially exposed in 2006 to determine the PFOA half-life.
The 2008 subjects included 66 men women and children from Arnsberg and 73 from the
reference community in their evaluation. The drinking water concentration monitoring results
(not detects estimated as 1A the LOD, 10 ng/L) and drinking water intake estimates obtained by
questionnaire and interview, were used to estimate PFOA exposures. Plasma PFOA samples
were collected during a two month period in late 2008. Plasma PFOA had declined in the serum
for both the Arnsberg residents (39.2%) and those from the reference community (13.4%). In
Arnsberg the decrease was greater for the exposed women and children than the men when
compared to the reference community, an observation that appeared to reflect the reported lower
drinking water intakes of the Arnsberg women and children (0.3±0.2 and 0.8±0.6 L/day
compared to 0.7±0.5 and 1.6±0.8 L/day, respectively). The estimate for the human half-life was
3.26 years (geometric mean; range 1.03-14.67 years). Regression analysis of the data also
suggested that the elimination rate may have been greater in the younger subjects and older
subjects.
Seals et al. (2011) determined half-life estimates for 602 residents of Little Hocking, OH
and 971 residents of Lubeck, WV who were part of the C8 study but had relocated to a different
area of the country. The half-life estimate was based on the decline in serum PFOA levels after
the time of the initial measurement and the years since the change in residential location
occurred. A background estimate (5 ng/ml) was subtracted from the serum measurements before
analysis. On average, the initial serum PFOA concentrations were higher in Little Hocking (60.6
ng/mL) than in Lubeck (31.0 ng/mL). Due to the nonlinearity in scatter plots of the natural log
for adjusted serum PFOA concentrations versus the years elapsed since relocation, the authors
used a two segment (Little Hocking- 4 years, Lubeck-9 years) linear spline regression approach
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in their analysis of the data. The slope of the line decreased for the second time segment as
compared to the first. In former residents of Little Hocking, a -21.4% change in serum PFOA
was observed in the first 4 years after leaving Little Hocking, and a -7.6% change was observed
after 4 years. In former Lubeck residents, the serum PFOA change was -7.8% for the first 9
years and 0.2% (a slight increase) afterwards. The half-life estimates for Little Hocking ranged
from 2.5-3.0 years (average 2.9 years) and those for Lubeck ranged from 5.9-10.3 years (average
8.5 years).
Based on their analysis, the authors suggested that, if their assumptions were correct, a
simple first order elimination model may not be appropriate for PFOA given that the rate of
elimination appeared to be influenced by both concentration and time. There was a difference in
the clearance for the two locations even though the range of years elapsed since relocation was
the same for both communities. The authors identified three potential limitations of their
analysis: the cross-sectional design, the assumption that exposure was uniform within a water
district, and a potential bias introduced by exclusion of individuals with serum values <15
ng/mL.
3M (Burris et al., 2000; Burris et al., 2002) conducted a half-life study on 26 retired
fluorochemical production workers from their Decatur, Alabama (n = 24) and Cottage Grove,
Minnesota (n = 3) plants. Blood was collected from the subjects between 1998 and 2004, a
period during which serum samples were drawn every 6 months over a 5-year period, depending
on the facility at which the subject had worked. Responses on questionnaires determined
whether any of the retirees had occupational exposures after retirement. The average number of
years that participants worked was 31 (range 20-36 years) and they had been retired an average
of 2.6 years at study initiation (range 0.4 - 11.5 years). The mean age of the retirees was 61 years
(range 55-75) at the beginning of the study.
The initial mean serum PFOA concentration of all of the subjects was 0.691 |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
sex of retiree, the number of years retired or worked at the production facility, or medication use
or health conditions.
Harada et al. (2005a) studied the relationship between age, sex and serum PFOA
concentration in residents of Kyoto, Japan. They found that females in the 20-50-year-old age
group (all with regular menstrual cycles) had serum PFOA concentrations that were significantly
lower than those in females over age 50 (all postmenopausal). Mean serum PFOA concentration
in the younger females was 7.89 ± 3.61 ng/ml versus 12.63 ± 2.42 ng/ml in the older females.
This age difference in serum PFOA concentrations was not seen in males, and serum PFOA
concentrations in males were comparable to those of the older females.
Harada et al. (2005b) also estimated the renal clearance rate of PFOA in humans and
found it to be only about 0.001% of the glomerular filtration rate, indicating the absence of active
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excretion in human kidneys. There was no significant difference in renal clearance of PFOA with
respect to sex or age group, and the mean value was 0.03 ± 0.013 ml/day/kg.
Animal. Kemper (2003) examined the plasma concentration profile of PFOA following gavage
administration in sexually mature Sprague-Dawley rats. Male and female rats (4/sex/group) were
administered single doses of PFOA by gavage at dose rates of 0.1, 1, 5, and 25 mg/kg PFOA.
After dosing, plasma was collected for 22 days in males and 5 days in females. Plasma
concentration vs. time data were then analyzed by non-compartmental pharmacokinetic methods
(Table 3-24 and Table 3-25). To further characterize plasma elimination kinetics, animals were
given oral PFOA at a rate of 0.1 mg/kg, and plasma samples were collected until PFOA
concentrations fell below quantitation limits (extended time).
Plasma elimination curves were linear with respect to time in male rats at all dose levels.
In males, plasma elimination half-lives were independent of dose level and ranged from
approximately 138 hours to 202 hours. To further characterize plasma elimination kinetics,
particularly in male rats, animals were given oral PFOA at a dose of 0.1 mg/kg, and plasma
samples were collected until PFOA concentrations fell below quantitation limits (2016 hours in
males). The estimated plasma elimination half-life in this experiment was approximately 277
hours (11.5 days) in male rats.
TABLE 3-24. Pharmacokinetic Parameters in Male Rats Following Administration of PFOA
Parameter
Tmax (hr)
Qnax (Hg/ml)
Lambda z (1/hr)
T1/2 (hr)
AUCiNp (hr ng/ml)
AUCWD
(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
(i.v.)
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)
From Kemper, 2003
(Mean (SD))
AUCINF: area under the plasma concentration time curve, extrapolated to infininty; AUCINF/D: AUCINF normalized to dose; Clp:
plasma clearance; Cmax: maximum plasma concentration; Lambda z: terminal elimination constant; T1/2: terminal elimination
half-life; Tmax: time to Cmax.
Plasma elimination curves were biphasic in females at the 5 mg/kg and 25 mg/kg dose
levels. In females, terminal elimination half-lives ranged from approximately 2.8 hours at the
lowest dose to approximately 16 hours at the high dose. The estimated plasma elimination half-
life in the extended time experiment was approximately 3.4 hours in females.
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TABLE 3-25. Pharmacokinetic Parameters in Female Rats Following Administration of PFOA
Parameter
Tmax (hr)
Qnax (Hg/ml)
Lambda z (1/hr)
T1/2 (hr)
AUCW (hr ng/ml)
AUCW/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
(i.v.)
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)
3.34
(0.32)
34.39
(3.29)
29.30
(3.06)
From Kemper, 2003
(Mean (SD))
• 14,
Gibson and Johnson (1979) administered a single dose of C-PFOA averaging 11.4
mg/kg by gavage to groups of 3 male 10-week old CD rats. The elimination half-life of carbon-
14 from the plasma was 4.8 days. NRC (2005) reported half-lives of 4-6 days for male rats and
2-4 hours for female rats; there was no mention of the strains studied. These values are
consistent with the half-lives reported by Kemper (2003) for male and female Sprague-Dawley
rats.
Lou et al. (2009) determined values of 21.7 days (95% confidence interval: 19.5-24.1)
for male CD1 mice and 15.6 days (95% confidence interval: 14.7-16.5) for females for use in
their pharmacokinetic model (see Section 3.5.1). NRC (2005) provided values of 12 days for
males and 20 days for females without any information on strains.
Butenhoff et al. (2004b) looked at the elimination half-life in monkeys treated for 6
months with 0, 3, 10 or 20 mg/kg/day via capsules. Elimination of PFOA from serum after
cessation of dosing was monitored in recovery monkeys from the 10 and 20 mg/kg dose groups.
For the two monkeys exposed to 10 mg/kg, serum PFOA elimination half-life was 19.5
(R2=0.98) days and indicated first-order elimination kinetics. For three monkeys exposed to 20
mg/kg, serum PFOA elimination half-life was 20.8 days (R2=0.82) and also indicated first-order
elimination kinetics although dosing was suspended at different time points because of weight
loss. The data from NRC (2005) which were provided by Butenhoff were about 21 days for
females and 30 days for males.
3.5.3 Volume of Distribution Data
Several researchers have attempted to characterize PFOA exposure and intake in humans
(Thompson et al., 2010; Lorber and Egeghy, 2011) through pharmacokinetic modeling. As an
integral part of model validation, the parameter for volume of distribution of PFOA within the
body was calibrated from the available data. In the models discussed below, volume of
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distribution was defined as the total amount of PFOA in the body divided by the blood or serum
concentration.
Two groups of researchers defined a volume of distribution of 170 ml/kg body weight for
humans for use in a simple, single compartment, first-order pharmacokinetic model (Thompson
et al., 2010; Lorber and Egeghy, 2011). The models developed by these groups were designed to
estimate intakes of PFOA by young children and adults (Lorber and Egeghy, 2011) and the
general population of urban areas on the east coast of Australia (Thompson et al., 2010). In both
models, the volume of distribution was calibrated using human serum concentration and
exposure data from the National Health and Nutrition Examination Survey (NHANES) and
assumes that most PFOA intake is from contaminated drinking water. Thus, in using the models
to derive an intake from contaminated water, the value of volume of distribution was calibrated
so that model prediction of elevated blood levels of PFOA matched those seen in residents.
Butenhoff et al. (2004b) calculated a volume of distribution from non-compartmental
pharmacokinetic analysis of data from cynomolgus monkeys. Three males and three females
were administered a single intravenous dose of 10 mg/kg and serum PFOA concentrations were
measured in samples collected up to 123 days post-dosing. The volumes of distribution of PFOA
at steady state (Vdss) were similar for both sexes at 181 ± 12 ml/kg for males and 198 ± 69
ml/kg for females.
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4.0 HAZARD IDENTIFICATION
4.1 Human Effects
4.1.1 Long-Term and Epidemiological Studies
Large-scale production of PFOA occurred in the United States for several decades. The
employees of the main production companies and the residents near the plants were exposed to
the compound over time. There are several studies that have examined the data available on
these occupational and residential populations. Both 3M and DuPont have been the primary U.S.
producers and users of perfluorinated compounds, and both companies have been offering
voluntary fluorochemical medical surveillance programs for many years to workers at plants that
produce or use perfluorinated compounds. The monitoring data collected by 3M and DuPont
were used in conjunction with mortality and health effect information in a number of
epidemiological studies on their worker populations. 3M discontinued manufacturing PFOA in
2000, but a subsidiary in Europe continued to manufacture and sell it through 2008.
The data from NHANES have been the subject of several studies investigating the
association between PFOA and health effects in the general population of the United States. The
NHANES studies have examined representative members of the U.S. population (-5000 adults
and children/year) through their surveys focusing on different health topics. The studies consist
of an interview (demographic, socioeconomic, dietary, and medical questions) and examination
(medical including blood and urine collection, dental, and physiological). Serum samples from
the studies were analyzed for PFOA concentration by solid-phase extraction coupled to isotope
dilution/high-performance liquid chromatography/ tandem mass spectrometry.
Members of the general population living in the vicinity of the West Virginia DuPont
Washington Works PFOA production plant in Parkersburg, WV, are the focus of an ongoing
study titled the C8 Health Project. Releases from the Washington Works plant, where PFOA
(C8) was used as a processing aid in the manufacture of fluoropolymers, contaminated the
ground water from six water districts near the plant resulting in exposures to the general
population. The purpose of the C8 Health Project is to assess if there are any "probable links"
between PFOA exposure and disease. During August 2005, through July 2006, about 69,000
study participants were identified. Eligible participants included those who had consumed
drinking water for at least one year up to and including December 4, 2004 from the Lubeck and
Mason County water districts in West Virginia, Belpre, Little Hocking, Tuppers Plains-Chester,
and Pomeroy water districts in Ohio, or private water source within the geographical boundaries
of the public water sources. The participants (n=69,030; 33,242 males, 35,788 females; <10 to
70+ years) donated a blood sample, filled out an extensive questionnaire, and received $400 in
compensation. Serum was analyzed for perfluorochemicals, including PFOA by liquid
chromatography separation with tandem mass spectrometry. Clinical chemistry tests were
carried out if serum quantities were sufficient. Medical records were used to validate diseases
reported by participants.
The results of these studies along with other population studies are described in the
following sections.
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4.1.1.1 Non Cancer Systemic Toxicity Studies
Several studies have been undertaken that examine the relationship between serum PFOA
concentration and various health outcomes suggested by the standard animal toxicological
database. Many of these monitored standard clinical chemistry parameters used within the
occupational and public health communities as indicators of health status. In the studies of
worker cohorts, the data collected were focused towards measures of cardiovascular risk, signs
of organ damage, and standard haematological endpoints. Within the general population, data
were focused on cardiovascular risk factors and diabetic or prediabetic conditions as well as
reproductive and developmental endpoints. In reviewing the epidemiological monitoring data
for PFOA it is important to remember that many individuals were exposed to other fluorocarbons
and, although only serum PFOA measurements are presented in the study summaries that follow,
many individuals had other PFCs in their serum as well.
Serum Lipids and Cardiovascular-Related Outcomes
Occupational Population Studies
Serum lipids have been a focus in many of the occupational studies because data from
animal studies shows alterations in standard profiles for total cholesterol and triglycerides
following exposure to PFOA. A number of the studies discussed in the following paragraphs
also examined endpoints in addition to those associated with cardiovascular disease.
Descriptions of the outcomes related to standard clinical biochemistry endpoints and
hematological effects are found in the sections that follow the lipid and cardiovascular-related
observations.
Olsen et al. (2000) analyzed data from voluntary medical surveillance examinations
conducted in 1993, 1995, and 1997, in PFOA production plant workers to determine if serum
PFOA concentration in workers was associated with serum lipid alterations. Cholesterol, low
density lipoprotein (LDL), high density lipoprotein (HDL), and triglycerides were measured in
male workers; n=l 11 in 1993; n= 80 in 1995, n= 74 in 1997. Multivariable regression analyses
were adjusted for age, body mass index (BMI), alcohol use, and cigarette use. Employees'
serum PFOA levels were stratified into 3 categories (<1, 1- <10, and >10 ug/mL). Categories
were chosen in order to maximize the number of employees in the highest exposure category.
Serum PFOA levels ranged from 0.0-80.0, 0.0-114.1, and 0.1-81.3 ug/mL for the years 1993,
1995, and 1997, respectively. No association was observed between serum PFOA concentration
and total cholesterol, LDL, HDL, or triglycerides in the workers in 1993, 1995, or 1997.
A cross-sectional analysis of the data from the 2000 medical surveillance program at the
Decatur and Antwerp manufacturing plants was undertaken to determine if there were any
associations between PFOA exposure and serum lipids of monitored employees (Olsen et al.,
200la, 2003). Multivariable regression analyses were conducted to adjust for possible
confounders including the following variables: production job (yes or no), plant, age, BMI,
cigarettes/day, drinks/day and years worked at the plant. Mean PFOA serum levels were 1.03
ug/mL for all male employees at the Antwerp plant and 1.90 ug/mL for all male employees at
the Decatur plant. Male production employees at the Decatur plant had significantly higher (p
<0.05) mean serum levels (2.34 ug/mL) than those at the Antwerp plant (1.28 ug/mL). Mean
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PFOA serum concentrations were 0.07 ug/mL for female employees at the Antwerp plant and
1.23 ug/mL for female employees at the Decatur plant. A positive significant association was
reported between serum PFOA and cholesterol (p = 0.05) and triglycerides (p = 0.002). Age was
also significant in both analyses. Alcohol consumed per day was significant in the cholesterol
model, while BMI and cigarettes smoked per day was significant for triglycerides.
A longitudinal analysis of the above data and previous medical surveillance results was
carried out to determine whether occupational exposure to fluorochemicals over time was related
to changes in serum lipids in employees of the Antwerp and Decatur facilities (Olsen et al.,
200Ib, 2003). Medical surveillance data from 1995, 1997, and 2000 were analyzed using
multivariable regression analysis. The plants were analyzed using 3 subcohorts that included
those who participated in 2 or more medical exams between 1995 and 2000. A total of 175 male
employees voluntarily participated in the 2000 surveillance and at least one other. Only 41
employees were participants in all 3 surveillance periods.
When mean serum PFOA levels were compared by surveillance year, PFOA levels in the
employees participating in medical surveillance at the Antwerp plant increased between 1994/95
and 1997 and then decreased slightly between 1997 and 2000. At the Decatur plant, PFOA serum
levels decreased between 1994/95 and 1997 and then increased between 1997 and 2000. When
analyzed using mixed model multivariable regression and combining Antwerp and Decatur
employees, there was a statistically significant positive association between PFOA and serum
cholesterol (p = 0.0008) and triglycerides (p = 0.0002) over time. When analyzed by plant and
also by subcohort, these associations were limited to the Antwerp employees (p = 0.005) and, in
particular, the 21 Antwerp employees who participated in all 3 surveillance years (p = 0.001).
However, the association between PFOA and triglycerides was also statistically significant (p =
0.02) for the subgroup in which employees participated in biomonitoring in 1994/95 and 2000.
There was not a significant association between PFOA and triglycerides among Decatur workers.
There were no significant associations between PFOA and changes over time in FtDL.
Limitations to the 2000 cross-sectional and longitudinal studies included but were not limited to
the consistent differences in serum PFOA concentration and demographics between employees
of the Decatur and Antwerp plants, possible confounding by other perfluorinated chemicals, and
differences in analytical techniques.
A cross-sectional study was undertaken to determine the relationship between serum
PFOA and lipids in a large population of active employees (Washington Works fluoropolymer
production plant) with potential exposure to PFOA (Sakr et al., 2007a). The employees who
volunteered (n=1025, 782 males, 243 females) to participate in the study had a physical
examination, provided fasting blood samples, and answered a medical and occupation history
questionnaire in 2004. The mean age and years of work were 46.5 and 19.6 years, respectively,
for males and 44.4 and 15.9 years, respectively, for females. Serum lipids including cholesterol,
FIDL, LDL, very low density lipoprotein (VLDL), triglyceride, uric acid (risk factor for
hypertension) and serum PFOA concentrations were determined from blood samples. Serum
PFOA concentration was analyzed by liquid chromatography tandem mass spectrometry. The
relationship between PFOA exposure and lipid levels was evaluated by analysis of variance
(ANOVA), %2 test, t test, and linear regression models. Confounders including age, BMI,
gender, alcohol consumption, and parental heart attack were considered in the models.
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Serum PFOA concentration in the workers ranged from 0.0046 to 9.55 ug/mL. For those
with current occupational exposure to PFOA, the range was 0.0174 to 9.55 ug/mL, and 0.0081 to
2.07 ug/mL for workers with intermittent occupational exposure. The range was 0.0086 to 2.59
ug/mL for workers with past occupational exposure and 0.0046 to 0.963 ug/mL for workers with
no occupational exposure. Serum PFOA was positively associated with cholesterol, VLDL, and
LDL (p<0.03) in the participating workers whether or not they were taking lipid-lowering
medication. An association between uric acid and serum PFOA was also observed. No
association was observed between serum PFOA and HDL or triglycerides. Limitations of the
study included the study design (cross-sectional surveys which cannot be used to determine
causality), and the limitations of the questionnaire which was not designed or validated for the
study.
Sakr et al. (2007b) conducted a longitudinal study of serum lipids in workers of the
Washington Works plant from 1979 to 2004. Employee medical records from the medical
surveillance program were used to obtain blood lipid (total cholesterol, LDL, HDL, triglycerides,
and uric acid), height, and weight data. As part of the medical surveillance program, employees
gave a detailed medical history and had a physical examination at least every 3 years. Serum
PFOA concentration was measured every 1 to 2 years in PFOA-exposed workers and every 3 to
5 years in non PFOA-exposed workers on a volunteer basis. This study included 454 workers
who had 2 or more serum PFOA measurements. The study population included 334 males and
120 females ranging in age from 24 to 66 years who had worked at the plant for at least one year
since 1979. A linear mixed effects regression model was used to analyze the data. Age, gender,
BMI, and decade of hire were potential confounders.
Serum PFOA concentration ranged from 0 to 22.66 ug/mL with a mean of 1.13 ug/mL
over the 23 year monitoring period in the study population. Serum PFOA concentration was
positively associated with total cholesterol after age, BMI, gender, and decade of hire adjustment
in the model. Limitations of the study included inclusion of workers who had at least 2 PFOA
measurements (only 16% of all workers in the PFOA area), lack of information on lipid-lowering
medications and alcohol intake, and the exposure and outcome were not measured on the same
date.
Olsen and Zobel (2007) examined the association between serum PFOA concentration
and serum lipids in fluorochemical workers based on data collected from the 3M medical
surveillance program in 2000. The fluorochemical workers consisted of males (age 21-67) from
the Antwerp, Belgium (n=196), Cottage Grove, MN (n=122), and Decatur, AL (n=188)
production facilities who volunteered to participate in the medical surveillance program and did
not take cholesterol lowering medication. Blood was collected for fluorochemical concentration
determination and serum lipid parameters including cholesterol, LDL, HDL, and triglycerides.
PFOA concentration was determined by liquid chromatography/tandem mass spectrometry.
Analysis of variance, analysis of covariance, logistic regression, and multiple regression models
were used to analyze the data with age, BMI, and alcohol consumption as covariates.
Serum PFOA concentration ranged from 0.04 to 92.03 ug/mL for the male workers (all
sites combined). The range was 0.01-7.04, 0.01-92.03, and 0.04-12.7 ug/mL for the Antwerp,
Cottage Grove, and Decatur sites, respectively. The mean serum PFOA concentration was 2.21,
1.02, 4.63, and 1.89 for all sites combined, and the Antwerp, Cottage Grove, and Decatur sites,
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respectively. Serum PFOA (all sites combined) was not associated with total cholesterol or
LDL. A negative association was observed between serum PFOA concentration (all sites
combined) and HDL. However, no association was observed at the individual sites for HDL. It
was hypothesized that the results may be due to uncontrolled residual confounding. Serum
triglyceride was positively associated with serum PFOA at all sites combined and the Antwerp
site. Non-adherence to fast requirement for blood collection and non-causal positive correlation
were provided as possible hypotheses for the results.
Costa et al. (2009) provided additional serum lipid data using 30 years of medical
surveillance data from workers of a PFOA production plant in Italy. The workers (n=53 males,
20-63 years of age) participated in the medical surveillance program yearly from 1978 to 2007.
The length of work exposure was 0.5 to 32.5 years. In 2007, 37 men were active workers, and
16 men were retired or had transferred to other departments and were no longer being exposed.
Non-exposed male workers, (n=107, 12 executives and 95 blue collar workers) from different
departments also participated in the medical surveillance program and served as controls. The
medical surveillance program consisted of a yearly physical examination and blood chemistry
tests including Apo-A and Apo-B lipoproteins, total cholesterol, HDL, triglycerides, and uric
acid analysis. Beginning in 2000, serum PFOA was monitored yearly except in 2005.
Determination of serum PFOA concentration was made by HPLC-electrospray-tandem mass
spectrometry. The relationship between PFOA exposure and serum lipids was evaluated by
ANOVA, t test, and multiple linear regression models. Confounders including age, BMI,
smoking, alcohol consumption, and shift work were considered in the models.
Serum PFOA concentration of the exposed workers ranged from 1.54-86.3, 0.73-91.9,
0.34-91.9, 0.38-74.7, 0.54-46.3, 0.54-41.9, and 0.2-47.0 ug/mL for the years 2000, 2001, 2002,
2003, 2004, 2006, and 2007, respectively. Serum PFOA concentrations in the workers decreased
after plant renovations partially automated the PFOA production process and adopted strict
working procedures with the use of protective devices in 2002. Total cholesterol and uric acid
were significantly increased (p<0.05) in 34 workers currently exposed to PFOA when compared
to 34 non-exposed workers (age and work matched). Regression analysis comparing those same
34 exposed workers to all non-exposed workers (n=107) revealed a significant effect (p<0.05) of
PFOA exposure on total cholesterol and uric acid levels. Serum PFOA concentration was
correlated (p<0.05) with total cholesterol and uric acid levels in 56 workers (formally, currently,
and never exposed) assessed concurrently from 2000-2007. No correlations were observed
between serum PFOA concentration and Apo-A or B lipoproteins, HDL, or triglycerides in any
of the analyses. Due to the small number of subjects, the authors were hesitant to draw firm
conclusions, but suggested further exploration into these correlations.
Leonard et al. (2008) examined mortality in employees at the DuPont Washington Works
plant in West Virginia after a cross sectional survey of 1,025 active employees revealed a
positive association between serum PFOA and cholesterol and LDL. The association led the
authors to target ischemic heart disease (IHD) as well as many other causes of death as
outcomes. The cohort consisted of 6,027 employees (4,872 males and 1,155 females) who had
ever worked at the plant from 1948 through 2002. The DuPont Epidemiology Registry and US
National Death Index were used to obtain causes of death. Standardized mortality ratios were
estimated using three reference populations; the populations of the US and West Virginia and the
DuPont regional worker reference population excluding workers at the Washington Works plant.
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A non-signifcant elevation for IHD mortality (SMR= 109, 95% CI: 96, 124) and a significant
increase in diabetes mortality [standardized mortality ratio (SMR) = 197, 95% CI: 123, 298]
were observed for Washington Works plant workers compared to the DuPont regional worker
reference population. No increases were observed in these two outcomes when compared to the
national and state populations. No significant associations were observed for the other causes of
death. Using populations spanning from 1948 to 2002 increased the possibilities for loss to
follow-up and inadequate records. Also caution must be exercised in interpreting the results of
mortality analyses that examine diseases that do not uniformly cause death.
Sakr et al. (2009) further examined IHD mortality in workers at the Washington Works
plant because of the association between PFOA exposure and increased serum lipid levels, a risk
factor for heart disease. The cohort consisted of 4,747 employees (4,460 males and 102 females)
who had worked in the plant from 1948 through 2002. The average duration of employment was
22.5 years. Serum PFOA levels were linked with job titles and used to determine PFOA
exposure for each job. Jobs were divided into categories for low, medium, or high exposures,
and intensity factors were assigned that corresponded to worker mean serum PFOA
concentration for all the jobs in each category. The cumulative exposure was calculated by
multiplying the time spent in each job by the mean intensity estimate from the job exposure
matrix. The exposure-response analysis of cumulative exposure to PFOA was used to estimate
relative mortality risks. There were 773 total deaths in employees and 239 of the deaths were
from IHD. Males accounted for 235 of the 239 deaths from IHD. Mortality from IHD was not
found to be associated with PFOA exposure. These results were similar to those reported by
Leonard et al. (2008) in which a non-significant elevation for IHD mortality was observed
relative to expected deaths for a regional DuPont employee population.
Lundin et al. (2009) also investigated occupational exposure to PFOA and death from
IHD and other vascular diseases in workers at the Cottage Grove manufacturing plant as reported
by Gilliland and Mandel (1993). Cerebrovascular disease, diabetes, and ischemic heart disease
cases were included as the causes of deaths of 3,993 workers. A total of 807 workers died in the
follow-up period. The cohort differed from Gilliland and Mandel (1993) in that the period of
enrolment was extended to 1997 versus 1983, later follow-up (2002 versus 1989), and the
inclusion criteria were modified. Standardized mortality ratios (SMRs) and 95% confidence
intervals (CIs) were calculated. Mortality from cerebrovascular disease and ischemic heart
disease in PFOA workers was not elevated compared to the general population of Minnesota.
Within the cohort, the risk for cerebrovascular disease, SMR=1.6, 95% CI 0.5-3.7, was elevated
in workers who had ever worked a job in which they were exposed to PFOA. The authors found
diabetes mortality to be associated with moderate PFOA exposure (Hazard Ratio 3.7, 95% CI:
1.4-10.1) among subjects whose cause of death had been identified as diabetes. No deaths from
diabetes had occurred among workers with high exposure, based on job classification.
The studies discussed above did not uniformly monitor the same endpoints. Table 4-1
provides a summary of the results to assist the reader in integrating the findings of the studies.
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TABLE 4-1. Association of Serum PFOA with Serum Lipids and Uric Acid in Studies of Occupational
Populations
Study
Olsenetal.,2000
Olsenetal.,2001a, 2003
Olsenetal., 200 Ib, 2003
Sakretal.,2007a
Sakretal.,2007b
Olsen and Zobel, 2007
Costa et al., 2009
TC
<— >
t
t
t
t
<— >
t
VLDL
NM
NM
NM
t
NM
NM
NM
LDL
<— >
NM
NM
t
<— >
<— >
NM
HDL
<— >
^
^
^H-
<— >
1
~
TG
<— >
t
t
<— >
<— >
t
<— >
UA
NM
NM
NM
t
NM
NM
t
t= positive association; j= negative association; <->= no association; TC= total cholesterol; VLDL=very low density lipoprotein;
LDL= low density lipoprotein; HDL=high density lipoprotein; TG= triglycerides; UA=uric acid; NM= not monitored
General Population Studies
Several epidemiology studies examined serum lipids, lipoproteins, and/or uric acid
among cohorts from the general population. Three studies focused on populations serviced by
water districts contaminated by the Washington Works PFOA production plant in Ohio and West
Virginia, and two utilized data from the NHANES program.
Emmett et al. (2006) examined the association of serum PFOA concentration with serum
total cholesterol in residents of the Little Hocking water district in Ohio. The study population
(n=371, 2->60 years of age) was a random sample of the population served by the Little Hocking
water service. The subjects completed questionnaires (demographic, occupational, health
conditions, etc) and provided blood samples. PFOA concentration was determined by high
performance liquid chromatography tandem mass spectrometry. Regression models were used to
analyze the data. The median serum PFOA concentration was 354 ng/mL. No association was
observed between serum PFOA and total cholesterol.
Steenland et al. (2009) examined the association of PFOA with serum lipids in adult
participants of the C8 Health Project (n=46,294; 18 to >80 years). Serum samples were
separated into deciles or quartiles for analysis. Total cholesterol, HDL, triglycerides, LDL, and
non-HDL (total cholesterol minus HDL cholesterol) were measured or calculated from blood
samples. The data were analyzed by linear regression using the log-transformed values for all
variables. Covariates of the model included gender, quartile body mass index, education,
smoking, regular exercise, and alcohol consumption. A logistic model was used to analyze high
cholesterol and serum PFOA concentration (quartiles). The mean serum PFOA concentration
was 80 ng/mL.
All lipid outcomes, except for HDL showed significant increasing trends with increasing
serum PFOA decile. There was a positive association between mean levels of serum PFOA and
total cholesterol, LDL cholesterol, triglycerides, total cholesterol/HDL ratio, and non-HDL. The
predicted increase in total cholesterol from lowest to highest serum PFOA concentration decile
was 11-12 mg/dL. No association between mean level of serum PFOA and HDL cholesterol was
observed.
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The odds ratio for high cholesterol (>240 mg/dL) increased from the lowest to the highest
quartile of serum PFOA concentrations: 1.00, 1.21 (95% CI: 1.12-1.31), 1.33 (95% CI: 1.23-
1.43), and 1.40 (95% CI: 1.29-1.51). The results of the study were consistent with occupational
studies that found a positive association between PFOA exposure and serum lipids. Data
interpretation was limited by the lack of cumulative PFOA exposure data for the individual
subjects and by the cross-sectional study design which prevented determination of whether an
increase in PFOA serum levels corresponded with an increase in cholesterol levels.
Frisbee et al. (2010) examined the association of PFOA with serum lipids in children
(n=6536; 1-11.9 years) and adolescent (n=5934; 12.0-17.9 years) participants of the C8 Health
Project. Total cholesterol, HDL, LDL, and triglycerides were measured from serum samples.
The data were analyzed by linear regression using the natural log-transformed values for all
variables or by analysis of covariance. For some analyses, data were grouped into quintiles with
age-group and sex-specific groupings. Covariates of the model included age, gender, body mass
index z score, exercise, and length of fast. Triglyceride values were used only if fast had lasted 6
hours or more. The mean serum PFOA concentration for children was 77.7 ng/mL (82.1 ng/mL
for males; 73.1 ng/mL for females) and for adolescents, the mean was 61.8 ng/mL (69.3 ng/mL
for males; 53.7 ng/mL for females).
Total cholesterol, LDL, and triglycerides were positively associated (p<0.02) with serum
PFOA concentration after covariable adjustment. Assessment of the quintile trends showed
significant differences (p<0.02) between the first and fifth quintile for total cholesterol and LDL
for children and adolescents of both sexes combined and separated. A significant difference
(p=0.04) was observed for fasting triglycerides in female children only. An increased risk of
abnormal total cholesterol and LDL were positively associated with serum PFOA. The odds
ratios were 1.0 first (reference), 1.1 (95% CI: 1.0-1.3, second), 1.2 (95% CI: 1.0-1.4, third), and
1.2 (95% CI: 1.1-1.4, fourth and fifth) for total cholesterol, and 1.0 (reference, first), 1.2 (95%
CI: 1.0-1.5, second), 1.2 (95% CI: 1.0-1.4, third and fourth), and 1.4 (95% CI: 1.2-1.7, fifth) for
LDL. An increased risk of abnormal fasting triglyceride and HDL was not associated with
serum PFOA.
Steenland et al. (2010) examined the association of serum PFOA concentrations with uric
acid levels in adult subjects (n=53,458; 20 to >80 years) participating in the C8 Health Project
from 2005-2006. The reference range for uric acid is 2.0 to 8.5 mg/dL. Serum samples were
separated into deciles or quintiles for analysis. The data were analyzed by linear and logistic
regression with uric acid as the outcome and PFOA as the exposure variable. Covariates of the
model included age, gender, body mass index, education, smoking, alcohol consumption, and
serum creatinine. The mean serum PFOA concentration was 86.4 ng/mL.
The mean uric acid level was 5.58 mg/dL with an interquartile range of 4.5-6.6 mg/dL.
The increase in uric acid from lowest to highest serum PFOA concentration decile was 0.2-0.3
mg/dL. The odds ratio for high serum uric acid levels increased from the lowest to the highest
quintile of PFOA serum concentrations: 1.00, 1.33 (95% CI: 1.24-1.43), 1.35 (95% CI: 1.26-
1.45), 1.47 (95% CI: 1.37-1.58), and 1.47 (95% CI: 1.37-1.58). The study showed that higher
serum PFOA concentrations were associated with higher incidence of high serum uric acid
levels. As was the case for the Steenland et al. (2009a) report, data interpretation was limited by
the lack of cumulative PFOA exposure data and by the cross-sectional study design.
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Lin et al. (2009) investigated the association between serum PFOA, HDL cholesterol, and
triglycerides in adolescents (12-20 years) and adults (>20 years) of the United States by
analyzing data from the 1999-2000 and 2003-2004 National Health and Nutrition Examination
Survey (NHANES) study. The study population included 1443 subjects (474 adolescents, 969
adults) at least 12 years of age who had a morning examination and triglyceride measurement.
There were 266 male and 208 female adolescents and 475 male and 493 female adults. Multiple
linear regression and logistic regression models were used to analyze the data. Covariates
included age, sex, race, smoking status, alcohol intake, and household income. Log-transformed
PFOA concentration was 1.51 and 1.48 ng/mL for adolescents and adults, respectively. Serum
PFOA concentration was not associated with HDL or triglycerides. Data interpretation was
limited by the cross-sectional study design which did not permit causal inference.
Nelson et al. (2010) examined the relationship between polyfluoroalkyl chemical serum
concentration, including PFOA, and lipid and weight outcomes in the general population of the
United States by analyzing data from the 2003-2004 NHANES study. The population (n=1445;
773 male, 672 female) included persons between the ages of 12 years and 80 years with no
missing covariate information who were not pregnant, breast-feeding, taking insulin or
cholesterol medicine, or undergoing dialysis. Cholesterol (total, HDL, LDL, VLDL) was
measured from serum samples. Weight, height, and waist circumference were used in
consideration of body size outcome. Data for covariates predicting cholesterol and body weight
including age, sex, race/ethnicity, socioeconomic status, saturated fat intake, exercise, alcohol
consumption at > 20 years of age, smoking, and parity were obtained from the questionnaires.
Regression analyses were performed for sex and the age groups 12-19 years, 20-59 years, and
60-80 years.
A positive association was found between total (TC) and non-HDL (TC-HDL, -70-80%
TC) cholesterol and serum PFOA (9.8; 95% CI, -0.2 to 19.7). No association was found between
serum PFOA concentration and HDL, LDL, or VLDL. No association was found between serum
PFOA concentration and body weight. Data interpretation was limited by the cross-sectional
study design and lack of multiple cholesterol measurements.
Table 4-2 provides a summary of the results from the general population studies using the
same format as Table 4-1.
TABLE 4-2. Association of Serum PFOA with Serum Lipids and Uric Acid in Studies of General
Populations
Study
Emmett et al., 2006
Steenlandetal.,2009a,b
Frisbee et al., 2010
Lin et al., 2009
Nelson etal., 2010
TC
<— >
t
t
NM
t
VLDL
NM
NM
NM
NM
<— >
LDL
NM
t
t
NM
<— >
HDL
NM
<— >
<— >
<— >
<— >
Non-HDL
NM
t
NM
NM
t
TG
NM
t
<— >
<— >
NM
UA
NM
t
NM
NM
NM
1= positive association; |= negative association; <-»= no association; TC= total cholesterol; VLDL=very low density lipoprotein;
LDL= low density lipoprotein; non-HDL= TC(VLDLJDL, LDL)-HDL; HDL=high density lipoprotein; TG= triglycerides;
UA=uric acid; NM= not monitored
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Clinical Chemistries (Biochemical and Hematological)
Occupational Population Studies
Many of the studies that investigated alterations in serum lipids also looked at alterations
in clinical chemistry and hematological parameters. The results that apply to occupational
cohorts are described below.
Olsen et al. (2000; 2001a,b; 2003) examined alkaline phosphatase (ALP), gamma-
glutamyl transpeptidase (GOT), aspartate aminotransferase (AST), alanine transaminase (ALT),
total and direct bilirubin, renal enzymes, blood glucose, and hematology in the workers of PFOA
production plants as described above. No association was observed between serum PFOA
concentration and the parameters explored in the workers in 1993, 1995, 1997, or 2000 by cross-
sectional or longitudinal analyses.
Sakr et al. (2007a) examined the relationship between serum PFOA and several clinical
chemistry parameters. A complete blood count, metabolic panel (glucose, BUN, creatinine, iron,
uric acid, electrolytes, creatinine kinase, lactic dehydrogenase (LDH), alkaline phosphatase
(ALP), protein, albumin, C-reactive protein), liver enzyme panel (AST, ALT, GOT, bilirubin),
and serum PFOA concentration were determined from the blood samples. Serum PFOA was
positively associated (p<0.05) with GGT in all of the participating workers. An association
between serum PFOA concentration and iron, LDH, calcium, and potassium was observed;
however, the authors concluded that the associations were unlikely to have any clinical
significance attributable to PFOA because of the small magnitude of the effect based on the
parameter estimates, and the expectation of finding statistically significant associations by
chance alone due to the number of comparisons performed.
Sakr et al. (2007b) also conducted a longitudinal study of liver enzymes in workers of the
Washington Works plant described previouly. Hepatic clinical chemistry (GGT, AST, ALT,
ALP, total bilirubin), height, and weight data were analyzed. Serum PFOA concentration was
positively associated with AST (p=0.009) and negatively associated with total bilirubin
(p=0.006) after adjustment for age, BMI, gender, and decade of hire in the model. No
association was observed between serum PFOA concentration and GGT, ALT, and ALP. A
limitation of the study was the lack of alcohol consumption data for the workers which could
lead to confounding of the liver enzyme results.
In addition to serum lipids, Olsen and Zobel (2007) examined the association between
serum PFOA concentration and liver enzymes in fluorochemical workers. Serum samples were
analyzed for glucose, ALP, GGT, AST, ALT, and bilirubin concentrations. Serum PFOA was
not associated with ALP or ALT. A negative association between total bilirubin and serum
PFOA concentration (p<0.05) was observed at all sites combined. At the Decatur site, serum
PFOA was positively associated with ALP, ALT, and GGT, and negatively associated with total
bilirubin (p<0.05). The authors suggested the associations may be due to demographics and
lifestyle rather than due to serum PFOA concentrations.
Costa et al. (2009) also examined correlations between serum PFOA concentration and
hematological parameters, a2 globulins, immunoglobulins (IgA, IgG, IgM), glucose, liver
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enzymes, and kidney enzymes in workers of a PFOA production plant in Italy. Serum PFOA
concentration was positively correlated with ALT, GOT, ALP, and a2 globulins (p<0.05), and
negatively correlated with total bilirubin (p<0.01) in 56 workers assessed concurrently over the
last 7 years. Body mass index was also correlated with ALT and ALP suggesting a possible
interaction. The authors hypothesized that PFOA may interfere with metabolism, but were
hesitant to draw firm conclusions due to the small number of subjects. No associations were
observed between serum PFOA concentration and hematology, immunoglobulins, glucose, and
kidney enzymes.
The occupational epidemiology studies that examined standard biochemical and
hematological endpoints looked for elevated levels of one or more enzyme indicative of effects
on the liver, but did not always examine the same enzymes. Two of the studies did not report on
renal enzymes, blood glucose levels, or hematology. Table 4-3 summarizes the observations
from the studies described above.
TABLE 4-3. Associations of Serum PFOA with Serum Clinical Biochemistry and Hematology Measures
Study
Olsenetal.,2000,
2001a,b; 2003
Sakretal, 2007a
Sakretal, 2007b
Olsen and Zobel, 2007
Costa et al., 2009
Liver Enzymes
<-»•
|GGT
IAST
IGGT, ALP, ALT (Decatur)
IGGT, ALP, ALT
Bilirubin
<-»•
<-»•
1
1
1
Renal Enzymes
<-»•
<-»•
NM
NM
<— >
Glucose
<— >
<— >
NM
NM
<— >
Hematology
<-»•
<-»•
NM
NM
<— >
t= positive association; J,= negative association; <->= no association The entries for liver enzymes identify only
those that were observed to have a positive association with serum PFOA. NM = Not Monitored
General Population Studies
Emmett et al. (2006) found no association between serum PFOA and liver enzymes, renal
enzymes, and hematologic parameters in 371 residents of the Little Hocking water district in
Ohio.
Lin et al. (2010) investigated the association between low-dose serum PFOA and liver
enzymes in the adult population of the United States by analyzing data from the 1999-2000 and
2003-2004 NHANES study. The study population included 2,216 adults (1076 men, 1140
women) older than 20 years who were not pregnant or nursing, had fasted more than 6 hours at
the time of examination, were negative for hepatitis B or C virus, had body weight, height,
educational attainment, and smoking status data available, and had serum tests for PFCs, liver
function, or the five components of metabolic syndrome. Regression models were used to
analyze the data and adjust for confounders.
Serum PFOA concentration was divided into quartiles (Ql: < 2.90; Q2: < 4.20; Q3: <
5.95; Q4: > 5.95 ng/ml). Unadjusted liver enzymes, serum ALT and log-GGT increased across
quartiles of PFOA (p < 0.012) but total bilirubin showed no trend. The linear regression models
were adjusted for:
• age, gender, and race/ethnicity
• age, gender, race/ethnicity, lifestyle (smoking status, drinking status, education level),
• biomarker data (BMI, metabolic syndrome, iron saturation status, insulin resistance).
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A positive association was found between serum PFOA concentration and serum ALT and log-
GGT. One unit increase in serum log-PFOA was associated with an increase of 1.86 units of
serum ALT and a 0.08 unit increase in log-GGT. An association between serum PFOA
concentration and serum ALT was significant in the following categories: non-Hispanic
Caucasians, lower education level, higher BMI, non-smoking, lower alcohol consumption,
higher insulin resistance, and those with metabolic syndrome. An association between serum
PFOA concentration and GGT was significant in the following categories: non-Hispanic
Caucasians, higher BMI, lower alcohol consumption, and higher insulin resistance. The ALT
level trend across serum PFOA quartiles was significant (p=0.003) in adjusted models with
subjects having a BMI > 30 kg/m2. Data interpretation were limited by the cross-sectional study
design, and the fact that other environmental chemicals (possible covariates or explanatory
variables) were not included in the study, medication use was not included, and liver tissue status
was unknown.
Similar to the above study, Gallo et al. (2012) investigated the correlation between serum
PFOA levels and liver enzymes in a total of 47,092 samples collected from members enrolled in
the C8 Health Project. The association of ALT, GGT, and direct bilirubin with PFOA was
assessed using linear regression models adjusted for various confounders. The In-transformed
values of ALT were significantly associated with In-PFOA and showed a steady increase in fitted
levels of ALT per decile of PFOA, leveling off after approximately 30 ng PFOA/mL. Fitted
values of GGT by deciles of PFOA showed a slight positive trend when adjusted for insulin
resistence and BMI, but this was not confirmed in the logistic model analysis. No association
was seen with direct bilirubin and PFOA levels. Limitations of the study include the cross-
sectional design and self-reported lifestyle characteristics. Only a small number of ALT values
were outside the normal range making the results difficult to interpret in terms of health.
Only the data from Lin et al. (2010) and Gallo et al. (2012) provide any suggestion of
effects on the liver at levels of serum PFOA observed in the general population. A significant
association with ALT and GGT in some population groups and in individuals with a higher BMI
was observed in the US adult population.
Although there were no changes in kidney enzymes in the studies discussed of worker
populations above, there is the potential for an impact of PFOA on the on the function of tubular
resorption as a result its utilization of tubular transporters for excretions (See Section 3-4).
Shankar et al. (2011) used data from the NHANES study to determine whether there was a
relationship between serum PFOA levels and chronic kidney disease. A total of 4,587 adult
participants (51.1% women) with PFOA measurements available from the 1999-2000 and 2003-
2008 cycles of the survey were examined. Chronic kidney disease was defined as glomerular
filtration rate of less than 60 mL/minute/1.73 m2. Serum PFOA levels were categorized into
quartiles: quartile 1 = <2.8 ng/mL; quartile 2 = 2.8-4.1 ng/mL; quartile 3 = 4.2-5.9 ng/mL;
quartile 4 = >5.9 ng/mL. The multivariable odds ratio for chronic kidney disease for individuals
in quartile 4 was 1.73 (95% CI: 1.04, 2.88; p for trend = 0.015) compared with individuals in
quartile 1. This association was shown to be independent for confounders of age, sex,
race/ethnicity, body mass index, diabetes, hypertension, and serum cholesterol level. However,
the authors noted that because of the cross-sectional nature of the study, the possibility of reverse
causality could not be excluded. A low glomerular filtration rate would diminish the removal of
PFOA from serum for excretion by the kidney, thus increasing the serum PFOA levels. Had
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glomerular filtration rate been the variable, those in the highest PFOA quartiles could have had
been those in the quartile with the lowest glomerular filtrations rates.
Diabetes and Related Endpoints
General Population Studies
MacNeil et al. (2009) examined the association of PFOA with type II diabetes in adult
participants of the C8 Health Project (n=54,468; 20 to >80 years). Serum PFOA concentration
was divided into deciles using the population distribution. Serum PFOA (deciles), body mass
index, gender, family history of diabetes, race, use of cholesterol-lowering medicine, and use of
blood pressure-lowering medicine were used to analyze the data in categorical and logistic
regression models for the outcome of type II diabetes. Serum fasting glucose levels were the
focus for a linear regression analysis of the study population (n=21,643) excluding type II
diabetics and those who had provided non-fasting blood samples. The mean serum PFOA
concentration for the entire study population was 86.8 ng/mL and 91.3 ng/mL for subjects with
type II diabetes validated by medical review (n=3539).
There was no association was between serum PFOA concentration and fasting serum
glucose level in participants who were not characterized as diabetics. The mean serum PFOA
concentration in subjects who self-reported type II diabetes (n=4278) was 92.9 ng/mL and 122.7
ng/mL in subjects diagnosed in the last 10 years (n=1055). No association was observed
between type II diabetes and serum PFOA concentration. The odds ratio by decile was 1.00,
0.71, 0.60, 0.72, 0.65, 0.65, 0.87, 0.58, 0.62, and 0.72. The results of the analysis indicated that
PFOA exposure is not associated with type II diabetes among population studied. Data
interpretation was limited by the cross-sectional study design which made it difficult to
determine if PFOA exposure preceded disease.
Metabolic syndrome is a combination of medical disorders and risk factors that increase
the risk of developing cardiovascular disease and diabetes. Lin et al. (2009) investigated the
association between serum PFOA and glucose homeostasis and metabolic syndrome in
adolescents (12-20 years) and adults (>20 years) as described previously. The National
Cholesterol Education Program Adult Treatment Panel III guidelines were used to define adult
metabolic syndrome and the modified guidelines were used to define adolescent metabolic
syndrome. In adults, serum PFOA concentration was associated with increased 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.
Nelson et al. (2010) examined the relationship between polyfluoroalkyl chemical serum
concentration, including PFOA, and insulin resistance as previously described. Fasting insulin
and fasting glucose were used to determine the homeostatic model assessment for insulin
resistance. No association was found between serum PFOA concentration and insulin resistance.
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4.1.1.2 Thyroid Effects
Occupational Population Studies
Thyroid hormones were monitored in the medical surveillance programs at the various
PFOA manufacturing plants and an association with serum PFOA concentration was examined
in several occupational studies.
Serum PFOA levels were obtained from volunteer workers of the Cottage Grove,
Minnesota PFOA plant in 1993 (n = 111) and 1995 (n = 80) as part of the medical surveillance
program (Olsen et al., 1998) and analyzed to determine a relationship between thyroid
stimulating hormone (TSH) and PFOA concentration. Employees were placed into 4 exposure
categories based on their serum PFOA levels: 0-1 ppm, 1- < 10 ppm, 10- < 30 ppm, and >30
ppm. Statistical methods used to compare PFOA levels and hormone values included:
multivariable regression analysis, ANOVA, and Pearson correlation coefficients. TSH was
significantly (p = 0.002) elevated in 10-<30 ppm exposure category for 1995 only (mean serum
TSH level was 2.9 ppm). However, mean TSH levels for the other exposure categories, including
the >30 ppm category, were all the same (1.7 ppm). In 1993, TSH was elevated in this same
exposure category, but was not statistically significant (p = 0.09) when compared to the other
exposure categories.
Olsen et al. (2001a, 2003) and Sakr et al. (2007a) found no association between serum
PFOA concentration and TSH, T4, and T3 levels in the Decatur, Antwerp, and Washington
Works plant employees. Costa et al. (2009) also found no association between serum PFOA
concentration and thyroid hormones levels in workers at a PFOA production plant in Italy.
Olsen and Zobel (2007) looked at the relationship between serum PFOA concentration
and TSH, serum and free T4, and T3 levels in workers at the Decatur, Antwerp, and Cottage
Grove production plants as described previously. No association between TSH, serum T4, and
PFOA concentration was observed. A negative association (p<0.01) between free T4 and serum
PFOA concentration was observed in the unadjusted and adjusted (age, BMI, and alcohol)
models for all locations combined; no association was observed for the individual locations. A
positive association (p<0.05) was observed between T3 and serum PFOA concentration in the
unadjusted and adjusted models for all locations combined, the Antwerp plant, and the Decatur
plant. The authors noted that the results were not considered clinically relevant because the
results were within normal reference range.
General Population Studies
Emmett et al. (2006) examined the association of serum PFOA with thyroid disease in
residents of the Little Hocking, Ohio, water district as described previously. No association was
observed between serum PFOA and thyroid disease. Serum PFOA was decreased (not
significantly different) in subjects with thyroid disease (387 ng/mL) compared to subjects
without thyroid disease (451 ng/mL).
Pirali et al. (2009) measured intrathyroidal levels of PFOA in thyroid surgical specimens
to determine if a relationship existed between PFOA and the clinical, biochemical, and
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histological phenotype of thyroid disease patients. Serum PFOA concentration was also
measured to determine if a relationship existed between thyroid tissue and serum PFOA levels.
Patients (n=28; 8 males, 20 females; 33-79 years) with benign multinodular goiters (n=15),
Graves' disease (n=7), malignant papillary carcinoma (n=5), and malignant follicular carcinoma
(n=l) 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 C/-test, Pearson and
Spearman's correlation tests, and Chi2 test with Fisher's correction were used to compare group
results. Regression analysis was used to test the effect of different variables independently of a
covariate.
The patients were divided into three different groups: Group I (toxic and nontoxic
multinodular goiter, n=12), Group II (differentiated thyroid cancer, n=6), and Group III
(Hashimoto's thyroiditis or Graves' disease, n=10). Thyroid PFOA concentration for the control
group, Group I, Group II, and Group III ranged from 1.0-6.0, 0.4-4.4, 1.4-4.0, and 1.0-4.6 ng/g,
respectively. Serum PFOA concentration for the control group, Group I, Group II, and Group III
ranged from 4-13.7, 1.2-16.6, 5.1-9.6, and 3.9-12.5 ng/ml, respectively. The concentration of
PFOA in the thyroid and serum was similar between control and thyroid patients at the time of
measurement. Age, sex, residence, working activity, malignant/nonmalignant conditions,
antibodies, thyroid hormone concentrations, and ultrasound parameters were not associated with
thyroid or serum PFOA concentration. There was also no correlation between serum and thyroid
PFOA concentration.
Bloom et al. (2010) investigated the associations between serum PFCs, including PFOA,
and thyroid stimulating hormone (TSH) and free thyroxine (T4). The serum samples came from
31 participants (27 males, 4 females; mean age 39 years) of the 1995-1997 New York State
Angler Cohort Study Dioxin Exposure Substudy. The study subjects completed a questionnaire
and provided a blood sample for serum analysis. The questionnaire contained questions about
sport-fish and game consumption, lifestyle, demographic factors, and medical history. The
serum samples were analyzed for TSH and free T4 in 2003 by immunometric chemiluminescent
sandwich assay and for PFCs in 2006 by ion pair extraction high-performance chromatography
with electrospray tandem mass spectrometry. Regression models were used to analyze the data
and adjust for confounders.
No subjects reported use of thyroid medication or physician-diagnosed goiter or thyroid
conditions. Mean TSH concentration (range 0.43-15.70 |iIU/mL) was within normal range
(0.40-5.00 |iIU/mL) with the exception of one subject. Mean free T4 (0.90-1.55 ng/dL) was
within normal range (0.80-1.80 ng/dL) for all subjects. The mean serum PFOA concentration
was 1.33 ng/mL and ranged from 0.57 to 2.58 ng/mL. The males had a significantly higher
serum PFOA concentration compared to females (1.47 ng/mL vs. 1.05 ng/mL; p = 0.047). There
was no association between serum PFOA concentration and TSH or free T4.
Melzer et al. (2010) examined the association between serum PFOA concentration and
thyroid disease in the general population of the United States by analyzing data from the 1999-
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2000, 2003-2004, and 2005-2006 NHANES study. The population included 3,966 adults (2066
women, 1900 men) older than 18 years. The participants answered a questionnaire, had a
physical examination, and provided blood and urine samples for analysis. Serum samples were
analyzed for PFOA concentration by solid-phase extraction coupled to isotope dilution/high-
performance liquid chromatography/ tandem mass spectrometry. Data on diseases diagnosed by
a physician and confounding factors including year of NHANES study, age, gender,
race/ethnicity, education, smoking status, BMI, and alcohol consumption were obtained from the
questionnaire. Regression models were used to analyze the data and adjust for confounders.
Serum PFOA concentration was divided into quartiles for each sex. In women, serum
PFOA concentration ranged from 0.1-123.0 ng/ml (Ql: 0.1-2.6; Q2: 2.7-4.0; Q3: 4.1-5.7; Q4:
5.7-123.0), and in men, serum PFOA concentration ranged from 0.1-45.9 ng/ml (Ql: 0.1-3.6;
Q2: 3.7-5.2; Q3: 5.3-7.2; Q4: 7.3-45.9). Women in PFOA quartile 4 were more likely to report
current thyroid disease [odds ratio (OR) = 2.24, 95% CI: 1.38-3.65, p=0.002] compared to
women in quartiles 1 and 2. No association between serum PFOA concentration and thyroid
disease was observed in men. Data interpretation was limited by the cross-sectional study
design, lack of specific thyroid disease diagnosis, and single serum samples for PFOA
measurements taken at the same time as disease status.
Chan et al. (2011) examined the association between hypothyroxinemia and serum PFOA
concentration in pregnant Canadian women (n=>271; 20.1-45.1 years of age, >22 weeks of
gestation) in a matched case-control study. Maternal serum from the second trimester was
collected between Dec 15, 2005 and June 22, 2006 as part of an elective prenatal screen for birth
defects. Serum samples were anlayzed for TSH and free T4 concentrations and PFOA. The
cases (n=96) had normal TSH concentrations and free T4 concentrations in the lowest 10th
percentile (<8.8 pmol/L). The controls (n=175) had normal TSH concentrations and free T4
concentrations between the 50th and 90th percentiles (12-14.1 pmol/L). Maternal age, weight,
and gestational age at blood draw, dichotomized at 50th percentiles were included as
confounders, and race was included at a covariate. Chi-square tests and regression models were
used to analyze the data. The geometric mean serum PFOA concentration in the cases was 3.10
nmol/L and 3.32 nmol/L in the controls. There was no association between serum PFOA
concentration and hypothyroxinemia in pregnant women. The authors noted that hormone levels
change over pregnancy and time of blood could have influenced the results. They also noted that
the population may have been skewed because only women who chose to be screened were
included.
There are no strong data linking serum PFOA to changes in thyroid hormones or thyroid
disorder and indicated by the Table 4-4 compilation of the studies in the occulational and general
population.
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TABLE 4-4. Association of serum PFOA with the prevalence of thyroid disease and thyroid hormone levels
in studies of general and worker populations
Study
Bloom etal. (2010)
Chan etal. (2011)
Costa et al. (2009)
Emmett et al. (2006)
Melzer etal. (2010)
Olsen etal. (1998)
Olsen etal. (200 la, 2003)
Olsen and Zobel (2007)
Pirali et al. (2009)
Sakr et al. (2007a)
Population
General
General
Occupational
General
General
Occupational
Occupational
Occupational
General
Occupational
Thyroid Disease
<— >
<—>•
NM
<— >
<->men
t women
NM
NM
NM
4-^>
NM
TSH
4-^>
4-^>
4-^>
NM
NM
t
4-^>
4-^>
NM
4-^>
T3
NM
NM
4-^>
NM
NM
NM
4-^>
t
NM
4-^>
T4
4-^>
4-^>
4-^>
NM
NM
NM
<— >
<-> serum
| free
NM
4-^>
t= positive association; J,= negative association; <->= no association; NM = Not Monitored
4.1.1.3
Steroid Hormones
Occupational Population Studies
Olsen et al. (1998) examined several hormones including cortisol, estradiol, follicle
stimulating hormone, dehydroepiandrosterone sulfate, 17 gamma-hydroxyprogesterone (a
testosterone precursor), free testosterone, total testosterone, luteinizing hormone, prolactin, and
sex hormone-binding globulin in male workers at the Cottage Grove, Minnesota, production
plant for 1993 and 1995. No association between serum PFOA and any hormone was observed,
but some interesting trends were noticed. When the mean measures of the various hormones
were compared by exposure categories, there was a statistically significant (p = 0.01) elevation in
prolactin in 1993 only for the 10 workers whose serum PFOA levels were between 10 and 30
ppm compared to the lower 2 exposure categories.
Estradiol levels in the >30 ppm PFOA group in both years were 10% higher than the
other PFOA groups, but the difference was not statistically significant. These results were
confounded by estradiol being correlated with BMI (r = 0.41, p < 0.001 in 1993, and r = 0.30, p
< 0.01 in 1995). The authors postulate that the study may not have been sensitive enough to
detect an association between PFOA and estradiol because measured serum PFOA levels were
likely below the observable effect levels suggested in animal studies (55 ppm PFOA in the CD
rat). Only 3 employees in this study had PFOA serum levels this high. They also suggest that the
higher estradiol levels in the highest exposure category could suggest a threshold relationship
between PFOA and estradiol.
In the Sakr et al. (2007a) study, an association was observed between serum PFOA and
serum estradiol (p=0.017) and testosterone (p=0.034) in male workers of the Washington Works
plant; however, circadian variations of hormones were not taken into consideration during
analysis. The biological signifificance of the results is unknown.
Costa et al. (2009) found no association between serum PFOA concentration and
estradiol or testosterone in workers at a PFOA production plant from 2000 to 2007.
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General Population Studies
Knox et al. (2011) examined the endocrine disrupting effects of perfluorocarbons in
women from the C8 Health by analyzing the relationship between serum PFOA, serum estradiol
concentration, and menopause onset. The population included women over age 18 years
(n=25,957). Serum PFOA and estradiol concentrations were determined from blood samples.
Women who were pregnant, had full hysterectomies, and taking any prescription hormones or
selective estrogen receptor modulators and fertility agents were excluded from estradiol analysis.
Serum PFOA concentration was grouped into quintiles (natural log transformation): 1=0.25-
11.2; 2=11.3-19.8; 3=19.9-36.7; 4=36.8-84.9; and 5=85.0-22412.0 ng/mL. Estradiol analysis
was calculated by age group, 18-42 years, >42 < 51 years, and >51 < 65 years. Menopause was
determined by questionnaire. Menopause analysis was calculated by age group, 30-42 years,
>42 < 51 years, and >51 < 65 years, and excluded those who reported hysterectomies. Logistic
regression models were adjusted for smoking, age, BMI, alcohol consumption, and regular
exercise. PFOA concentration in women who had a hysterectomy was significantly higher than
women who had not had a hysterectomy. Serum PFOA and estradiol concentrations were not
associated. Menopause analysis in the oldest group of women showed that all quintiles were
significantly higher than the lowest, and in women between the ages of 42 and 51 years, quintiles
3-5 were significantly higher than the lowest. The authors suggested that PFOA may be
associated with endocrine disruption in women. Data interpretation was limited by the cross-
sectional study design and survey-reported menopause without age or independent confirmation.
7.1.1.4. Immunotoxicity
As described previously, no association between serum PFOA concentrations and
immunoglobin levels was found in male workers (Costa et al., 2009).
Fei et al. (201 Ob) examined the association between prenatal PFOA exposure and
hospitalizations for infectious diseases in early childhood. A total of 577 hospitalizations
occurred with 363 being due to infectious disease. A lower risk of hospitalization due to
infections was associated with higher maternal PFOA concentration. The incidence rate ratio
(IRR) was significantly different only for the second quartile of exposure (IRR=0.71, 95% CI
0.53-0.94) compared to the first. A slightly higher risk for hospitalizations was observed in girls
with higher maternal PFOA concentrations (IRR= 1.00, 1.20, 1.63, 1.74 for first, second, third,
and fourth quartiles, respectively). The risk for boys was below 1.0 for each quartile. Overall,
there was no association between hospitalizations due to infectious diseases and maternal PFOA
exposure.
Okada et al. (2012) investigated the relationship between maternal PFOA concentration
and infant allergies and infectious diseases during the first 18 months of life as well as cord
blood IgE levels. The prospective birth cohort was based on infants delivered at the Sapporo
Toho Hospital in Sapporo, Hokkaido, Japan between July 2002 and October 2005. PFOA levels
were measured in maternal serum taken after the second trimester (n=343) and total IgE
concentration was measured in cord blood (n=231) at the time of delivery. Infant allergies and
infectious diseases were assessed in a maternal self-administered questionnaire at 18 months
post-delivery. Polynomial regression analyses, adjusted for potential confounders, were
performed on log-transformed data. Mean maternal PFOA concentration was 1.4 ng/mL and
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cord blood IgE level was 0.62 lU/mL. Cord blood IgE level decreased significantly with high
maternal PFOA concentration among female infants, but not male infants. No significant
associations were observed between maternal PFOA levels incidence of food allergy, eczema,
wheezing, or otitis media in infants at 18 months of age. Limitations of the study include the
small sample size, potential selection bias of the population, and accurate diagnosis of disease in
the infants.
Antibody responses to diphtheria and tetanus toxoids following childhood vaccinations
were assessed in context of exposure to perfluorinated compounds (Grandjean et al., 2012). The
prospective study included a birth cohort of 587 singleton births during 1999-2001 from the
National Hospital in the Faroe Islands. Serum antibody concentrations were measured in
children at age 5 years prebooster, approximately 4 weeks after the booster, and at age 7 years.
Prenatal exposures to perfluorinated compounds were assessed by analysis of serum collected
from the mother during week 32 of pregnancy; postnatal exposure was assessed from serum
collected from the child at 5 years of age. Multiple regression analyses with covariate
adjustments were used to estimate the percent difference in specific antibody concentrations per
2-fold increase in PFOA concentration in both maternal and 5-year serum. Maternal PFOA
serum concentration was negatively associated with antidiphtheria antibody concentration (-
16.2%) at age 5 before booster. The biggest effect was found in comparison of antibody
concentrations at age 7 with serum PFOA concentrations at age 5 where a 2-fold increase in
PFOA was associated with differences of-36% (95% CI, -52% to -14%) and -25% (95% CI, -
43% to -2%) for tetanus and diphtheria, respectively. Additionally at age 7, a small percentage
of children had antibody concentrations below the clinically protective level of 0.1 lU/mL. The
odds ratios of antibody concentrations falling below this level were 4.20 (95% CI, 1.54 to 11.44)
for tetanus and 3.27 (95% CI, 1.43 to 7.51) for diphtheria when age 7 antibody levels were
correlated with age 5 PFOA serum concentrations.
The association between serum levels of perfluorinated compounds and childhood asthma
was investigated by Dong et al. (2013). The cross-sectional study included a total of 231
children aged 10-15 years with physician-diagnosed asthma and 225 age-matched non-asthmatic
controls. Between 2009 and 2010, asthmatic children were recruited from two hospitals in
Northern Tiawan, while the controls were part of a cohort population in seven public schools in
Northern Tiawan. Serum was collected for measurement often 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 asthenia severity
score (including frequency of attacks, use of medicine, and hospitilazation) during the previous
four weeks. Associations of perflourinated compound quartiles with concentrations of
immunological markers and asthma outcomes were estimated using multivariable regression
models. Nine often perfluorinated compounds were detectable in >84.4% of all children with
levels generally higher in asthmatic children compared with non-asthmatics. Serum
concentrations of PFOA in asthmatic and non-asthmatic children were 1.5±1.3 and 1.0±1.1
ng/mL, respectively; four other compounds were measured at higher concentrations with the
highest levels for PFOS and perfluorotetradecanoic acid. The adjusted odds ratios for asthma
association with the highest versus lowest quartile levels were significantly elevated for seven of
the compounds. For PFOA the odds ratio was 4.05 (95%CI: 2.21, 7.42). In asthmatic children,
absolute eosinophil counts, total IgE, and eosinophilic cationic protein concentration were
postitively associated with PFOA levels with a significant monotonic trend with increasing
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serum concentration. None of these biomarkers was significantly associated with PFOA levels
in non-asthematic children. Serum PFOA levels were not significantly associated with asthma
serverity scores among the children with asthma.
4.1.1.4 Reproductive & Developmental Endpoints
Several studies have examined the relationship between PFOA exposures and
reproductive, gestational, and developmental endpoints as well as postnatal growth and
maturation. The reproductive endpoints included sperm count (Joensen et al., 2009) and
fecundity (Fei et al., 2009). The gestational endpoints included gestational age (Nolan et al.,
2009), and fetal growth (Fei et al., 2007, 2008a; Washino et al., 2009). The developmental
endpoints included birth weight (Apelberg et al., 2007; Monroy et al., 2008; Nolan et al., 2009;
Stein, et al., 2009), miscarriage or pre-term birth (Stein, et al., 2009), birth defects (Stein et al.,
2009), and infant development during the first two years (Anderson et al., 2010; Fei et al., 2008b,
2010a,b). Postnatal growth and maturation included behavioural assessment (Fei and Olsen,
2011; Hoffman et al., 2010), onset of puberty (Lopez-Espinosa et al., 2011; Christensen et al.,
2011), and risk of adult obesity (Halldorsson et al., 2012). As a group, the studies do not suggest
an effect of serum PFOA at the levels examined on reproductive and developmental outcomes.
A few studies indicate an impact on anthropometric measures (e.g. body weight, birth length) but
the associations are not strong.
Andersen et al. (2010) and Fei et al. (2007, 2008a, b, 2009, 2010a,b) conducted a series
of studies examining the association of maternal plasma PFOA concentration and various
reproductive and developmental outcomes in Danish women and their offspring. The women
(n=1400) and their infants were randomly selected from the Danish National Birth Cohort, and
the study included those who provided their first blood samples between gestational weeks 4 and
14 and gave birth to a single live-born child without congenital malformation. The women
participated in telephone interviews (at 12 and 30 weeks gestation and when children were 6 and
18 months of age) and filled out a food frequency questionnaire. As the children aged, more
questionnaires were completed by the mothers with regard to behavioral health and motor
coordination. Highly structured questionnaires were used to gather information on possible
confounders including infant sex, maternal age, parity, socio-occupational status, prepregnancy
BMI, and smoking during pregnancy. The National Hospital Discharge Register was used to
obtain birth weight, gestational age, placental weight, birth length, head and abdominal
circumference data, Apgar scores based on heart rate, respiratory effort, reflex, irritability,
muscle tone, and skin color. Plasma PFOA concentration was determined from the first blood
samples of 1399 women, from the second blood samples of 200 women, and from cord blood
samples of 50 infants by solid phase extraction high performance liquid chromatography-tandem
mass spectrometry. PFOA concentrations were divided into quartiles (Fei et al., 2009, 2010a),
with the lowest quartile designated as the reference group, as follows: 6.97 ng/mL. Regression models were
used to analyze the data. Results of these studies are included in the discussion of results for
specific endpoints below.
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Reproductive Outcome
Fei et al. (2009) examined the association between plasma PFOA concentration and
longer time to pregnancy as a measure of fecundity in 1240 women. The time to pregnancy
(TTP) was categorized as follows: immediate pregnancy (<1 month), 1-2, 3-5, 6-12, and >12
months. Having >12 months TTP or having used fertility treatment to get pregnant were used to
define infertility. A total of 620 women had a TTP within the first 2 months of trying to get
pregnant and 379 had a TTP of >6 months with 188 of those women having a TTP of >12
months. The mean plasma PFOA concentration was 5.6 ng/mL for women who planned their
pregnancies, and for TTP <6 months, 6-12 months, and >12 months was 5.4, 6.0, and 6.3 ng/mL,
respectively. Plasma PFOA concentration was significantly greater (p<0.001) in women who
had TTP >6 months compared to those whose TTP was <6 months. The women with a TTP >6
months were more likely to be older, have middle socio-occupational status, and have a history
of spontaneous miscarriage or irregular menstrual cycles. The adjusted odds for infertility
increased 60-154% among women with >3.91 ng/mL plasma PFOA concentration compared to
women with <3.91 ng/mL plasma concentration. The fecundity odds ratio was 0.72, 0.73, and
0.60 for the three highest PFOA concentration quartiles. In the likelihood ratio test, the trend
was significant (p<0.001). Although the results of the study suggest that plasma PFOA
concentration may reduce fecundity, the authors noted that selection bias, the unknown quality of
the sperm, unknown frequency and timing of intercourse, and abnormal hormone levels may
have an impact on the results and fecundity.
Nolan et al. (2009, 2010) published two studies that examined birth outcome in infants
born to mothers whose partial or full public water source was from the Little Hocking Water
Association (LHWA) in Washington County, OH, or from the same geographic area but with
water not provided by the LHWA. The population serviced by LHWA had serum levels of
PFOA (6.78 |ig/L) that were approximately 50 times higher than serum levels of the general
population. Archival data for births, including birth weight, gestation age, plurality, neonatal
sex, race, mother's age, and mother's zip code, in Washington County, OH, from January 1,
2003 to September 1, 2005 were obtained for analysis. For mothers with public water service
provided completely, partially, or not at all by LHWA, the incidence of preterm births was
10.7%, 11.3%, and 13.4%, respectively (Nolan et al., 2009). The national incidence of preterm
birth is 12.7%. No differences were found between gestational age of infants born to mothers
whose partial or full water source was serviced by LHWA and infants born to mothers whose
water source was serviced by another provider. Exposure to high levels of PFOA through the
drinking water was not associated with preterm births.
Congenital anomalies were diagnosed in 1.8%, 1.9%, and 2.0% of the mothers with water
provided completely, partially, or not at all by LHWA, respectively (Nolan et al., 2010). When
adjusted for confounders, no statistically significant differences were found. Complications with
labor and delivery were observed in 32.5%, 35.9%, and 41.9% of the mothers with water
provided completely, partially, or not at all by LHWA, respectively. Mothers with water
provided by LHWA did have in increased likelihood of having dysfunctional labor, but the
reported cases were low in number. Mothers with one or more maternal risk factors were 37.5%,
34.4%, and 39.3% of the populations with water provided completely, partially, or not at all by
LHWA, respectively. Adjusted regression models showed no statistical differences across water
service. An increased likelihood (crude OR 11, 95% CI: 1.8-64) of anemia and dysfunctional
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labor (crude OR 5.3, 95% CI: 1.2-24) in mothers with water provided by LHWA was found, but
the reported cases were low in number. No association was found between PFOA and increased
incidence of congenital anomalies, most labor and delivery complications, and maternal risk
factors.
Stein et al. (2009) examined the association between PFOA serum levels and pregnancy
outcome in a population of women living in the Mid-Ohio Valley near Parksburg, WV from
2000 to 2006. The women were part of the C8 Health Project. Pregnancies were restricted to
those occurring in the 5 years prior to C8 Health Project enrollment (August 2005 through July
2006) to ensure that the serum PFOA levels measured during enrollment were the same levels
that occurred during pregnancy. Restrictions on residency including living in the same water
district from pregnancy to enrollment were also applied. Preeclampsia and birth defects were
included in the analysis as well. Maternal age, education, and smoking status were possible
cofounders. A total of 1,845 pregnancies for 1,505 women met the requirements for the study.
Regression models were used for statistical analysis. The median PFOA serum level was 21.2
ng/mL, and the mean serum PFOA level was 48.8 ng/mL.
No association was observed between serum PFOA levels and miscarriage, low weight
birth, or pre-term birth. Serum PFOA levels above the median were weakly associated with
preeclampsia. A weak association (odds ratio 1.7, CI: 0.8-3.6) was observed between PFOA
levels above the 90th percentile and birth defects. Data interpretations were limited by the
quality of the self-reported pregnancy outcome especially in preeclampsia, miscarriage, and
preterm births.
Joenson et al. (2009) examined the association between PFCs, including PFOA, and
testicular function in 105 Danish men who provided semen and blood samples as part of
reporting for the military draft in 2003. The men chosen had the highest (ranging from 30.1 to
34.8 nmol/L; n=53; 18.2-24.6 years) and lowest (ranging 10.5-15.5 nmol/L; n=52; 18.2-25.2
years) testosterone concentrations. Regression models were used to analyze associations
between PFCs and testicular function. Median serum PFOA concentration was 4.4, 5.0, and 4.9
ng/mL in the high testosterone, low testosterone, and combined groups, respectively. A
nonsignificant negative association was observed between serum PFOA concentration and semen
volume, sperm concentration, sperm count, sperm motility, or sperm morphology. No
association was observed between serum PFOA concentration and testosterone, estradiol, sex
hormone binding globulin (SHBG), luteinizing hormone (LH), follicle-stimulating hormone
(FSH), and inhibin B.
Anthropometric Endpoints
Apelberg et al. (2007) measured PFOA in the cord blood of 293 newborns (singleton
births) born November 26, 2004 through March 16, 2005 in Baltimore, MD, at Johns Hopkins
Hospital. Maternal and infant data including maternal birth cohort, social class, place of
residence, past pregnancies, insurance type, BMI, age, race, education, marital status, parity,
gestational age, smoking status, and infant sex were collected from the hospital database and
forms filled out at time of delivery. All samples of cord blood contained PFOA (geometric mean
1.6, range 0.3-7.1 ng/mL). Higher concentrations (p<0.03) of PFOA were measured in infants
born to obese mothers. Concentration was evenly distributed by race, but not by sex. Male
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infants had significantly lower (p<0.01) PFOA concentrations than females. Lower
concentrations (p<0.02) were also measured in instances of multiparous birth. No correlations
were observed between PFOA and maternal age, gestational age, education, insurance type,
marital status, smoking status, and place of residence. Birth weight, head circumference, and
ponderal index were negatively associated with PFOA.
At birth, Fei et al. (2007, 2008a) examined the association between plasma PFOA
concentration and infant birth weight, gestation age, placental weight, birth length (skeletal
growth), head circumference (brain), and abdominal circumference (liver size). The mean
plasma PFOA concentration from the first blood sample, the second blood sample, and the cord
blood sample was 5.6, 4.5, and 3.7 ng/mL, respectively. Plasma PFOA concentration ranged
from below the limit of quantitation (<1.0 ng/ml) to 41.5 ng/ml. Mean infant birth weight was
3623 g; 24 infants were categorized as low birth weight infants (<2500 g). The mean gestational
age was 280.4 days, and 53 were preterm births and 139 were post-term births. Plasma PFOA
concentration was significantly inversely associated with birth weight among normal-weight
women (adjusted P=-10.63 g; 95% CI, -20.79 to -0.47 g). A statistically significant decrease was
observed in birth length in the second and fourth quartiles of PFOA exposures compared to the
first quartile after adjustment for confounders. When maternal PFOA concentration was
considered as a continuous variable in relation to fetal growth indicators, a statistically
significant negative association between abdominal circumference and birth length were
observed after adjustment for confounders. The authors noted that any associations may not be
causal, but rather due to chance or confounding factors.
Monroy et al. (2008) examined the association between maternal serum and umbilical
cord PFCs, including PFOA, and birth weight. The population (n=101 pregnant mothers) was
selected from those recruited from January 2004 to June 2005, to participate in the Family Study
in Canada. As part of the study, the women provided blood samples during the second trimester
of pregnancy and filled out an obstetrical history questionnaire. Another blood sample was
provided at delivery as well as blood from the umbilical cord. Age, occupation, pre-pregnancy
height and weight, education, and infant gender and birth weight were obtained post delivery.
Paired t-test and linear regression analysis were used to analyze the data. The mean serum
PFOA concentration was 2.54, 2.24, and 1.94 ng/mL at 24-28 weeks, delivery, and in the
umbilical cord, respectively. There was no association between serum or umbilical cord PFOA
concentration and birth weight.
As mentioned above, Nolan et al. (2009, 2010) published two studies that examined birth
outcome in infants born to mothers whose partial or full public water source was from the Little
Hocking Water Association (LHWA) in Washington County, OH, or from the same geographic
area but with water not provided by the LHWA. The population serviced by LHWA had serum
levels of PFOA (6.78 |ig/L) that were approximately 50 times higher than serum levels of the
general population. Archival data for births, including birth weight, gestation age, plurality,
neonatal sex, race, mother's age, and mother's zip code, in Washington County, OH, from
January 1, 2003 to September 1, 2005 were obtained for analysis. Analysis of variance and
regression models or logistic regression models were used for statistical analysis. The noted
limitations included lack of individual exposure levels, possible misclassification of mothers by
zip code, misclassification of socioeconomic status, possible reporting error in the database used
to abstract the data, and using data from only live births for analysis.
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Nolan et al. (2009) examined the association between birth weight with PFOA in a total
of 1,555 singleton neonates (777 male, 778 female): 11% were born to mothers fully serviced by
LHWA, 13% born to mothers partially serviced by LHWA, and 76% born to mothers not
serviced by LHWA. For mothers with public water service provided completely, partially, or not
at all by LFIWA, the incidence of low birth weight was 3.6%, 3.8%, and 8.1%, respectively. The
national incidence of low birth weight is 8.2%. No differences were found between birth
weights of infants born to mothers whose partial or full water source was serviced by LHWA and
infants born to mothers whose water source was serviced by another provider.
Washino et al. (2009) investigated the correlation between prenatal exposure to
perfluorinated chemicals and reduced fetal growth in a population of women (n=428, average
age 30.5 years) living in or near Sapporo, Japan. Blood samples were obtained from the women
either after the second trimester or after delivery for PFOA and PFOS concentration
determination. PFOA was detected in 92.8% of the blood samples, and PFOS was detected in all
samples. Maternal serum PFOA concentrations ranged from <0.5 (limit of detection) to 5.2
ng/mL. A regression model adjusted only for gestational age showed a correlation between
logic-transformed PFOA concentrations and decreased birth weight. The same model showed a
correlation between logio-transformed PFOA concentrations and decreased birth weight in male
infants. However, when the regression model was adjusted for multiple factors, including
maternal age, education level, smoking status, BMI, parity, how infant was delivered, infant sex,
gestational age, and when blood sample was taken, no correlation was found. No correlations
were observed between PFOA and birth length, chest circumference, or head circumference.
Data interpretation is limited by small sample size, selection bias, and possible errors in birth
length, chest circumference, and head circumference measured due to molding during birth.
Hamm et al. (2009) investigated whether increasing maternal exposure to PFCs,
including PFOA, was associated with fetal growth and gestation length. Pregnant women living
in and around the city of Edmonton, Alberta, Canada, (n=252; age >18 years) submitted serum
samples at 15-16 weeks gestation as part of a screening for birth defects between December 15,
2005 and June 22, 2006. The serum samples were included in the study if the mother had given
birth to a live singleton with no evidence of malformations at greater than or equal to 22 weeks
of gestation. Maternal serum PFOA concentrations ranged from below the limit of detection to
18 ng/mL and had a mean of 2.1 ng/mL. A total of 21 births (8.3%) were classified as preterm
deliveries and 16 infants (6.3%) were considered small for gestational age. The mean infant
birth weight was 3349 g and the mean birth weight z-score was 0.062. The natural log of
maternal serum PFOA concentration was associated with changes in birth weight of-37.4 g
(95% CI: -86.0-11.2) after adjustment for covariates. No association was observed between the
untransformed serum PFOA concentration and birth weight. There was no association between
maternal serum PFOA concentration and birth weight z-score, gestation length, or preterm
delivery. Data interpretation may have been limited by the small sample size, selection bias, and
possible errors in birth length, chest circumference, and head circumference measurements.
Andersen et al. (2010) examined the association between maternal plasma PFOA
concentration and offspring weight, length, and BMI at 5 and 12 months of age. The mothers
(n=1010) reported the information during an interview and weight and length measurements
were used to calculate BMI. Maternal plasma PFOA concentration was inversely associated
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with weight at 5 months (P -30.2, 95% CI -59.3- -1.1), BMI at 5 months (P -0.067, 95% CI -
0.129- -0.004), weight at 12 months (P -43.1, 95% CI -82.9- -3.3 ), and BMI at 12 months in
boys (P -0.078, 95% CI -0.144- -0.011) in models adjusted for maternal age, parity,
prepregnancy BMI, smoking, gestational age at blood draw, socioeconomic status, and
breastfeeding. No associations were observed between maternal plasma PFOA concentration
and the endpoints for female children in the adjusted models.
Neurodevelopmental Endpoints
Fei et al. (2008b) examined the association between plasma PFOA concentration in
pregnant women and motor and mental developmental milestones of their children. The mothers
self-reported the infant's fine and gross motor skills and mental development at 6 and 18 months
of age. There was no association between maternal plasma PFOA concentration and Apgar
score or between maternal plasma PFOA concentration and fine motor skills, gross motor skills,
or cognitive skills at 6 and 18 months of age. The children born to women having higher plasma
PFOA concentrations reached developmental milestones at the same times as children born to
women having lower plasma PFOA concentrations. The authors concluded that there was no
association between maternal early pregnancy levels of PFOA and motor or mental
developmental milestones in offspring.
Fei and Olsen (2011) examined the association between prenatal PFOA exposure and
behavior or coordination problems in the children at age 7. Behavioral problems were assessed
using the Strengths and Difficulties Questionnaire (SDQ), and coordination problems were
assessed using the Developmental Coordination Disorder Questionnaire (DCDQ) completed by
the mothers. A total of 787 mothers completed the SDQ and 537 completed the DCDQ for
children aged 7.01-8.47 years (mean age 7.15 years). The mean maternal PFOA concentration
was 5.7 ng/mL, and PFOA levels were divided into quartiles:
-------
ratio for serum PFOA and parental report of ADHD and ADHD medication use was 1.19 (95%
CI 0.95-1.49). Data interpretation were limited by the cross-sectional study design, random
misclassification error resulting from using current PFOA levels as proxy measures of
etiologically relevant exposures, and other confounders that were not included in the available
data.
Postnatal Development
Lopez-Espinosa et al. (2011) examined the association of serum PFOA concentration and
the age of puberty in exposed children of the Mid-Ohio Valley. Data from the C8 Health project
(sex steroid hormone levels, self-reported menarche status) along with detailed date of birth
information were used to determine age of puberty in males (n=3076) and females (n=2931)
aged 8-18 years. Serum PFOA concentrations were divided into quartiles: <11.4, 11.4-23, >23-
58, and >58 ng/mL. Confounders including age at survey, BMI, BMI z-score, height, family
income, ethnicity, smoking status, alchohol consumption, date and time of sample collection
were included in the logistic regression models used to analyze the data. The median PFOA
concentrations were 26 and 20 ng/mL in boys and girls, respectively. No association between
PFOA concentration and puberty was observed for boys. Reduced odds of having reached
puberty was associated with higher PFOA exposure in girls (OR=0.57, 95% CI 0.37-0.89).
There were 130 days of delay between the highest and lowest quartile. Reduced odds of
experiencing menarche at a younger age (10-15 years) was also observed (OR 0.83, 95% CI
0.74-0.93). The results suggested that PFOA was associated with a later age of menarche. The
authors expressed caution in interpretion of the data because of lack of serum PFOA
concentration prior to puberty, PFOA concentration was measured after the attainment of
puberty, and lack of secondary sexual maturation data (physical, Tanner criteria, and biomarker
measurements).
Christensen et al. (2011) used data from a prospective cohort study in the United
Kingdom to perform a nested case-control study examining the association between age at
menarche and gestational exposure to perfluorinated chemicals including PFOA. The study
population from the Avon Longitudinal Study of Parents and Children included single-birth
female subjects who had completed at least 2 puberty staging questionnaires between the ages of
8 and 13 years and whose mothers provided at least one analyzable prenatal serum sample. If
more than one serum sample were available, the earliest sample provided was used for analysis.
The study does not provide information as to when samples were collected. The females were
divided into two groups including a those who experienced menarched prior to age 11.5 years
(n=218) and a random sample of those who experienced menarche after age 11.5 (n=230).
Confounders including the mother's pre-pregnancy BMI, age at delivery, age at menarche,
educational level, and the child's birth order and ethnic background were included in linear and
logistic regression models used to analyze the data. The median maternal serum PFOA
concentrations were 3.9 and 3.6 ng/mL for the early menarche and non-early menarche groups,
respectively. There was no association between age at menarche and prenatal PFOA exposure.
The authors noted that a modest non-significant association between the odds of earlier
menarche and prenatal serum PFOA concentrations above the median. For all models, the
confidence intervals included the null value of 1.0. The limitations of the study included having
a small sample size, using a single maternal gestational serum sample for perfluorinated
chemical measurement, and missing information from the controls and cases.
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Halldorsson et al. (2012) examined prenatal exposure to PFOA and the risk of being
overweight at 20 years of age in a prospective study. A birth cohort consisting of 665 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 3.7±2.0 ng/mL with quartiles of 2.4±0.6, 3.3±0.4, 4.2±0.5, and 5.8±1.9
ng/mL. In covariate adjusted analyses, female offspring whose mothers were in the highest
quartile had 1.6 kg/m2 (95% CI: 0.6, 2.6) higher BMI and 4.3 cm (95% CI: 1.4, 7.3) higher waist
circumference compared to offspring whose mothers were in the lowest quartile. Female
offspring of mothers in the highest versus lowest PFOA quartile were also more likely to be
overweight [relative risk (RR) 3.1 (95%CI: 1.4, 6.9)] and to have a waist circumference >88 cm
at 20 years of age [3.0 (95%CI: 1.3, 6.8)]. Among female participants who provided blood
sample at clinical examination (n = 252), maternal PFOA concentration was positively
associated with insulin, leptin and the leptin-adiponectin ratio; and inversely associated with
adiponectin levels. PFOA was not associated with overweight or obesity in male offspring.
4.1.2 Cancer
Occupational Population Studies
Several occupational studies examining cancer mortality have been conducted at 3M's
Cottage Grove facility in Minnesota. These studies have reported associations with prostate
cancer mortality and cerebrovascular disease mortality (Alexander 2001a,b; Gilliland and
Mandel, 1993).
Lundin et al. (2009, previously described) further examined occupational exposure to
PFOA and death from cancer in Cottage Grove employees explored by Gilliland and Mandel
(1993) using a longer period of enrolment and later follow-up. Cancers outcomes including
liver, bladder, pancreas, testes, and prostate were examined in the 3,993 workers. Mortality from
liver, pancreatic, prostate, and testicular cancer in PFOA workers was not elevated compared to
the general population of Minnesota. Within the cohort, the SMR for prostate cancer, SMR=2.1,
95% CI 0.4-6.1, was elevated in workers who had ever worked a job in which they were exposed
to PFOA.
Researchers suggested that pancreas acinar cell adenomas seen in rats exposed to PFOA
may be the result of increased cholecystokinin (CCK) levels secondary to blocked bile flow
(Obourn et al., 1997). CCK is a peptide hormone that stimulates the digestion of fat and protein,
causes the increased production of hepatic bile, and stimulates contraction of the gall bladder.
As a result, CCK was measured in male workers (n=74 males) at the 3M's Cottage Grove plant
in 1997 as part of the medical surveillance program (Olsen et al., 1998, 2000). Employees'
serum PFOA levels were stratified into 3 categories (<1, 1- <10, and >10 ppm). The mean CCK
values for the 3 PFOA categories were 33.4, 28.0, and 17.4 pg/mL, respectively. The means in
the 2 serum categories < 10 ppm were at least 50% higher than in the > 10 ppm category. A
statistically significant (p = 0.03) negative association between mean CCK levels and the three
PFOA categories was observed. A multiple regression model of the natural log of CCK and
serum PFOA levels continued to display a negative association after adjusting for potential
confounders. As stated previously (Olsen et al., 2000), no abnormal liver function tests,
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hypolipidemia, or cholestasis were observed in the workers. The authors suggested that the lack
of a positive association between PFOA and CCK in workers may have resulted from serum
PFOA levels too low to cause an increase in CCK provided that the same mechanism that
increases CCK levels in rodents exists in humans.
Limited data are available on mortality and cancer incidence at DuPont's Washington
Works Plant in Parkersburg, WV. Several studies of this population have been published
(Walrath and Burke, 1989; Karns and Fayerweather, 1991; DuPont, 2003). In the most recent
report (DuPont, 2003) cancer incidence for active employees was reported for the 1959-2001
time period and mortality data were reported for active and retired employees for 1957 through
2000 (DuPont, 2003). No information on employee exposure, lifestyle factors, employee
demographics, or information on other chemicals used at the plant, is presented in the DuPont
(2003) report.
Cancer cases were identified through a combination of company health and life insurance
claims and company cancer and mortality registries. Standardized incidence ratios (SIRs) were
only calculated for those cancers for which 5 or more cases were observed, which included 14
types of cancer. Two of those cancer types were elevated and statistically significant (p = 0.05):
bladder [standardized incidence ratio (SIR) = 1.9; 95% CI (1.15-3.07)] and kidney and urinary
organs [SIR = 2.3 (95% CI = 1.36-3.65)]. All of the reported cases were male. Some other cancer
types with elevated SIRs (not statistically significant at p = 0.05) included myeloid leukemia
(2.02), cancer of the larynx (1.77), multiple myeloma and immunoproliferative (1.72), malignant
melanoma of skin (1.3), testicular cancer (1.46), and brain cancer (1.2).
Two separate analyses of leukemia incidence were conducted prior to the DuPont (2003)
effort (Walrath and Burke, 1989; Karns and Fayerweather, 1991). Walrath and Burke (1989)
reported a statistically significant (p = 0.10, significance level chosen by researchers) elevated
odds ratio (OR) of 2.1 for leukemia incidence for male employees working at the plant from
1956-1989. Eight cases (all male) of different types of leukemia were identified. The OR
remained elevated when the workers were divided into wage and salaried employees (2.2 and
2.0, respectively). In the follow-up case-control study by Karns and Fayerweather (1991), four
controls were selected from the plant for each case, matched on gender, age and payroll status.
Matched odds ratios were significantly elevated (p = 0.10) for employees who had previously
worked as custodians and engineers, 8.0 (90% CI, 1.1-60.0) and 7.9 (90% CI, 1.0 - 76.0),
respectively and remained elevated (although not statistically significant) for these same job
categories within the plant (OR= 4.0 and 5.1, respectively). Matched ORs were also reported
based on the area of the plant where the cases worked; however, no statistically significant
association (p = 0.10) was identified.
The data reported above do not show any statistically significant (p = 0.10) elevations in
leukemia deaths, possibly because the number of cases was very small and divided among
different types of leukemia. The Washington Works data provide some insight as to where more
medical surveillance should be concentrated at this plant but provide little information about the
relationship of PFOA to mortality or cancer incidence since no exposure information, use of
other chemicals, or lifestyle information was collected on these employees.
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General Population Studies
Eriksen et al. (2009) examined the association between plasma PFOA concentration and
the risk of cancer in the general Danish population. The study population was chosen from
individuals (50-65 years of age) who had enrolled in the prospective Danish cohort Diet, Cancer,
and Health study between December 1, 1993 and May 31, 1997. The Danish Cancer Registry
and Danish Pathology Data Bank were used to identify cancer patients diagnosed between
December 1, 1993, and July 1, 2006. The cancer patients (n=1240) consisted of 1,111 men and
129 women whose median age was 59 years having prostate cancer (n=713), bladder cancer
(n=332), pancreatic cancer (n=128), and liver cancer (n=67). The individuals (n=772) in the
subcohort comparison group were randomly chosen from the cohort study and consisted of 680
men and 92 women whose median age was 56 years. The participants answered a questionnaire
upon enrollment in the cohort study, and data on known confounders were obtained from the
questionnaires. Study participants provided plasma samples which were analyzed for PFOA
concentration by high pressure liquid chromatography coupled to tandem mass spectrometry.
The plasma PFOA concentrations for cancer patients were as follows: men 6.8 ng/ml,
women 6.0 ng/ml, prostate cancer 6.9 ng/ml, bladder cancer 6.5 ng/ml, pancreatic cancer 6.7
ng/ml, and liver cancer 5.4 ng/ml. The plasma PFOA concentrations for the subcohort
comparison group were 6.9, 5.4, and 6.6 ng/mL for the men, women, and combined,
respectively. Incidence rate ratios, crude and adjusted for confounders, did not indicate an
association between plasma PFOA concentration and prostate, bladder, pancreatic, or liver
cancer. The plasma PFOA levels in the population were lower than those observed in
occupational cohorts. This study is novel in that it is the first to examine PFOA levels and
cancer in the general population.
More recently, Vieira et al. (2013) investigated the relationship between PFOA exposure
and cancer among the residents living near the DuPont plant in Parkersburg, WV. This analysis
included incident cases of 18 cancers diagnosed from 1996-2005 in five Ohio counties and eight
West Virginia counties which included public water districts contaminated with PFOA. The
dataset included 7,869 cases from Ohio geocoded to residence and 17,238 cases from West
Virginia linked to water district. Exposure levels and serum PFOA concentrations were
estimated from data collected previously from participants in the C8 Health Project and linked to
the residence or water district of the cancer cases. Indivudual-level exposure was categorized as
very high, high, medium, low, or unexposed based on serum concentrations of >110 |ig/L, 30.8-
109 |ig/L, 12.9-30.7 |ig/L, 3.7-12.9 |ig/L, and unexposed (background levels not given),
respectively. Logistic regression was applied to individual-level data to calculate odds ratios
(OR) and confidence intervals (CI) for each cancer category.
Data were first analyzed by water district. The adjusted odds ratios were increased for
testicular cancer (OR: 5.1, 95% CI: 1.6, 15.6; n=8), and for kidney cancer (OR: 1.7, 95% CI: 0.4,
3.3; n=10) in Little Hocking water district, and for kidney cancer (OR: 2.0, 95% CI: 1.3, 3.1;
n=23) in the Tuppers Plains water district. Residents of Little Hocking also had increased odds
ratio for non-Hodgkin lymphoma (OR: 1.6, 95% CI: 0.9, 2.8; n=14) and prostate cancer (OR:
1.4, 95%CI:0.9, 2.3; n=36).
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Analysis of exposure category and cancer type showed an association between the very
high PFOA exposure category and several cancers, but ORs for lower exposure categories
generally did not support a positive dose-response relationship. Kidney cancer was positively
associated with very high and high exposure categories (OR: 2.0, 95% CI: 1.0, 3.9; n=9 and OR:
2.0, 95% CI: 1.3, 3.2; n=22, respectively) while ORs for medium and low exposure categories
were close to the null compared to the unexposed. The largest OR was for testicular cancer with
the very high exposure category (OR: 2.8, 95% CI: 0.8, 9.2; n=6) but the estimate was imprecise
due to small numbers. ORs for the other cancer categories were all <1.0. Ovarian cancer, non-
Hodgkin's lymphoma, and prostrate cancer were positively associated with the very high
exposure category, but showed weaker or negative associations for the other exposure categories
(Vieiraetal., 2013).
4.2 Animal Studies
4.2.1 Acute Toxicity
Oral Exposure
Dean and Jessup (1978) reported an oral LDso of 680 mg/kg and 430 mg/kg for male and
female CD rats, respectively. Glaza (1997) reported an oral LD50 of greater than 500 mg/kg in
male Sprague-Dawley rats and between 250 and 500 mg/kg in females. Gabriel (1976c) reported
an oral LDso of less than 1000 mg/kg for male and female Sherman-Wistar rats. According to
the Hodge Sterner Scale, these LD50 values suggest that PFOA can be classified as moderately
toxic after acute oral exposures.
Inhalation Exposure
Rusch (1979) reported no mortality in male or female Sprague-Dawley rats following
inhalation exposure to 186,000 mg/m3 for one hour. Kennedy et al. (1986) reported a 4-hour
LCso of 980 mg/m3 for groups of 6 male rats exposed to PFOA as a dust in air. As reported in a
later publication (Kennedy et al., 2004) body weight loss, irregular breathing, and red discharge
around the nose and eyes were observed. Corneal opacity and corrosion were seen at
concentrations greater than or equal to 810 mg/m3.
Dermal/Ocular Exposure
The dermal LD50 in New Zealand White rabbits was determined to be greater than 2000
mg/kg (Glaza, 1995). Kennedy (1985) determined a dermal LD50 of 4300 mg/kg for rabbits, 7000
mg/kg for male rats, and 7500 mg/kg for female rats. The animals lost body weight, exhibited
lethargy, labored breathing, diarrhea, and severe skin irritation (Kennedy et al., 2004). PFOA is
an ocular irritant in rabbits when the compound is not washed from the eyes (Gabriel, 1976d),
but is not an irritant in rabbits when washed from the eye (Gabriel, 1976a). Markoe (1983) found
PFOA to be a skin irritant in rabbits, while Gabriel (1976b) did not.
4.2.2 Short-Term Studies
Oral Exposure
Monkey. In a range-finding study, Thomford (200la) administered PFOA to male Cynomolgus
monkeys as an oral capsule containing 0, 2, or 20 mg/kg-day PFOA for 4 weeks. There were 3
monkeys in the 2 and 20 mg/kg-day groups and one monkey in the control group. Animals were
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observed twice daily for mortality and moribundity and were examined at least once daily for
signs of poor health or abnormal behavior. Body weights were recorded weekly and food
consumption was assessed qualitatively. The monkeys were fasted overnight and blood samples
were collected one week prior to the start of the study and on day 30 for measurement of serum
PFOA, clinical hematology, and clinical chemistry, plus analysis for hormones (estradiol,
estrone, estriol, thyroid stimulating hormone, total and free triiodothyronine, and total and free
thyroxin). Blood samples were also collected from each animal on day 2 (approximately 24
hours after the first dose) for clinical chemistry measurements.
At scheduled necropsy, liver samples were collected for palmitoyl coenzyme A (CoA)
oxidase activity (a biomarker for peroxisome proliferation) and serum PFOA. Liver, testes, and
pancreas were collected and assayed for cell proliferation using proliferation cell nuclear
antigen. Bile was collected from each animal for measurement of bile acid. The adrenals, liver,
pancreas, spleen, and testes from each animal were examined microscopically.
All animals survived to scheduled sacrifice. There were no clinical signs of toxicity in the
treated groups and there was no effect on body weight. Low or no food consumption was
observed for one animal given 20 mg/kg-day. There were no effects on the hormones measured
with the exception of estrone, which was notably lower in the 2 and 20 mg/kg-day PFOA groups.
There was no evidence of peroxisome proliferation or cell proliferation in the liver, testes or
pancreas of treated monkeys. No adverse effects were noted in either gross or clinical pathology
studies. Under the conditions of this study, the no observed adverse effect level (NOAEL) was
20 mg/kg and no lowest observed adverse effect level (LOAEL) was established.
Rat. Pastoor et al. (1987) dosed male Crl:CD (SD) BR rats (n=6/group) for 1, 3, or 7 days with 0
or 50 mg PFOA/kg. Liver sections were collected at necropsy and stained with hematoxylin and
eosin. Sections were also examined by electron microscopy. DNA content was also determined
from the livers of rats dosed for 7 days. Treatment with 50 mg PFOA/kg for 7 days caused a
17% decrease (p<0.05) in mean body weight. Pair fed control rats had a 24% decrease in body
weight. Body weight was not different in rats treated for 1 or 3 days compared to control rats.
Liver weight of rats treated for 1 day was not different from control liver weight. The relative
liver weight of rats treated for 3 days was significantly increased (p<0.05) compared to control
relative liver weight. Absolute and relative liver weights were significantly increased (p<0.05)
after the 7 day treatment with PFOA. A 57% decreased (p<0.05) was observed in relative
hepatic DNA/g liver, but no difference was observed between total amount of hepatic DNA/liver
when compared to the total amount of DNA/liver in control rats.
The hepatocytes of rats treated with PFOA for three days were enlarged with partially
occluded sinusoids, had numerous basophilic granules, eosinophillic granular material in the
cytoplasm, and fewer perinuclear glycogen vacuols compared to control hepatocytes. Enlarged
with hyperplastic smooth endoplasmic reticulum, increased mitochondria, increased
peroxisomes, decreased rough endoplasmic reticulum, and increased autophagosomes with
electron-dense material were also observed in the hepatocytes.
Loveless et al. (2008) administered 0, 0.3, 1, 10, or 30 mg linear PFOA/kg by oral gavage
to groups of male CD rats (n=10/group) for 29 days. Body weight was recorded on days 0, 3,
and 6-28. At necropsy, blood was collected for hematology, clinical chemistry, and
corticosterone measurements. Tissues were collected for weight and microscopic examination.
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Body weight, weight gain, hematocrit, and hemoglobin were reduced at >10 mg PFOA/kg/day.
Increased reticulocytes and hematopoieses were observed in the rats dosed with 30 mg
PFOA/kg/day. Total and non-HDL cholesterol were significantly reduced at 0.3 and 1
mg/kg/day compared to control. HDL cholesterol was significantly decreased at 0.3, 1, and 10
mg/kg/day. Triglyceride levels were significantly decreased at all doses except 1 mg/kg.
Absolute liver weight (>1 mg/kg/day) and relative liver weight (>10 mg/kg/day) were
significantly increased, minimum to mild (0.3, 1 mg/kg/day) and moderate (>10 mg/kg/day)
hepatocellular hypertrophy, and focal necrosis (>10 mg/kg/day) were observed. Although not
statistically significant, serum corticosterone was increased at > 10 mg/kg/day. Data on several
immunological endpoints were reported as part of the Loveless et al. (2008) publication. The
immunological data are included in Section 4.3.2 of this report.
Cui et al. (2009) exposed male Sprague-Dawley rats (10/group) to PFOA (96% a.i.) at 0,
5, or 20 mg/kg/day for 28 days by gavage once daily. The activity of the rats was observed over
the course of the study. All rats were sacrificed after the final exposure. The rats dosed with 5
mg/kg/day exhibited hypoactivity, decreased food consumption, cachexia, and lethargy during
the third week of the study. Rats dosed with 20 mg/kg/day also exhibited sensitivity to external
stimuli. The visceral index (hepatic, renal, gonal weight/animal's body weight) used to evaluate
hyperplasia, swelling, or atrophy was significantly increased in the treated animals compared to
control animals. In the liver, treatment with 5 or 20 mg PFOA/kg caused hepatic hypertrophy,
fatty degeneration, acidophilic lesions as well as angiectasis (gross dilation) and congestion in
the hepatic sinusoid or central vein. In the lung, treatment with 5 or 20 mg PFOA/kg caused
pulmonary congestion and focal or diffuse thickened epithelial walls. No effects were observed
in the kidneys of the low dose animals, but turbidness and tumefaction (swelling) in the
epithelium of the proximal convoluted tubule was observed at 20 mg PFOA/kg. Under the
conditions of this study, the LOAEL was 5 mg/kg/day based on liver and pulmonary effects, and
no NOAEL was established.
Male Sprague-Dawley rats (n= 10/group) were fed diets containing 0 or 300 pm PFOA for
1, 7, or 28 days in two studies (Elcombe et al., 2010). The mean daily intake for study 1 and
study 2 were 19 and 23 mg/kg/d, respectively. A group of rats was fed diets containing 50 ppm
Wyeth 14,643, a 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. Liver
DNA content and concentration were determined, and plasma was collected for analysis (study 1
only) of liver enzymes, cholesterol, triglycerides, and glucose. Hepatic cell proliferation and
apoptosis were also determined.
In both studies, body weight significantly decreased (p<0.05) after 7 and 28 days on the
PFOA diet. Body weight was not affected by Wyeth 14,643. Absolute liver weight was
significantly increased (p<0.05) in rats fed PFOA diets for 7 days in the first study and in rats
treated for 7 and 28 days in the second study (Table 4-4). The liver-to-body-weight ratio was
significantly higher in rats fed PFOA diets for 7 and 28 days in both studies. Absolute liver
weight and liver-to-body-weight ratio were significantly increased in Wyeth 14,643 diet fed rats
in both studies.
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After 1 day of eating the PFOA diet, plasma AST was significantly decreased and
triglycerides were significantly increased. After 7 and 28 days on the PFOA diet, total
cholesterol, triglycerides, and glucose levels were significantly decreased. AST was also
significantly decreased after 28 days on the PFOA diet. After 1 day on the Wyeth 14,463 diet,
AST and total cholesterol were significantly decreased. After 7 and 28 days on the Wyeth
14,643 diet, ALT, total cholesterol, triglycerides, and glucose levels were significantly
decreased. AST was not significantly decreased in rats fed Wyeth 14,643 diets for 28 days, but it
was after 7 days. Liver DNA concentration was significantly decreased (p<0.05) in all PFOA-
exposed rats except those treated for 1 day in the second study, but liver DNA content was not
altered by PFOA. In the Wyeth 14,643 rats, liver DNA concentration was significantly
decreased after 1 and 7 days in the first study and 7 and 28 days in the second study. Liver DNA
content in the Wyeth 14,643 treated rats was significantly increased after 7 and 28 days in both
studies.
Labelling indices for hepatic cell proliferation, as measured by BrdU incorporation, was
significantly increased after day 1 and 7 in study 1 in both PFOA (p<0.05) and Wyeth 14,643
(p<0.01) diet fed rats. Samples from control livers at day 29 were not available for comparisons.
In study 2, labeling was significantly increased (p<0.05) at all time points in both groups of rats
compared to labeling in control rats (Table 4-5). Apoptosis of hepatic cells was not altered by
treatment with PFOA at any time point. In rats fed diets containing Wyeth 14,643 for 28 days,
hepatic apoptosis was significantly decreased (p<0.01) compared to apoptosis observed in
control livers.
TABLE 4-5. Hepatic Effects of Rats Exposed to PFOA
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
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
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*
From Elcombeetal., 2010
* Significantly different from control (p<0.05); ND=No Data
Histological examination of the livers of PFOA and Wyeth 14,643 diet fed rats showed
decreased glycogen after 1, 7, and 28 days. An increase in hepatocellular hypertrophy was
observed after 7 and 28 days on the diets, fatty vacuolation after 7 days, and increased
hepatocellular hyperplasia was observed after 28 days on the diets. The hepatic observations
were similar in both studies, and findings in Wyeth 14,643-diet fed rats were generally more
pronounced or severe compared to those in PFOA-diet fed rats.
Mouse. Kennedy (1987) fed male and female CRL:CD-1 mice diets containing 0, 30, 300, or
3000 ppm PFOA for 14 days. At necropsy body weight and liver weight were recorded and
analyzed. No histological evaluations were conducted. All mice died at 3000 ppm. Body
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weight was decreased at 300 ppm and one female died at that dose. Both male and female mice
had significantly increased (p<0.05) absolute and relative liver weights at all doses compared to
the control. The LOAEL was 30 ppm based on increased liver weight, and no NOAEL was
established. Kennedy (1987) used lower doses in a follow-up study lasting 21 days. Male and
female mice were fed diets containing 0, 0.01, 0.03, 0.1, 0.3, 1, 3, 10, or 30 ppm PFOA.
Absolute and relative liver weights for male and female mice were significantly increased
(p<0.05) at > 3 ppm PFOA. The LOAEL was 3 ppm based on increased liver weight, and the
NOAEL was 1 ppm.
Loveless et al. (2008) administered 0, 0.3, 1, 10, or 30 mg linear PFOA/kg by oral gavage
to groups of male CD-I mice (n=20/group) for 29 days. Body weight was recorded on days 0, 3,
and 6-28. At necropsy, blood was collected for hematology, clinical chemistry, and
corticosterone measurements. Tissues were collected for weight and microscopic examination.
Body weight was significantly reduced at 10 and 30 mg/kg/day. An increase in neutrophils and
monocytes was observed at >10 mg/kg/day along with a decrease in eosinophils. Serum
corticosterone levels were significantly increased in mice dosed with 10 mg/kg and elevated in
those dosed with 30 mg/kg/day. Total cholesterol and triglycerides were significantly decreased
at >10 mg/kg/day. HDL was significantly reduced at >1 mg/kg/day. In mice treated with 30
mg/kg and water, triglycerides and HDL levels were significantly decreased compared to control
levels. Absolute and relative liver weights were significantly increased at >1 mg PFOA/kg/day.
Increased incidences of microscopic lesions in the liver included mild hepatocellular hypertrophy
at 0.3 mg/kg/day, moderate to severe hypertrophy and individual cell necrosis at >1 mg/kg/day,
and increased hepatocellular mitotic figures, fatty changes, and bile duct hyperplasia at >10
mg/kg/day. Data on several immunological endpoints were reported as part of the Loveless et al.
(2008) publication. The immunological data are included in Section 4.3.2 of this report.
Son et al. (2008) administered 0, 2, 10, 50, or 250 mg/L PFOA (0, 0.49, 2.64, 17.63, or
47.21 mg/kg PFOA) in the drinking water to 4 week old male ICR mice for 21 days. Food and
water consumption, and body weight were recorded daily. At sacrifice, blood was collected and
the liver and kidneys were removed and weighed. Plasma from the blood was used to determine
levels of ALT, AST, BUN, and creatinine. Sections of the liver and kidney were processed and
stained with hematoxylin and eosin or stained for caspase 3 (a biomarker for apoptosis).
Expression of mRNA for tumor necrosis factor-a, interleukin-lp, and transforming growth
factor-p were determined using reverse transcriptase-polymerase chain reaction (RT-PCR).
The mice exposed to 250 mg/L PFOA had significantly reduced (p<0.05) food and water
consumption, and body weight gain (p<0.05) compared to control mice. Body weight gain was
also significantly reduced (p<0.05) in mice receiving 50 mg/L PFOA in the drinking water. In
all PFOA-exposed mice, relative liver weight was significantly increased in a dose dependent
manner (p<0.05) compared to liver weight of control mice. Relative kidney weight was not
affected by PFOA exposure. At > 10 mg/L PFOA, plasma ALT activity was significantly
increased, and at > 50 mg/L PFOA, plasma AST activity was significantly elevated compared to
the activity level in control mice. Exposure to PFOA did not affect BUN or creatinine.
The livers of mice exposed to >50 mg/L PFOA were characterized by enlarged
hepatocytes with acidophilic cytoplasm and the presence of eosinophils. No apoptotic bodies in
the liver were observed with staining for caspase 3. Exposure to PFOA did not affect kidney
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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, interleukin-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.
Wolf et al. (2008a) gavage dosed wild-type 129Sl/SvlmJ mice (n=7-8/group) and
PPARa-null mice (129S4/SvJae-PPARatmlGonz/J, n=6-8/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/group) with 0, 1, or 10
mg PFOA/kg for 7 days to characterize hepatic effects resulting from exposure. The mice were
sacrificed 24 hours following the last dosing. Blood was collected for serum and the livers were
removed and weighed. Liver sections were stained with hematoxylin and eosin for examination
by light microscopy and with uranyl acetate for transmission electron microscopy. Liver
sections were also processed for immunohistochemistry of proliferating cell nuclear antigen
(PCNA). Hepatocyte hypertrophy and vacuolation, observed in both strains of wild-type mice,
were assigned a score from 0-4 based on severity with 0 being no lesions observed and 4 being
panlobular hypertrophy with cytoplasmic vacuolation. Hepatic lesions in 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-
1000 hepatocyte nuclei/animal. Slides were read blind to treatment but with knowledge of
genetic status.
Compared to control values, the absolute and relative liver weights, lesion score, and
labeling index were significantly increased (p<0.05) in a dose dependent manner in both strains
of wild-type mice exposed to PFOA and were also significantly increased (p<0.05) in the wild-
type 129Sl/SvlmJ mice exposed to Wyeth 14,643 (Table 4-6). The absolute and relative liver
weights and lesion score were significantly increased (p<0.05) in a dose dependent manner in all
PFOA-exposed 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 not different from control
values.
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TABLE 4-6. Hepatic Effects in PFOA-treated Mice
Group
Liver weight (g)
Relative liver weight
(%)
Lesion score
Labeling Index
Wild-type CD-I mice
Control
1 mg/kg PFOA
10 mg/kg 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 PFOA
3 mg/kg PFOA
10 mg/kg PFOA
50 mg/kg 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 PFOA
3 mg/kg PFOA
10 mg/kg PFOA
50 mg/kg 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
From Wolf etal.,2008a
* Statistically different from control, p< 0.05
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 were similar, but treatment with PFOA resulted in a different
response. Table 4-7 summarizes the cellular characteristics of the hepatocytes for the control,
POFA-treated, and Wyeth 14,643-treated wild-type and PPARa-null mice on the basis of their
glycogen content, golgi-bodies and associated rough endoplasmic reticulum (ER), mitochondrial,
peroxisomes, and lipid-like cytoplasmic vacuoles.
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TABLE 4-7. Mouse Hepatocyte infrastructure After PFOA or Wythe 14,643 Treatment
Mou se/treatment
Wild-type/
Control
Wild-type/
PFOA
Wild-type/
Wyeth
PPARa-null/
Control
PPARa-null/
PFOA
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
Numerous*
Scattered
From Wolf etal.,2008a.
*Described as electron-dense, nonmembrane-bound spaces morphologically consistent with lipids ranging from the size of
mitochondria to the size of nuclei. The vacuoles were believed to be an accumulation of PFOA.
It is apparent from the data in Table 4-7 that both PFOA and Wyeth 14,643 behaved
similarly in the wild-type strains but differently in the 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 non-membrane bound lipid-like vacuoles
throughout the cytoplasm. At the high dose (10 mg/kg/day), there was an increase in the labeling
index that was not observed with Wyeth 14,643. The authors concluded that the large lipid-like
vacuoles in the hepatocytes of PFOA-dosed 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, and 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, or
0.3 mg/kg/day PFOA (n=4-6/group) for 2 weeks and sacrificed 18-20 hours following the last
dose. Blood was collected and analyzed for triglycerides and cholesterol concentrations and
ALT measurements. Livers were collected and analyzed for triglycerides and cholesterol
concentration, and histopathological changes. The differences in the observations for the three
strains of mice are summarized in Table 4-8.
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TABLE 4-8. Relative Response of 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)
I compared to PPARa-null
(0.1, 0.3 mg/kg/day),
mPPARa (0.3 mg/kg/day)
ND
I 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
I compared to PPARa-null
(0.1, 0.3 mg/kg/day; t
compared to control (0.3
mg/kg/day)
PPARa-null
1 compared to control (0. 1
mg/kg/day)
ND
ND
ND
ND
ND
ND
t compared to mPPARa
(all doses)
From Nakamura et al., 2009
Mice were dosed with with 0, 0.1, or 0.3 mg/kg/day PFOA
|=significant increase (p<0.05); |= significant decrease (p<0.05); ND= no differences
Body weight of the hPPARa mice was slightly lower than mPPAR and PPA-null mice
prior to PFOA treatment and remained lower throughout the dosing regimen. Treatment with
PFOA did not affect plasma ALT or triglyceride concentrations in any group. The 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 4-
7). 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 PFOA
based on increased liver weight and increased liver triglycerides and cholesterol concentrations.
The NOAEL for mPPARa mice was 0.1 mg/kg PFOA. The NOAEL for PPARa-null mice was
0.3 mg/kg because the changes in absolute liver weight were not dose related and the increase in
relative liver weight was not significantly different from the control. The NOAEL for hPPARa
mice was 0.3 mg/kg PFOA, the highest dose tesed. However, a nonsignificant, dose related
increase was observed in plasma cholesterol.
Minata et al. (2010) examined hepatobiliary injury in mice treated with PFOA. Male
wild-type 12984/SvlmJ mice (n=39) and PPARa-null (129S4/SvJae-PparatmlGonz/J mice (n=40)
were orally dosed with 0, 12.5, 25, or 50 umol/kg PFOA (~0, 5.4, 10.8, or 21.6 mg/kg PFOA) for
4 weeks. At the end of 4 weeks, animals were sacrificed and blood, liver, and bile were collected
for clinical chemistry analysis and determination of PFOA concentration. Sections of the liver
were processed for histological examination, oxidative DNA damage, and multidrug resistance
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protein 2 (Mdr2) and tumor necrosis factor a (TNF-a) mRNA expression. Bile acid and
phospholipid contents in bile were determined as well as the protein expression of canalicular
bile salt export pump (BSEP) and canalicular multidrug resistance-associated protein 2 (MRPs).
Absolute and relative liver weights in all PFOA treated wild-type and 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 and in PPARa-null mice at
50 umol/kg compared to the concentrations of their respective controls. Plasma ALT was not
different from control in the treated mice. In wild-type mice, total bilirubin was significantly
decreased at 12.5 umol/kg and significantly increased at 50 umol/kg. In PPARa-null mice, total
bilibrubin was significantly increased at 50 umol/kg. Total bile acid was significantly increased
at 50 umol/kg in PPARa-null mice. Total cholesterol was significantly decreased in wild-type
mice at 25 and 50 umol/kg and total triglyceride was significantly increased at 12.5 and 25
umol/kg. Total cholesterol was significantly decreased at 12.5 and 25 umol/kg and significantly
increased at 50 umol/kg 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, or 50
umol/kg. A dose dependent increase in eosinophilic cytoplasmic changes consistent with
peroxisome proliferation was observed in liver parenchyma, but no fat droplets or focal necrosis
were observed in wild-type mice. An increase in bile duct epithelium thickness suggested slight
cholangiopathy in wild-type mice at 25 and 50 umol/kg. Increased apoptosis in hepatic cells,
hepatic arterial walls, and bile duct epithelium was observed at 25 and 50 umol/kg in wild-type
mice. Ultrastructure examination of livers from PFOA treated wild-type mice showed decreased
glycogen granules, degranulated or disrupted rough endoplasmic reticulum, nuclear vacuoles,
extensive peroxisome proliferation, and slight mitochondria proliferation.
In PPARa-null mice treated with 12.5, 25, or 50 umol/kg PFOA, hypatocellular
hypertrophy, cytoplasmic vacuolation, and increased microvesicular steatosis were observed.
These observations are consistent with Wolf et al. (2008a). At 50 umol/kg, 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. Increased apoptosis was observed in bile
duct epithelium at 25 and 50 umol/kg in PPARa-null mice. Ultrastructure examination of livers
from PFOA treated PPARa-null mice showed decreased glycogen granules, degranulated or
disrupted rough endoplasmic reticulum, increased cytoplasmic lipid accumulation, mitochondria
proliferation, and mitochondrial changes including swelling and decreased matrix density.
Peroxisome proliferation was not observed. Ultrastructure of bile duct showed degradation of
cytoplasmic structure, vacuolization, disintegrations of nuclei and organelles, and periductal
infiltration of fibroblasts and macrophages as well as fibrosis.
The marker for oxidative damage, 8-hydroxydeoxyguanosine, and TNF-a were not
elevated or upregulated in wild-type mice. In PPARa-null mice, 8-hydroxydeoxyguanosine was
elevated in the liver at 50 umol/kg and TNF-a mRNA was significantly increased at 25 and 50
umol/kg. The transporter Mdr2 moves biliary phospholipids from hepatocytes to bile and was
significantly upregulated in wild-type mice at all doses, but only at 12.5 umol/kg in PPARa-null
mice. The BSEP transports bile acid from hepatocytes to bile and was significantly decreased in
wild-type mice at 50 umol/kg, significantly increased in PPARa-null mice at 12.5 umol/kg, and
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significantly decreased at 50 umol/kg. The transporter MRP2 also transports bile acid and was
significantly decreased at 50 umol/kg in both groups of rats.
Under the conditions of the study, the LOAEL for male wild-type and PPARa-null mice
was 12.5 umol/kg PFOA (-5.4 mg/kg) based on increased liver weight. A NOAEL was not
established.
Li et al. (2011) investigated the involvement of mouse and human PPARa in PFOA-
induce testicular toxicity. Wild-type, PPARa-null, and humanized PPARa male 129/Sv mice
were given PFOA daily by gavage at doses of 0, 1, or 5 mg/kg/d for six weeks. Body weight and
testis weight were not affected by treated in any group. Absolute and relative-to-body weights of
the epididymis and seminal vesicle plus prostate gland were decreased only in high-dose wild-
type 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-
typ 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 significane was not attained.
Using real-time quantitative PCR, the mRNA levels for several genes associated with
testicular cholesterol synthesis, transport, and testosterone biosynthesis were examined. Levels
of 3-hydroxy-3-methylglutaryl coenzyme A (HMG-CoA) synthase, HMG-CoA reductase, and
aromatase were not changed after treatment in any group. Expression of steriodogenic acute
regulatory protein (which transports cholesterol into mitochondria) was inhibited in wild-type
mice at the high dose and in humanized PPARa mice at both doses; peripherial 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 of 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 was decreased in both groups of wild-type and high-dose humanized
PPARa mice and superoxide dismutase 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.
Inhalation Exposure
No data on the effects of short term inhalation exposures to PFOA were identified in the
literature.
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Dermal Exposure
Fairley et al. (2007) investigated the role of dermal exposure to PFOA in an experiment
to evaluate toxicity in BALB/c mice. The mice were exposed to 0, 0.01%, 0.1%, 0.25%, 0.5%,
1.0%, or 1.5% PFOA (equivalent to 0, 0.25, 2.5, 6.25, 12.5, 25, or 50 mg/kg PFOA). It was
applied to the dorsal surface of both ears daily for 4 days. The mice were sacrificed 6 days later.
Dermal PFOA exposure did not cause reductions in body weight or signs of inflammation at the
application site. A significant increase in liver weight was observed in mice dosed with > 6.25
mg/kg PFOA (p<0.01) compared to control liver weight. Under the conditions of the study, the
LOAEL was 6.25 mg/kg PFOA based on increased liver weight, and the NOAEL was 2.25
mg/kg PFOA.
4.2.3 Subchronic Studies
Oral Exposure
Monkey. Goldenthal (1978b) administered rhesus monkeys (2/sex/group) doses of 0, 3, 10, 30
or 100 mg/kg-day PFOA by gavage for 90 days. Animals were observed twice daily and body
weights were recorded weekly. Blood and urine samples were collected once during a control
period, and at 1 and 3 months for hematology, clinical chemistry and urinalysis. Organs and
tissues from animals that were sacrificed at the end of the study and from animals that died
during the treatment period were weighed, examined for gross pathology, and processed for
histopathology.
All monkeys in the 100 mg/kg-day group died between weeks 2-5 of the study. Signs and
symptoms which first appeared during week 1 included anorexia, frothy emesis, swollen face
and eyes, decreased activity, prostration and body trembling. Three monkeys from the 30 mg/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 vs. 3.78 kg for the control). The
results of the urinalysis, and hematological and clinical chemistry analyses were comparable for
the control and the 3 and 10 mg/kg-day groups at one and three months.
At necropsy, there were significant decreases in the absolute heart and brain weight and
relative liver weight in 10 mg/kg-day females. At 3 mg/kg-day the relative pituitary weight in
males was significantly increased. The biological significance of these weight changes is
difficult to assess, as they were not accompanied by morphologic changes.
In animals that died, one male and two females from the 30 mg/kg-day group and all
animals from the 100 mg/kg-day group had marked diffuse lipid depletion in the adrenal glands.
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All males and females from the 30 and 100 mg/kg-day groups also had slight to moderate
hypocellularity of the bone marrow and moderate atrophy of lymphoid follicles in the spleen.
One female from the 30 mg/kg-day group and all animals in the 100 mg/kg-day group had
moderate atrophy of the lymphoid follicles in the lymph nodes.
The one male in the 30 mg/kg-day group that survived until terminal sacrifice had slight
to moderate hypocellularity of the bone marrow and moderate atrophy of lymphoid follicles in
the spleen. Under the conditions of this study, the male LOAEL was 3 mg/kg-day based on
increased relative pituitary weight, and no NOAEL was established. The female LOAEL was 10
mg/kg based on decreased heart and brain weight, and the NOAEL was 3 mg/kg.
Rat. Goldenthal (1978a) administered ChR-CD rats (5/sex/group) dietary levels of 0, 10, 30,
100, 300, and 1000 ppm PFOA for 90 days. These dose levels are equivalent to 0, 0.56, 1.72,
5.64, 17.9, and 63.5 mg/kg-day in males, and 0.74, 2.3, 7.7, 22.36 and 76.47 mg/kg-day in
females. Animals were observed twice daily and body weight and food consumption were
recorded weekly. Blood and urine samples were collected during the pretest period and at 1 and
3 months of the study for hematology, clinical chemistry and urinalysis. At necropsy, the organs
from the control, 100, 300, and 1000 ppm groups were weighed and examined for
histopathologic lesions; livers from the 10 and 30 ppm groups were also examined
microscopically.
There were no treatment-related changes in behavior or appearance and no treatment
related deaths. There was a decrease in body weight gain for male rats at the 300 and 1000 ppm
dose levels. At 13 weeks, mean body weight of males in the 1000 ppm group was significantly
less than that of controls. There were no treatment related effects on the hematologic,
biochemical or urinary parameters.
Relative kidney weights were significantly increased in males in the 100, 300, and 1000
ppm groups. However, absolute kidney weights were comparable among groups, and there were
no histopathological lesions. Absolute liver weights were significantly increased in males in the
30, 300 and 1000 ppm groups and in females in the 1000 ppm group. The mean absolute liver
weight for the 0, 10,30, 100, 300, and 1000 ppm groups was 13.4, 14.3, 19.1, 18.6, 20.1, and
19.2 g, respectively. Relative liver weights were significantly increased in males in the 300 and
1000 ppm groups and in females in the 1000 ppm group. Discoloration on the surface of the liver
was observed in male rats in the 1000 ppm group. Hepatocellular hypertrophy (focal to
multifocal in the centrilobular to midzonal regions) was observed in 4/5, 5/5, and 5/5 males in
the 100, 300, and 1000 ppm groups, respectively. Hepatocyte necrosis was observed in 2/5, 2/5,
1/5, and 2/5 males in the 30, 100, 300, and 1000 ppm groups, respectively. Under the conditions
of this study, the LOAEL for males is 30 ppm (1.72 mg/kg-day) based on liver effects and the
NOAEL is 10 ppm (0.56 mg/kg-day); the LOAEL for females is 1000 ppm (76.5 mg/kg-day)
based on increased liver weight, and the NOAEL is 300 ppm (22.4 mg/kg-day).
In a dietary study, reported in Palazzolo (1993) and Perkins et al. (2004), male ChR-CD
rats (45-55 per group) were administered concentrations of 1, 10, 30, or 100 ppm PFOA for 13
weeks. These doses are equivalent to 0.06, 0.64, 1.94, and 6.50 mg/kg-day. There were two
control groups (a nonpair-fed control group and a control group pair-fed to the 100 ppm dose
group); both fed the basal diet. Following the 13-week exposure period, 10 animals per group
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were fed basal diet for an 8-week recovery period. The animals were observed twice daily for
clinical signs of toxicity, and body weights and food consumption were recorded weekly. Food
consumption was recorded daily for the pair-fed animals.
A total of 15 animals per group were sacrificed following 4, 7, or 13 weeks of treatment;
10 animals per group were sacrificed after 13 weeks of treatment and following the 8 week
recovery period. Serum samples collected from 10 animals per group at each scheduled sacrifice
during treatment and from 5 animals per group during recovery were analyzed for estradiol, total
testosterone, luteinizing hormone, and PFOA. The level of palmitoyl CoA oxidase was analyzed
from a section of liver that was obtained from 5 animals per group at each scheduled sacrifice.
Weights of the brain, liver, lungs, testis, seminal vesicle, prostate, coagulating gland, and urethra
were recorded, and these tissues were examined histologically. In addition, the brain, liver,
lungs, testis, seminal vesicle, and prostate were preserved in glutaraldehyde for electron
microscopic examination.
In the analysis of the data, animals in groups exposed to 1, 10, 30, and 100 ppm PFOA
were compared to the control animals in the nonpair-fed group, while the data from the pair-fed
control animals were compared to animals exposed to 100 ppm PFOA. At 100 ppm, significant
reductions in body weight and body weight gain were seen compared to the pair-fed control
group during week 1 and the nonpair-fed control group during weeks 1-13. Body weight data in
the other dosed-groups were comparable to controls. At 10 and 30 ppm, mean body weight gains
were significantly lower than the nonpair-fed control group at week 2. These differences in body
weight and body weight gains were not observed during the recovery period. Animals fed 100
ppm consumed significantly less food during weeks 1 and 2, when compared to the nonpair-fed
control group. Overall, there was no significant difference in food consumption. There were no
significant differences among the groups for any of the hormones evaluated although there
appeared to be some indication of elevated estradiol for the 100 ppm group at week 5.
Significant increases in absolute and relative liver weights and hepatocellular
hypertrophy were observed at weeks 4, 7, or 13 in the 10, 30 and 100 ppm groups (Table 4-9).
There was no evidence of any degenerative changes. Hepatic palmitoyl CoA oxidase activity was
significantly increased at weeks 4, 7, and 13 in the 30 and 100 ppm groups. At 10 ppm, hepatic
palmitoyl CoA oxidase activity was significantly increased at week 4 only. During recovery,
however, no liver effects were observed, indicating that the treatment-related liver effects were
reversible. Under the conditions of this study, the LOAEL is 10 ppm (0.64 mg/kg-day) based on
increases in absolute and relative liver weights with hepatocellular hypertrophy. The NOAEL is
1.0 ppm (0.06 mg/kg-day).
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TABLE 4-9. Liver Effects In Male Rats
Parameter
Palmitoyl CoA
Oxidase (lU/g)
Hepatocellular
Hypertrophy
Hepatocellular
Necrosis,
Coagulative
Absolute Liver
Weight (g)
Liver/Body
Weight (%)
Week
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
3.97±0.37
3.75±0.29
3.53±0.28
Ob
5±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
4.07 ±0.27
3.76±0.37
3.23±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
3.73 ±0.23
3.64±0.33
3.24±0.30C
0.64 (10 ppm)
14±3.8C
18 ±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
4.49±0.32C
4.12±0.37
3.69±0.32
1.94 (30 ppm)
24 ± 11.4°
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
5.77 ±0.60C
5.14±0.53C
4.21±0.56C
6.5 (100 ppm)
37± 14.8cd
54± 35.3cd
17 ±4.5cd
14/15
15/15
15/15
2/15
1/15
0/15
25.44 ±1.89C
27.76±3.51C
26.78±5.47C
6.73 ±0.49C
6.06±0.59C
5.49±0.84C
FromPalazzolo 1993; Perkins et al. 2004
Mean ±SD; NA= not available
a non-pair-fed controls;b pair-fed controls;'
significant (p < 0.05) with pair-fed control
statistically significant (p < 0.05) with non-pair-fed control; statistically
Inhalation and Dermal Exposure
No data on the effects of subchronic inhalation or dermal exposures to PFOA were
identified in the literature.
4.2.4 Neurotoxicity
Johansson et al. (2008) gave male neonatal Naval Medical Research Institute mice
(NMRI, 3-4 litters, -5-6 male pups/litter) a single gavage dose of 0, 0.58, or 8.7 mg PFOA/kg in
a lecithin/peanut oil emulsion on postnatal day 10, the approximate peak time of rapid brain
growth in mice. Spontaneous behavior (locomotion, rearing, and total activity) and habituation
in response to a placement in unfamiliar environment were tested in 10 mice in each group at
ages 2 and 4 months. Each test period was divided into three 20 minute periods. The habituation
ratio was determined by dividing the activity for the third 20 minute period by the activity for the
first period. A high habituation ratio indicated that movement patterns of the exposed animals
when placed in an unfamiliar test chamber differed from control by displaying comparatively
low activity for the first 20 minutes and comparatively higher activity for the last 20 minutes.
Exposure to PFOA did not affect body weight or body weight gain in male NMRI mice
following treatment. Compared to controls, the habituation ratio for rearing and locomotion in
the high dose animals was elevated compared to controls at 2 and 4 months with a significantly
higher ratio (p<0.01) at 4 months when compared to 2 months. At 4 months, the changes in
activity patterns for the high dose were significant (p<0.01) compared to controls for locomotion,
rearing, and total activity. The results at the low dose were less pronounced, with a significant
impact on locomotion and slight changes in on rearing behavior.
At 4 months of age, mice were tested for nicotine-induced behavior and behavior in the
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elevated plus-maze. Increased activity is the expected response to nicotine injection (80 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 were also tested in an elevated plus-maze which determined whether they
would select an enclosed environment (the expected response) over an open environment. No
significant differences were observed in the PFOA exposed mice in this test. Under the
conditions of this study, the clear LOAEL was 8.7 mg/kg based on locomotion, rearing, and total
activity; habituation ratio; and response to nicotine at 2 and 4 months. There were significant
differences in locomotion and total activity at 4 months in the low dose animals which support
identifying the 0.58 mg/kg dose as a marginal LOAEL. However, the data at the low dose are
less compelling as those for the high dose.
In a follow-up to their original study, Johansson et al. (2009) gave male neonatal NMRI
mice (3-4 litters, -5-6 male pups/litter) a single gavage dose of 0 or 8.7 mg PFOA/kg on
postnatal day 10. Protein levels of calcium/calmodulin-dependent protein kinase II (CaMKII),
growth-associated protein-43 (GAP-43), synaptophysin and tau protein were determined in the
cerebral cortex and hippocampus. CaMKII regulates synapotogenesis and synaptic plasticity,
GAP-43 modulates axon sprouting and growth, synaptophysin is a membrane glycoprotein in
presynaptic vesicles, and tau protein is responsible for outgrowth of neuronal processes and
microtubule assembly and maintenance.
Levels of CaMKII protein in the hippocampus were significantly higher (58%, p<0.05) in
mice exposed to PFOA than levels in control mice, but not changed in the cerebral cortex.
Levels of GAP-43 protein in the hippocampus were significantly higher (17%, p<0.05) in PFOA-
exposed mice than levels in control mice, but not changed in the cerebral cortex. Synaptophysin
levels in mice exposed to PFOA were significantly increased in the hippocampus (52%) and
cerebral cortex (82%). Tau protein levels in PFOA-exposed mice were increased 92% and 142%
(p<0.001) in the hippocampus and cerebral cortex, respectively, above levels in control mice.
The authors concluded that alterations of these proteins could be a factor in the altered behavior
of adult mice that were exposed to PFOA as neonates because they are required for normal brain
development.
Onishchenko et al. (2011) exposed pregnant C57BL/6/Bkl mice (n=6/group) to 0 or 0.3
mg PFOA/kg/day in the diet from GDI-end of pregnancy. The behavior of the weaned offspring
was analyzed in locomotor, circadian activity, elevated plus-maze, and forced swim tests at 5-8
weeks of age. Muscle strength and motor coordination tests were given at 3-4 months of age.
The distance traveled over 30 minutes was registered in 5 minute intervals in the locomotor test.
For the circadian activity test, the activity of the mice in social groups was monitored for 48
hours after placement in new cages. Anxiety-like behavior was determined using the elevated
plus maze. Depression-like behavior was determined in the forced swim test by tracking the
time spent floating passively for 2 seconds or longer. Muscle strength (3 trials) was measured by
how long (within 60 seconds) it took the mouse to fall off an upside down lid onto the cage floor.
Motor coordination (4 trials) was measured by how long the mice remained on a rotating drum as
a rotarod accelerated from 4 to 40 rpm over 5 minutes.
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Prenatal exposure to PFOA did not alter offspring locomotor activity, anxiety-related
behavior, depression-like behavior, or muscle strength. In the circadian activity tests, male
offspring exposed to PFOA were significantly more active (p=0.013) and the female offspring
were significantly less active (p=0.036) than than control offspring during the first hour of the
test. PFOA-exposed male offspring were significantly more active (p<0.05) than control males
from the dark phase of Day 1 through the dark phase Day 2. Both male and female offspring
exposed to PFOA had significantly (p<0.05) less inactive periods during the light phase
compared to their respective controls. In the accelerating rotarod test, female offspring exposed
to PFOA exhibited decreased fall latency over the 4 trials compared to control females, but no
effect of treatment was observed in male offspring. The authors concluded that prenatal
exposure to PFOA resulted in sex-related alterations in offspring behavior and motor function.
4.2.5 Developmental/Reproductive Toxicity
Reproductive Effects
A standard oral two-generation reproductive toxicity study of PFOA in Sprague-Dawley
rats was conducted (York, 2002; Butenhoff et al., 2004a; York et al., 2010). Five groups of male
and female SD rats (30/sex/group) were administered PFOA by gavage at doses of 0, 1, 3, 10,
and 30 mg/kg-day. The parental generation (FO) rats (n=30/sex/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/sex/group) were dosed similarly, beginning at weaning. The F2 generation
rats were maintained through lactation day 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 were also determined. Food
consumption, body-weight gain, selected organ weights, gross pathology, and appropriate
histopathology were evaluated.
FO Males. One FO male rat in the 30 mg/kg-day dose group was sacrificed on day 45 of the study
due to adverse clinical signs. Statistically significant increases in clinical signs were also
observed in male rats in the high-dose group that included dehydration, urine-stained abdominal
fur, and ungroomed coat. Significant reductions in body weight and body weight gain were
reported for most of the dosage period and continuing until termination of the study in the 3,10,
and 30 mg/kg-day dose groups. Absolute food consumption values were also significantly
reduced during these periods in the 30 mg/kg-day dose group, while significant increases in
relative food consumption values were observed in the 3, 10, and 30 mg/kg-day within those
same periods.
No treatment-related effects were reported at any dose level for any of the reproductive
parameters assessed. No treatment-related effects were seen at necropsy or upon microscopic
examination of the reproductive organs.
At necropsy, statistically significant reductions in terminal body weights were seen at 3,
10, and 30 mg/kg-day (6%, 11%, and 25% decrease from controls, respectively; p<0.05). The
absolute weight of the liver was statistically significantly increased in all dose-groups (Table 4-
10). Absolute kidney weights were statistically significantly increased in the 1, 3, and 10 mg/kg-
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day dose groups, but significantly decreased in the 30 mg/kg-day group. Organ weight-to-
terminal body weight ratios for the liver and for the left and right kidney were statistically
significantly increased in all treated groups. Organ weight-to-terminal body weight ratios for the
brain were statistically significantly increased for the 3, 10, and 30 mg/kg-day dose groups;
organ weight-to-brain weight ratios were significantly reduced for some organs at the high dose
group, and significantly increased for other organs among all treated groups; these changes
occurred in a pattern typically associated with decrements in body weight. The only histologic
finding was increased thickness and prominence of the zona glomerulosa and vacuolation in the
cells of the adrenal cortex observed in 2/10 males in the 10 mg/kg-day group and 7/10 males in
the 30 mg/kg-day dose group.
TABLE 4-10. Organ weight data from FO males
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*
From Butenhoff et al. 2004a.
Mean±S.D.; n= 29-30; significantly different from control: *p<0.05, **p<0.01.
Under the conditions of the study, the LOAEL for FO parental males is 1 mg/kg-day, the
lowest dose tested, based on significant increases in absolute and relative liver and kidney
weight. A NOAEL for the FO parental males could not be determined.
FO Females. There were no treatment-related effects on clinical signs, body weight, food
consumption, organ weights, or histology of the organs. There were no treatment-related effects
on any of the reproductive parameters assessed, and no treatment-related effects were seen at
necropsy or upon microscopic examination of the reproductive organs. The NOAEL for FO
parental females is 30 mg/kg-day, the highest dose tested.
Fl Generation. Pup body weight on a per litter basis (sexes combined) was significantly
reduced (p<0.01) by 8-10% throughout the first two weeks of lactation in the 30 mg/kg-day
group; at weaning the mean body weights were reduced 4.5%, but the difference was not
statistically significance. Although there were no effects on the viability and lactation indices,
the total number of dead pups during lactation was increased in the 30 mg/kg-day groups; the
difference was statistically significant on lactation days 6-8. No other effects were noted, and
there were no treatment-related findings for the pups necropsied at weaning.
Fl Males. Significant increases in treatment-related deaths (5/60 animals) were reported in Fl
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males in the high dose group between days 2-4 postweaning. One rat was moribund sacrificed on
day 39 postweaning and another was found dead on day 107 postweaning. Clinical signs
included a significant increase in emaciation at 10 and 30 mg/kg-day, and in urine-stained
abdominal fur, decreased motor activity, and abdominal distention at 30 mg/kg-day.
Mean body weight was significantly reduced in the in the 30 mg/kg-day group beginning
on postweaning day 8, in the 10 mg/kg-day group beginning on postweaning day 36, and
towards the end of the study in the 1 and 3 mg/kg-day groups. Terminal mean body weight was
reduced in all treated groups. For all groups, there was a significant, dose-related reduction in
mean body weight gain for the entire dosing period (days 1-113). Absolute food consumption
values were significantly reduced at 10 and 30 mg/kg-day during the entire precohabitation
period (days 1-70 postweaning), while relative food consumption values were significantly
increased.
Statistically significant delays (p< 0.01) in the average day of preputial separation were
observed in high-dose animals versus concurrent controls (52.2 days of age versus 48.5 days of
age, respectively). There were no other effects on any of the reproductive parameters assessed,
and at necropsy no effects on reproductive organs were noted.
The absolute and relative weights of the liver were statistically significantly increased in
all treated groups (p < 0.01). The absolute weights of the left and/or right kidneys were
statistically significantly increased in the 1 and 3 mg/kg-day dose groups and statistically
significantly decreased in the 30 mg/kg-day dose group (Table 4-11). Organ weight-to-terminal
body weight and brain weight ratios for the kidney were statistically significantly increased in all
treated groups. Histopathology was not conducted on the kidney. All other organ weight changes
observed (thymus, spleen, left adrenal, brain, prostate, seminal vesicles, testes, and epididymis)
were probably due to decrements in body weight and not a reflection of target organ toxicity.
Treatment-related microscopic changes were observed in the adrenal glands of high-dose
animals (cytoplasmic hypertrophy and vacuolation of the cells of the adrenal cortex) and in the
liver of animals treated with 3, 10, and 30 mg/kg-day (hepatocellular hypertrophy).
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TABLE 4-11. Organ weight data from Fl males
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
From Butenhoff et al. 2004a.
Mean±S.D.; n= 29-30; significantly different from control: *p<0.05, **p<0.01.
For Fl males, the LOAEL for developmental/reproductive toxicity was 10 mg/kg-day
based on significantly reduced body weight and body weight gain during the lactation period,
and the NOAEL was 3 mg/kg-day. The LOAEL for adult systemic toxicity in the Fl males is 1
mg/kg-day based on significant, dose-related decreases in body weights and body weight gains
(observed prior to and during cohabitation and during the entire dosing period), and in terminal
body weights; and significant changes in absolute and relative liver weights. A NOAEL for adult
systemic toxicity in the Fl males could not be determined.
Fl Females. 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 weights were observed in high-dose animals on days 8, 15, 22, 29, 50, and 57
postweaning, during precohabitation (recorded on the day cohabitation began, when 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 (p<0.01) delays in sexual maturation (the average day of vaginal
patency) were observed in high-dose animals versus concurrent controls (36.6 days of age versus
34.9 days of age, respectively). Prior to mating, the study authors noted a statistically significant
increase in the average numbers of estrous stages per 21 days in high-dose animals (5.4 versus
4.7 in controls). For this calculation, the number of independent occurrences of estrus in the 21
days of observation was determined. This calculation can be used as a screen for effects on the
estrous cycle, but should be followed by a more detailed analysis.
Both 3M (2002) and the U.S. EPA (2002b) conducted a more detailed analysis of the
estrus cycle data. The 3M analysis of the data concluded that there were no differences in the
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number of females with >3 days of estrus or with >4 days of diestrus in the control and high dose
groups. This conclusion is consistent with that of the U.S. EPA (2002b). The cycles were
evaluated as having either regular 4-5 day cycles, uneven cycling (defined as brief periods with
irregular pattern) or periods of prolonged diestrus (defined as 4-6 day diestrus periods) extended
estrus (defined as 3 or 4 days of cornified smears), possibly pseudopregnant, (defined as 6-
greater days of leukocytes) or persistent estrus (defined as 5-or greater days of cornified smears).
The two groups were not different in any of the parameters measured. Thus, the increase in the
number of estrous stages per 21 days that was noted by the study authors was an outcome of the
approach used for the calculations and is not biologically meaningful. There were no effects on
the other reproductive parameters assessed, and at necropsy no effects on reproductive organs
were noted.
No treatment-related effects were observed in the terminal body weights of the Fl
females. The absolute weight of the pituitary, the pituitary weight-to-terminal body weight ratio,
and the pituitary weight-to-brain weight ratio were statistically significantly decreased at 3
mg/kg-day and higher. Since there is not a clear dose-response relationship and histologic
examination did not reveal any lesions, the biological significance of the pituitary weight data is
problematic. No other differences were reported for the absolute weights or ratios for other
organs evaluated. No treatment-related effects were reported following macroscopic and
histopathologic examinations.
For 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. There was a statistically significant increase (p< 0.01) in the number of pups
found dead on lactation day 1 in the 3 and 10 mg/kg-day groups. An independent statistical
analysis was conducted by U.S. EPA (2002c), and no significant differences were observed
between dose groups and the response did not have any trend in dose. There were no 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.
Developmental Effects
Prenatal Development. Prenatal developmental toxicity studies of PFOA have been conducted
in rats, and mice and are summarized below. A greater number of studies have been conducted
using mice as experimental animals than rats. Most studies have used the oral route of exposure.
Rat. Pregnant Sprague-Dawley rats were gavage dosed with 0, 3, 10 or 30 mg/kg-day PFOA
during days 4-10, 4-15, or 4-21 of gestation, or from gestation day 4 to lactation day 21
(Hinderliter et al., 2005; Mylchreest, 2003). Clinical observations and body weights were
recorded daily. On gestation days 10, 15, and 21,5 rats/group/time-point were sacrificed and the
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number, location and type of implantation sites were recorded. Embryos were collected on day
10, and placentas, amniotic fluid, and embryos/fetuses were collected on days 15 and 21.
Maternal blood samples were collected at 2 hours ±30 minutes post-dose. The remaining 5 rats
per group were allowed to deliver. On lactation days 0, 3, 7, 14, and 21, the pups were counted,
weighed (sexes separate), and examined for abnormal appearance and behavior. Randomly
selected pups were sacrificed and blood samples were collected. On lactation days 3,7, 14, and
21, the dams were anesthetized and milk and blood samples were collected; dams were removed
from their litters 1-2 hours prior to collection. Plasma, milk, amniotic fluid extract, and tissue
homogenate (placenta, embryo, and fetus) supernatants were analyzed for PFOA concentrations
by HPLC-MS.
All dams survived and there were no clinical signs of toxicity. In the 30 mg/kg-day
group, mean body weight gain was approximately 10% lower than the control group during
gestation, and mean body weights were approximately 4% lower than controls throughout
gestation and lactation. The number of implantation sites, resorptions, and live fetuses were
comparable among groups on days 10, 15, and 21 of gestation. One dam in the 3 mg/kg-day
group and two dams in the 30 mg/kg-day group delivered small litters (litter size of 3-6 pups as
compared to 12-19 pups/litter in the control group); however, given the small sample size the
biological significance of this finding is unclear. There were no clinical signs of toxicity in the
pups, and pup survival and pup body weights were comparable among groups. Under the
conditions of this study, the maternal LOAEL was 30 mg/kg-day for decreased body weight gain
during gestation, and the NOAEL was 10 mg/kg-day. The developmental NOAEL was 30
mg/kg-day.
Mouse. Lau et al. (2006) conducted a developmental toxicity study of PFOA in mice in order to
ascertain whether there was a sex difference in the bioaccumulation of PFOA in the mouse, and
to evaluate the effects of PFOA on pre- and postnatal development in offspring exposed during
pregnancy. In that study, groups averaging 9-45 timed-pregnant CD-I mice were given 0, 1, 3, 5,
10, 20, or 40 mg/kg PFOA daily by oral gavage on GD1-17. Maternal weight was monitored
during gestation. Dams were divided into two groups. In the first group, dams were sacrificed on
GDIS and underwent maternal and fetal examinations that included measure of maternal liver
weight, and examination of the gravid uterus for numbers of live and dead fetuses and
resorptions. Maternal blood was collected and analyzed for PFOA serum concentration. PFOA
levels in the fetuses were not examined. Live fetuses were weighed and subjected to external
gross necropsy and skeletal and visceral examinations. In the second group of dams, an
additional dose of PFOA was administered on GDIS. Dams were allowed to give birth on GD19.
The day following parturition was designated as PND 1. Time of parturition, condition of
newborns, and number of live offspring were recorded. The number of live pups in each litter
and pup body weight were noted for the first four days after birth and then at corresponding
intervals thereafter. Eye opening was recorded beginning at PND 12. Pups were weaned on PND
23 and separated by sex. The time to sexual maturity was determined by monitoring vaginal
opening and preputial separation beginning on PND 24. Two to four pups per sex per litter were
randomly selected for observation of postnatal survival, growth, and development. Estrous
cyclicity was determined daily by vaginal cytology. After weaning, dams were sacrificed and the
contents of the uteri examined for implantation sites. Postnatal survival was calculated based on
the number of implantations for each dam. Benchmark dose values (BMDs and BMDLs) were
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calculated for the maternal and developmental toxic endpoints.
Signs of maternal toxicity were observed following exposure to PFOA during pregnancy.
Statistically significant (p < 0.05) dose-related increases in maternal liver weight were also
observed, beginning at 1 mg/kg/day. Dose-related decreases in maternal weight gain during
pregnancy were observed beginning at 5 mg/kg/day, with statistical significance (p < 0.05) seen
in the 20 and 40 mg/kg/day dose groups. Under the conditions of the study, a maternal LOAEL
of 1 mg/kg was indicated based on increased liver weight, and a NOAEL was not established.
Signs of developmental toxicity were observed following in utero exposure to PFOA. The
number of implantations in treated mice was comparable to control mice. Statistically significant
(p < 0.05) increases in full-litter resorptions were reported at doses of > 5 mg/kg/day, with
complete loss of pregnancies at the highest dose group of 40 mg/kg/day. A 20% reduction (p <
0.05) in live fetal body weight at term was reported at 20 mg/kg/day. A statistically significant
increase in prenatal loss was observed in the 20 mg/kg/day dose group. Ossification (number of
sites) of the forelimb proximal phalanges was significantly decreased at all doses except 5
mg/kg. Ossification of hindlimb proximal phalanges was significantly decreased at all doses
except 3 and 5 mg/kg. Reduced ossification (p < 0.05) of the calvaria and enlarged fontanel was
observed at 1, 3, and 20 mg/kg and at > 10 mg/kg in the supraoccipital bone. Statistically
significant (p < 0.05) increases in minor limb and tail defects were observed in the fetuses at
doses > 5 mg/kg/day. Under the conditions of the study, a prenatal developmental LOAEL of 1
mg/kg was indicated based on increased skeletal defects, and the NOAEL was not established.
Slight, but statistically significant (p < 0.05), increases in the average time to parturition
were observed at 10 and 20 mg/kg/day. Increases (p < 0.05) in stillbirths and neonatal mortality
(or decreases in postnatal survival) were observed at doses > 5 mg/kg/day, with as much as a
30% increase in these effects seen at 10 and 20 mg/kg/day dose group. Postnatal survival and
viability at the 1 and 3 mg/kg/day dose groups was comparable to controls. At doses > 3
mg/kg/day, a trend in growth retardation (body weight reductions of 25-30%; p < 0.05), was
observed in the neonates at weaning. Body weights were at control levels by 6 weeks of age for
females, and by 13 weeks of age for males. A trend for increasing body weight (-6-10% greater
than controls) was observed in animals dosed with 5 mg/kg at 13 weeks and in animals dosed
with 1 and 3 mg/kg at 48 weeks. Deficits in early postnatal growth and development were also
manifested by significant (p < 0.05) delays in eye opening at doses > 5 mg/kg/day. Slight delays
(P < 0.05) in vaginal opening and in time to estrous were observed at 20 mg/kg/day in females;
in contrast significant accelerations (p < 0.05) in sexual maturation were observed in males, with
preputial separation occurring 4 days earlier than controls at the 1 mg/kg/day dose and 2-3 days
earlier in the 3-10 mg/kg/day dose groups, but the 20 mg/kg/day dose group was only slightly
delayed compared to controls. Under the conditions of the study, a LOAEL for developmental
toxicity of 1 mg/kg for males was indicated based on accelerated pubertal development, and a
NOAEL was not established. For females, the developmental LOAEL was 3 mg/kg based on
growth retardation, and the NOAEL was 1 mg/kg.
For maternal toxicity, BMD5 and BMDL5 estimates for decreases in maternal weight gain
during pregnancy were 6.76 and 3.58 mg/kg, respectively. For increases in maternal liver weight
at term, BMDs and BMDLs estimates were 0.20 mg/kg and 0.17 mg/kg, respectively. BMDs and
BMDL5 estimates for the incidence of neonatal mortality (determined by survival to weaning) at
5 mg/kg/day were 2.84 and 1.09 mg/kg, respectively. Significant alterations in postnatal growth
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and development were observed at 1 and 3 mg/kg/day, with BMDs and BMDLs estimates of 1.07
and 0.86 mg/kg, respectively, for decreased pup weight at weaning; and 2.64 and 2.10 mg/kg,
respectively, for delays in eye opening. The BMD5 and BMDL5 estimates for reduced phalangeal
ossification were < 1 mg/kg. BMD5 and BMDL5 estimates for reduced fetal weight at term were
estimated to be 10.3 and 4.3 mg/kg, respectively.
Male and female 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 (WT)] and PPARa-null mice were orally dosed from GD1-17 with 0,
0.1, 0.3, 0.6, 1, 3, 5, 10, or 20 mg PFOA/kg. Heterozygous (HET) litters were also produced by
mating WT and PPARa-null males with WT and PPARa-null dams to determine if genetic
background affected survival. The HET litters were sacrificed on PND15. Survival at birth was
recorded and live offspring counted and weighed by sex. Litters were counted and offspring
weighed on PND1-10, 14, 17, and 22. Weaning occurred on PND 22, and all dams and one
pup/litter were sacrificed. Blood was collected and the uteri were stained for implantation
counts.
There was no effect of treatment on maternal weight or maternal weight gain (excluding
non-pregnant females and those with full litter resorptions), number of implants, or pup weight at
birth. WT dams exposed to >0.6 mg/kg/day and 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 WT dams and 20 mg/kg/day in PPARa-null dams, 100% litter loss occurred.
Relative liver weight was significantly increased in WT adult females dosed with >1 mg/kg and
in PPARa-null adult females dosed with >3 mg/kg.
Body weight in WT offspring born of dams dosed with 1.0 mg/kg was significantly
reduced (p<0.05) compared to control offspring body weight gain on PND9, 10, and 22 (males)
and PND7 tolO and PND22 (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 WT litters exposed to >0.6 mg/kg, but was not
affected in PPARa-null litters. Survival was significantly decreased (p<0.05) for WT and HET
pups born to WT dams dosed with 1 mg/kg and for HET pups born to PPARa-null dams dosed
with 3 mg/kg. Offspring born of WT dams showed a dose-related trend for delayed eye opening
compared to control offspring (significantly delayed at 1 mg/kg/day, p<0.05), but no difference
in day of eye opening was observed in the offspring born of PPARa-null dams. At weaning,
relative liver weight was significantly increased (p<0.05) in WT offspring gestationally exposed
to >0.1 mg/kg and in 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 may
also contribute. Under the conditions of the study, the maternal/reproductive LOAEL for WT
mice was 0.6 mg/kg/day based on increased percentage of litter loss, and the NOAEL was 0.3
mg/kg/day. The developmental LOAEL for WT offspring was 0.1 mg/kg/day based on
increased liver weight, and the NOAEL was not established. The maternal LOAEL for PPARa-
null mice was 3 mg/kg/day based on increased liver weight, and the NOAEL was 1 mg/kg/day.
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The developmental LOAEL for 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 or 3 mg PFOA/kg on
GDs 1-17 by oral gavage (Albrecht et al., 2013). Controls received the water vehicle. Females
were either sacrificed on GDIS (n = 5-8/group) or allowed to give birth and then sacrificed,
along with their litters (n = 8-14), on PND20. Livers from dams, fetuses, and pups were weighed
and collected for histopathological evaluation and RNA analysis. Gene expression results are
given in section 4.3.4. Mammary gland whole mounts were prepared from female pups on
PND20 for quantification of ductal length and number of terminal end buds.
Evaluation on GD 18 showed no effects of PFOA administration on maternal body
weight, body weight gain, gravid uterine weight, number of implantations per dam, or number of
resorptions per litter in dams of any genotype. For animals allowed to litter, the average day of
parturition was slightly later in PFOA treated humanized mice compared with the controls.
Body weight of dams during lactation, the number of pups born per litter, pup body weight
during lactation, and the onset of pup eye opening were similar between treated and control
groups for all genotypes. Offspring survival during PNDs 1-5 was significantly reduced in the
wild-type PFOA treated group, but not the other genotypes.
Maternal liver weight was significantly increased in the treated groups of all genotypes
on GDIS and in wild-type animals on PND20. Maternal liver weight was not affected on
PND20 in the PPARa-null or PPARa-humanized mice. Relative fetal liver weight on GDIS was
significantly increased in fetuses from treated wild-type and humanized dams. On PND20,
relative liver weight was increased only in pups from treated wild-type dams. Microscopic
evaluation of the maternal liver showed centrilobular hepatocellular hypertrophy in all PFOA
treated groups on GDIS and PND20, with decreased incidence and severity by PND20. On
GDIS the liver lesions were graded as mild in the wild-type mice, minimal-to-mild in the
humanized mice, and minimal in the null mice. The morphological features of the liver lesions
differed slightly between genotypes and are described in more detail in section 4.4.1. Only wild-
type fetuses and pups from treated dams showed similar liver lesions.
The average length of mammary gland ducts and the average number of terminal end
buds per mammary gland per litter were not different between control and treatment groups from
all genotypes.
Yahia et al. (2010) gavage dosed pregnant ICR mice (n=5/group) with 0, 1, 5, or 10 mg
PFOA/kg from GDO-17 or 18. The dams dosed from GDO-17 were sacrificed on GDIS, and the
fetal skeletal morphology was evaluated. Dams dosed from GDO-18 were allowed to give birth
and their offspring were either processed for pathological examination or observed for 4 days for
neonatal mortality. Maternal liver, kidney, brain, and lungs were histologically examined after
necropsy. Serum was collected for clinical chemistry and lipid analysis. Body weight was
significantly decreased in dams receiving 10 mg/kg. Maternal absolute liver weight was
significantly increased (p<0.05) at doses > 5 mg/kg and relative liver and kidney weights were
significantly increased at all doses. Hepatic hypertrophy was localized to the centrilobular
region at the two lower doses and was diffuse at the highest dose. Renal cells in the outer
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medullar and proximal tubule were slightly hypertrophic at all doses. Treatment with 10 mg/kg
caused a significant increase in AST, ALT, GOT, and ALP and a significant decrease in total
serum protein, albumin, globulin, triglycerides, phospholipids, total cholesterol, and free fatty
acids. At 5 mg/kg, total serum protein and globulin were significantly decreased, and
phosphorus was increased. At 1 mg/kg, BUN and phosphorus were significantly increased. The
maternal LOAEL was 1 mg/kg based on significantly increased relative liver and kidney weight,
and no NOAEL was established.
Live fetal birth weight was significantly decreased at the two highest doses. There was
no difference in the percentage of live fetuses between treated and control groups. At 10 mg/kg,
increased incidence of cleft sternum, delayed phalanges ossification, and delayed eruption of
incisors was observed. Delayed parturition was observed in dams treated with 10 mg/kg, and
-58% of all pups born to those dams were stillborn. Death occurred within 6 hours of delivery
in the remaining pups, and whole body edema was observed in some of the pups. The body
weight of the live pups born to dams treated with 5 or 10 mg/kg was significantly reduced
compared to control pup body weight. By PND4, 16% of offspring born to dams dosed with 5
mg/kg had died. No pathological differences were observed in the lungs or brains of treated and
control offspring. The developmental LOAEL was 5 mg/kg based on decreased body weight and
decreased survival rate, and the NOAEL was 1 mg/kg.
Suh et al. (2011) examined placental prolactin-family hormone and fetal growth
retardation in mice treated with PFOA. Pregnant CD-I mice (n=10/group) were treated with 0,
2, 10, or 25 mg/kg PFOA from GDI 1 to GDI6. Dams were sacrificed on GDI6 and uteri were
removed and examined. Three placentas/group were analyzed using histochemistry and the
numbers of glycogen tropholblast cell (GlyT) in the junctional zone and sinusoidal trophoblast
giant cells (S-TGC) in the labyrinth zone were counted and compared. Trophoblast cells express
prolactin-family genes. mRNA from three placentas/group were analyzed using situ
hybridization, northern blot hybridization, and RT-PCR for mouse placental lactogen (mPL)-II,
prolactin like protein (mPLP)-E and F, Pit-la and P isotype (transacting factors of mPL and
mPLP genes).
A significant difference in maternal body weight was observed from GD13-16 in dams
treated with 25 mg/kg PFOA compared to controls. At >2 mg/kg PFOA, placental weight was
significantly decreased and the number of resorptions and dead fetuses was significantly
increased. At >10 mg/kg PFOA, fetal weight and the number of live fetuses were significantly
decreased. There were no differences in the number of implantation sites among the groups, and
post-implantation loss was 3.87, 8.83, 30.98, and 55.41% for the 0, 2, 10, and 25 mg/kg PFOA
groups, respectively.
The placentas of dams dosed with >10 mg/kg PFOA displayed necrotic changes. Parietal
and sinusoidal TGC and GlyT cell frequency in the placental junctional and labyrinth zones was
significantly decreased (>0.05) in a dose-dependent manner in treated dams. At 25 mg/kg
PFOA, S-TGCs showed signs of atrophy with crushed cell nucleus. A significant 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 express may be secondary to insufficient
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trophoblast cell differentiation and increased cell necrosis. These effects reduced placental
efficiency and contributed in part to fetal growth retardation.
Other Developmental Studies/Mammary Gland Development. The following studies used
study designs and/or examined endpoints not usually included in developmental studies. The
studies were conducted to determine critical periods of exposure focusing on mammary gland
development in dams and female offspring. Researchers focused on effects resulting from
indirect exposure of offspring via treatment of pregnant animals and direct exposure of
peripubertal animals starting about the time of weaning.
Indirect gestational and/or lactational exposures
An overview of experiments designed to assess the developmental effects of PFOA
exposure to CD-I mice during gestation is given in Table 4-12. Additional details of the studies
are described in the text below. All studies included measures of maternal and offspring survival
and body weight.
TABLE 4-12. Overview of studies in which pregnant CD-I mice were administered PFOA
Dose
(mg PFOA
/kg/day)
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
Timing
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 GD 7 and
continued to F2
generation
Endpoints
Body weight; mammary gland GD 18 (dams)
and PND 10 and 20 (dams, female pups)
Body weight; developmental landmarks and
growth to PND 189; mammary gland of
female pups up to 18 months
Body weight; developmental landmarks and
growth to PND 245; mammary gland of
female pups up to 18 months
Mammary gland of dams and female pups on
PND 1, 3, 5, 10
Liver weight; mammary gland of female
pups PND 7, 14, 21, 28, 42, 63, 84
Body weight; reproductive parameters;
mammary gland of FO, Fl, and F2 females
Reference
White et al.,
2007
Wolfetal.,
2007; White et
al., 2009
Wolfetal.,
2007; White et
al., 2009
White et al.,
2009
Macon et al.,
2011
White et al.,
2011
White et al. (2007) orally dosed pregnant CD-I mice with 0 or 5 mg PFOA/kg on
gestation days (GD) 1-17 (n=14), 8-17 (n=16), or 12-17 (n=16) to determine if decreased
neonatal body weights and survival were linked to gestational exposure or lactational changes in
milk quantity or quality. The control mice (n=14) were dosed with vehicle on GD1-17. A subset
of dams was sacrificed on GD 18. The remaining dams were allowed to give birth, and pups
were pooled and randomly redistributed among the dams of the respective treatment groups.
Litters were equalized to 10 pups per litter. Half of dams and litters were sacrificed on PND 10
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and remaining dams and litters were sacrificed on PND20. Mammary glands were collected
from dams and female pups at time of sacrifice.
Treatment with PFOA did not affect maternal weight gain, number of implants, or the
number of live fetuses. There was a significant increase (p<0.05) in prenatal loss in dams
exposed during GD1-17. Body weight of pups exposed gestationally to PFOA was significantly
decreased (p<0.05) at all time points measured. On GDIS, stunted alveolar development was
observed in the mammary gland of dams treated with PFOA on GD1-17 compared to the
mammary glands of the control dams which were saturated with milk-filled alveoli. Dams
treated with PFOA on GD 1-7 or 8-17 exhibited significant mammary gland epithelial
differentiation delays on PND10 as evidenced by an excessive amount of adipose tissue. In
comparison, mammary glands from control dams on PND10 were well differentiated, full of
alveoli filled with milk, and contained few apoptotic bodies and little adipose tissue. At PND20,
the mammary glands of all PFOA-treated dams were similar in appearance to the mammary
glands of control dams at PND10. Mammary gland differentiation in dams treated on GD12-17
was not statistically different from control dams.
The pups were also impacted by their in utero PFOA exposure. Their mammary glands
exhibited significantly stunted epithelial branching and longitudinal growth at PND10 and 20.
Very little mammary gland development occurred between PND 10-20 in the offspring of dams
exposed to PFOA even though postnatal growth and body weight gain increased in parallel to
that of the controls. Thus, at the only dose tested, 5 mg/kg, effects were observed on the dam
and pup mammary gland, decreased pup body weights and decreased survival for the pups
exposed diring GD 1-17.
In the study by Wolf et al. (2007), CD-I mice were orally dosed with 0 or 5 mg PFOA/kg
on GD7-17 (n=14), 10-17 (n=14), 13-17 (n=12), and 15-17 (n=12) or with 20 mg/kg on GD15-
17 (n=6) to determine if there was a specific window during which PFOA exposure produced
developmental effects. The developmental results from this study were published by Wolf et al.
(2007) and the mammary gland effects were published by White et al. (2009). On PND22, all
dams and one male and female pup from each litter were necropsied. Blood samples were
collected and livers from dams and offspring were removed and weighed. Uterine implantation
sites were counted. The 4th and 5th inguinal mammary glands were removed from female
offspring and analyzed at various intervals up to 18 months of age (White et al., 2009). The
remaining offspring were observed until PND 189.
Maternal weight gain was increased in dams exposed to PFOA beginning on GD7 and
10, but there was no affect on number of uterine implantation sites, litter loss, or number of pups
per litter at birth. Male pup weight at birth was significantly decreased (p<0.05) in dams dosed
with 5 mg/kg on GD7-17 or 10-17 or with 20 mg/kg on GDI5-17. By PND78, all male offspring
had recovered to control body weight levels. On PND161, the offspring of dams dosed during
GD13-17 weighed significantly more than control. Litters exposed to 20 mg/kg on GD15-17
experienced decreased survival (p<0.01) during postnatal days 1 through 22. Maternal relative
liver weight was significantly increased in all PFOA-treated dams except those treated during
GDI5-17. Relative liver weight in all male and female pups was significantly increased
(p<0.01). Eye opening and growth of body hair were delayed in pups exposed GD7-17 and 10-
17 (Wolf etal., 2007).
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Mammary gland developmental scores for all offspring of dams exposed to PFOA were
significantly lower compared to scores from offspring of control dams at PND29 and 32.
Delayed ductal elongation and branching and delayed appearance of terminal end buds were
characteristics of delayed mammary gland development at PND32. At 18 months of age,
increases in epithelial density and in density of darkly staining foci were observed in offspring of
PFOA-exposed dams. Based on histopathologic observation the darkly staining foci were
hypothesized to be hyperplasia of ductal epithelium, inflammatory cell influx into ductal regions,
increased stromal density, or altered differentiation of ductal epithelia. The average number of
foci per mammary gland was 1.5, 29.8, 17.9, 32.8, and 25.5 for the control, GD15-17, GD13-17,
GDI0-17, and GD7-17 groups, respectively (White et al., 2009). The 5 mg/kg dose was
associated with increased maternal and pup liver weight, altered pup mammary gland
development, and delayed pup eye opening and growth of body hair. The 20 mg/kg dose was
associated with decrease postnatal pup survival.
The objective of a second component of the Wolf et al. (2007)/White et al. (2009) study
was to determine if postnatal body weight deficits, neonatal lethality, and developmental delays
caused by PFOA exposure were the result of gestational exposure, lactational exposure, or a
combination of gestational and lactational exposure. Pregnant CD-I mice were orally dosed with
0 (n=48), 3 (n=28), or 5 (n=36) mg PFOA/kg on GDI through 17 and their offspring cross-
fostered at birth to create 7 treatment groups: control, in utero exposure only (3U and 5U),
lactation exposure only (PFOA stored in milk during gestation and released during lactation; 3L
and 5L), and in utero and lactation exposure (3U+L and 5U+L). On PND22, all dams and one
male and female pup from each litter were necropsied. Blood samples were collected and the
liver was removed from dams and offspring and weighed. Implantation sites were counted from
the uteri of dams. The 4th and 5th inguinal mammary glands were removed from female offspring
and analyzed at various intervals up to 18 months of age (White et al., 2009). The remaining
offspring were observed until PND245.
Maternal weight and weight gain were higher in PFOA treated dams compared to control
dams. Whole litter loss was significantly increased (p<0.05) at 5 mg/kg, but no differences in
the number of implantation sites were observed between the treated and control mice. Absolute
and relative livers weights of PFOA treated dams from both dose groups were significantly
increased (p<0.001) compared to absolute and relative liver weights of control dams 23 days
after the last dose (PND22). No difference in the number of live pups born per litter were found
between treated and control mice, but male and female pup birth weight was reduced (p<0.01) in
dams receiving 5 mg/kg (Wolf et al., 2007). The 3 mg/kg dose was a LOAEL for increases in
liver weight in the dams while 5 mg/kg was a LOAEL for the pups based on whole litter loss and
significantly reduced male and female birth weight.
A dose-dependent increase of PFOA was observed in the serum of dams treated with
PFOA providing a reservoir for lactational transfer. The control dams that nursed offspring
exposed in utero (3U and 5U) had low concentrations of PFOA in their serum from maternal
grooming behavior which allowed for low level lactational transfer.
Body weight of male and female pups (3U+L, 5U, and 5U+L) was significantly reduced
as early as PND2 and 1, respectively, and remained reduced throughout the lactation period.
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Body weight recovery to control levels was reached by male offspring within two weeks of
weaning, but recovery in female offspring in the 5U and 5U+L groups did not occur until after
PND85.
Postnatal survival in 5U+L pups was significantly decreased compared to control survival
beginning at PND4 and continuing throughout lactation. Survival in the other groups was not
different from control survival. Eye opening and body hair growth were significantly delayed in
the 3U+L, 5U, and 5U+L offspring. The relative liver weight was significantly increased in all
offspring regardless of exposure scenario (Wolf et al., 2007).
All female offspring of PFOA-exposed dams had reduced mammary gland
developmental scores at PND22 except for females in the 3L group. At PND42, mammary gland
scores from females in the 3U+L group were the only ones not statistically different from control
scores. This may have been due to inter-individual variance and multiple criteria used to
calculate mammary gland development scores. Despite the fact that the mammary gland scores
from the 3U+L group were not significantly different from scores of the control animals, they did
exhibit reduced branching and mammary fat pads with reduced parenchymal density. All
offspring of dams exposed to PFOA exhibited delayed mammary gland development at PND63
including those exposed only through lactation (3L and 5L). A higher density of dark staining
foci was observed in the mammary glands of these animals at 18 months of age (White et al.,
2009).
White et al. (2009) also reported the results from pregnant CD-I mice orally dosed with
0 (n=56) or 5 (n=56) mg PFOA/kg from GD8 through 17 to determine the timing of the
mammary gland development deficits observed following gestational or lactational exposure to
PFOA. The groups were cross-fostered at birth to create 4 treatment groups: control, in utero
exposure only (5U), lactation exposure only (5L), and in utero and lactation exposure (5U+L).
Dams and litters were sacrificed on PND1, 3, 5, and 10. Blood and liver samples were collected
for PFOA analysis. The 4th and 5th inguinal mammary glands were collected from dams and
female offspring and analyzed.
Maternal weight gain in treated dams was significantly higher compared to control
weight gain, but there were no effects of treatment on litter size or pup birth weight or weight at
PND1. Significantly decreased body weight occurred in the pups of the 5U+L group on PND3
and in all PFOA-exposed pups on PND 5 and 10. Relative liver weight of the treated dams was
significantly increased (p<0.05) compared to relative liver weight of control dams. On PND1,
liver to body weight ratios were significantly increased (p<0.05) in pups exposed in utero (5U,
5U+L); serum PFOA levels were 65,000-70,000 ng/mL. The liver to body weight ratio was
increased in pups exposed lactationally by PND5; serum PFOA levels were approximately
15,000 ng/mL (White et al., 2009).
On PND1, the mammary glands of PFOA-exposed dams were similar to glands seen in
late pregnancy, prior to parturition. In control dams nursing offspring from PFOA-exposed
dams, reduced alveolar filling was noted as early as PND3 presumably a result of exposure of the
dam from maternal grooming behavior. The delayed lactational morphology in dams treated
with PFOA and control dams nursing offspring from PFOA-treated dams was persistent up to
PND 10 (terminal sacrifice). Delayed mammary gland development was observed as early as
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PND1 in all female offspring from PFOA-exposed dams including those exposed through
lactation only (5L). Delayed mammary gland development persisted through the study duration
(White et al., 2009).
Macon et al. (2011) gavage dosed CD-I mice with 0, 0.3, 1.0, or 3.0 mg PFOA/kg from
GDI to GDI? (n=13 dams/group). Offspring were sacrificed on PND 7, 14, 21, 28, 42, 63, and
84, and blood, liver, brain, and the fourth and fifth mammary glands were collected from female
pups.
Body weight in male and female offspring was not affected through PND 84. Absolute
liver weight was significantly increased at >0.3 mg/kg in females and at >1.0 mg/kg in males on
PND 7, and at 3.0 mg/kg in females at PND14. Relative liver weight was significantly increased
at >0.3 mg/kg in males and females on PND7, at >1.0 mg/kg in females on PND14, and at 3.0
mg/kg in males and females on PNDs 14, 21, and 28. No dose-related differences were observed
absolute and relative brain weights.
Delayed mammary gland development of female pups was observed as early as PND7 at
>1.0 mg/kg and PND14 at >0.3 mg/kg and persisted until the end of the study. The delayed
development was characterized by reduced epithelial growth and the presence of numerous
terminal end buds at PND 63 and 84. The LOAEL was 0.30 mg/kg based on significantly
increase liver weight and delayed mammary gland development. No NOAEL was established.
Macon etal. (2011) also gavage dosed CD-I mice with 0, 0.01, 0.1, or 1.0 mg PFOA/kg
from GD10 to GDI7 (n=5-8 dams/group) to examine the effects of low doses of PFOA on
mammary gland development. Female offspring were sacrificed on PND1, 4, 7, 14 and 21, and
blood, liver, and the fourth and fifth mammary glands were collected.
No differences in body weight or brain weight were observed for male or female
offspring. At 1 mg/kg, absolute and relative liver weights were significantly increased at PNDs
4 and 7. Relative liver weight was also significantly increased at PND14. Mammary gland
development was delayed by exposure to PFOA, especially longitudinal epithelial growth. At
PND21, all treatment groups had significantly lower developmental scores. At the highest dose,
poor longitudinal epithelial growth and decreased number of terminal end buds were observed.
The LOAEL was 0.01 mg PFOA/kg based on delayed mammary gland development. No
NOAEL was established.
White et al. (2011) examined the extended consequences of PFOA-induced altered
mammary gland development in a multigenerational study in CD-I mice. Pregnant mice (FO,
n=10-12 dams/group) were gavage dosed with 0, 1, or 5 mg PFOA/kg from GDI-17. A separate
group of pregnant mice (n=7-10 dams/group) were gavage dosed with either 0 or 1 mg PFOA/kg
from GD1-17 and received drinking water containing 5 ppb PFOA beginning on GD7. 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 breeding and early gestation, to
simulate a chronic low-dose exposure. FO females were sacrificed on PND22. Fl offspring
were weaned on PND22 and bred at 7-8 weeks of age. F2 litters were maintained through
PND63. Groups of Fl and F2 offspring (n=l-2 offspring/litter from 5-7 litters/group) were
sacrificed on PND 22, 42, and 63. A group of F2 offspring (n=6-10/group) was also sacrificed
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on PND10. A lactational challenge experiment was performed on PND10 with Fl dams and F2
offspring.
FO Females. Gestational exposure to 5 mg PFOA/kg significantly increased prenatal loss in FO
mice and significantly decreased the number of live offspring and the postnatal survival of the
viable pups. Weight gain and number of implants did not differ among the groups. At PND 22,
control FO dams displayed weaning-induced mammary involution. At PND22, 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 (-PND10).
Fl Generation. Offspring body weight at PND42 was significantly reduced for those whose
dams received 5 mg/kg; at PND63, body weight was significantly reduced at 1 mg/kg + 5 ppb in
the drinking water. Liver-to-body weight ratios were significantly increased at 1 mg/kg on
PND22 and at 5 mg/kg on PNDs 22 and 42.
Fl Females. There was no indication of toxicity in Fl females. Exposure to PFOA did not
affect prenatal loss or postnatal survival, although Fl females that had been gestationally
exposed to 5 mg/kg, had significantly fewer number of implants. On PNDs 22, 42, and 63, the
mammary glands of all mice exposed to PFOA were significantly delayed in development. Fl
females were bred and examined on LD10 and 22. On LD10, delayed mammary gland
development (lactation morphology) was observed in all groups. Although, morphological
differences were observed, there were no differences among the groups in the lactational
challenge experiment. In the lactational challenge experiment, dams were removed from their
litters for 3 hours, then returned to their litters and allowed to nurse for 30 minutes. The time
from the dam's return to the litter and nursing initiation was recorded. The litters were weighed
before and after nursing to estimate volume of milk produced. By LD22, only Fl dams in the 1
mg/kg + 5 ppb PFOA and 5 mg/kg groups displayed delayed mammary gland development.
F2 Generation. A significant reduction in body weight was observed in control + 5 ppb drinking
water PFOA offspring on PND 1, but there was no difference by PND 3. Offspring from the 1
mg/kg + 5 ppb drinking water PFOA group had a significant increase in weight compared to
controls on PNDs 14, 17, and 22. Liver-to-body weight ratios were not different among the
groups. Mammary gland development in groups with 5 ppb PFOA in the drinking water was
significantly delayed at PND42. Mammary glands in the remaining groups displayed delayed
development, but control F2 females had unusually low scores at PNDs 10 and 22 which may
have reduced the ability to detect statistically significant effects in other treatment groups.
In summary, this study showed delay mammary gland development in dams and female
pups exposed continuously to a concentration as low as 5 ppb in the drinking water for three
generations. Direct dosing of dams with 1 or 5 mg/kg during gestation also resulted in delayed
mammary gland development in dams and their offspring at both doses as well as reduced pup
survival and decreased body weight at 5 mg/kg.
Direct peripubertal exposures
Yang et al. (2009) gavage dosed 21 day old female BALB/c mice (5/group) with 0, 1, 5,
or 10 mg PFOA/kg for 5 days/week for 4 weeks to determine the effects of peripubertal PFOA
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exposure on puberty and mammary gland development. At necropsy, uteri and livers were
weighed and processed for histological examination. Mammary glands were collected and
processed for histological and whole mount examination. A significant decrease in body weight
was observed following exposure to 10 mg/kg. The mammary glands of female BALB/c mice
treated with 5 or 10 mg/kg had reduced ductal length, decreased number of terminal end buds,
and decreased stimulated terminal ducts compared to the mammary glands of control mice.
BrdU incorporation into the mammary gland revealed a significantly lower number of
proliferating cells in the ducts and terminal end buds/terminal ducts at 5 mg/kg (not tested at 10
mg/kg). Absolute and relative liver weight was significantly increased in all treated BALB/c
mice. The absolute and relative uterine weight was significantly decreased in all treated mice
compared to uterine weight in control mice. Vaginal opening was significantly delayed in mice
dosed with 1 mg/kg, and did not occur at 5 or 10 mg/kg. Under the conditions of this study, the
LOAEL was 1 mg/kg/day based on delayed vaginal opening, increased liver weight and
decreased uterine weight, and no NOAEL was established.
Yang et al. (2009) also dosed 21 day old female C57BL/6 mice in the same manner as the
BALB/c mice and examined the effects of PFOA on mammary gland development and vaginal
opening. The body weight effects were similar in both strains with 10 mg/kg causing
significantly reduced body weight. At 5 mg/kg, PFOA had a stimulatory effect on the mammary
glands. There was a significant increase in the number of terminal end buds and stimulated
terminal ducts. Ductal length was not affected. Mammary gland development was inhibited in
mice dosed with 10 mg/kg with no terminal end buds or stimulated terminal ducts present and
very little ductal growth. Absolute and relative liver weight was significantly increased in all
treated mice. The absolute and relative uterine weight was significantly increased in C57BL/6
mice dosed with 1 mg/kg and significantly decreased in C57BL/6 mice dosed with 10 mg/kg.
There was no difference in uterine weights between mice treated with 5 mg/kg and control mice.
Vaginal opening was delayed in C57BL/6 mice dosed with 5 mg/kg and did not occur in mice
dosed with 10 mg/kg. Under the conditions of this study, the LOAEL was 1 mg/kg/day based on
and increased liver and uterine weights, and no NOAEL was established.
Zhao et al. (2010) conducted several experiments in C57BL/6 mice to determine the
potential mechanism by which peripubertal PFOA exposure resulted in the stimulation of
mammary gland development observed by Yang et al. (2009). In experiments to determine if
PFOA has a hormonal effect on mammary gland development, C57BL/6 mice (n=10/group)
were ovariectomized (OVX) at 3 weeks of age, allowed one week to recover, and treated with 0
or 5 mg PFOA/kg bw for 4 weeks. Abdominal and inguinal mammary glands were collected at
sacrifice, prepared as wholemounts, and scored for growth and development. The mammary
glands of the OVX control and PFOA-treated OVX mice were similarly stunted in growth as
evidenced by no outgrowth of ducts or presence of terminal end buds. This was in contrast to the
stimulatory effect of PFOA observed by Yang et al. (2009) in intact mice.
In experiments to determine if PFOA-affected mammary glands respond to hormone
treatment, intact C57BL/6 mice were dosed with 0 or 5 mg/kg bw PFOA for 4 weeks starting at
21 days of age. After the last dose, the mice were ovariectomized, allowed to recover for 1
week, and injected s.c. for 5 days with 17p-estradiol (E, 1 |ig/0.2 ml/mouse), progesterone (P, 1
mg/0.2 ml/mouse), or both (E+P, 1 |ig+l mg/0.2 ml/mouse). The mice were sacrificed 24 hours
after the last hormone injection. Abdominal and inguinal mammary glands were collected at
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sacrifice, prepared as whole mounts, and scored for growth and development. In the mammary
glands of mice treated with PFOA and E, stimulated terminal ducts were observed, and in
PFOA-treated mice given P or E+P stimulated terminal ducts and an increased number of side
branches were observed. The results showed that PFOA increased the mouse mammary gland
response to exogenous estrogen and progesterone.
In experiments to determine if PFOA-induced mammary gland development stimulation
is related to PPARa expression and to determine the impact of PFOA on steroid hormones and
growth factors, female C57BL/6 and PPARa-null C57BL/6 mice (n=5-10 mice/group) were
gavage dosed with 0 or 5 mg/kg bw PFOA 5 days/weeks for 4 weeks starting at 21 days of age
(Zhao et al., 2010). Vaginal opening was monitored daily and estrous cycle state was determined
at sacrifice after 4 weeks of treatment. At necropsy, blood was collected for measurement of
serum steroid hormones and binding proteins. Portions of the mammary glands, ovaries, and
livers were collected and processed for histological examination. RNA was extracted from the
livers for quantitative RT-PCR and PCR array for selected genes related to metabolism of drugs,
toxic chemicals, hormones, and micronutrients. Portions of mammary gland were used in
western blot analysis of several enzymes, local growth factors, and receptors. Among those were
aromatase which aids in converting testosterone to estradiol and androstenedione to estrone,
hydroxysteroid 1?P dehydrogenase 1 (HSD17P1) which aids in converting estrone to estradiol,
and hydroxysteroid 3p dehydrogenase 1 (HSD3P1) which aids in converting pregnenolone to
progesterone and androstenediole to testosterone. Growth factors critically involved in
mammary gland development including amphiregulin (Areg), insulin like growth factor I (IGF-
I), and hepatocyte growth factor (HGFa), and markers of cell proliferation including cyclin Dl
and proliferating cell nuclear antigen (PCNA) were analyzed by western blot. Areg mediates
estrogen receptor a (ERa) function and is a ligand for epidermal growth factor receptor (EGFR).
These receptors were also 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
progesterone levels were significantly increased (p<0.05) in PFOA-treated mice in proestrus and
estrus compared to control mice in the same stages of the estrous cycle. Serum sex hormone
binding globulin and albumin levels were not significantly changed by treatment with PFOA.
The effect of PFOA on aromatase, HSD17P1, and HSD3P1 activity in the ovaries of
C57BL/6 PPARa-null mice was examined. In C57BL/6 mice, HSD17P1 and HSD3P1 proteins
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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 may increase serum steroid hormone levels in the ovaries.
Due to the increased progesterone levels observed in PFOA-treated mice, the expression
of liver metabolic enzymes was analyzed. Liver metabolic function may affect steroid hormone
serum levels which play a role in mammary gland development. In 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 but not limited to CYPlal, CYPla2, and HSD17P2, was significantly down-regulated
(p<0.05) in C57BL/6 mice treated with PFOA, but expression in PFOA-treated PPARa-null
mice was comparable to 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 may
not contribute to PFOA-induced mammay gland development stimulation.
Inhalation Exposure
There were two components (inhalation and oral) to the Staples et al. (1984)
developmental toxicity study of PFOA in Sprague-Dawley rats. The oral component was
presented earlier. The study design consisted of whole-body dust inhalation exposure for 6
hours/day, on GD 6-15. The MMAD ranged from 1.4 to 3.4 jim and the geometric standard
deviation (GSD) ranged from 4.3 to 6.0. The study was carried out in two trials with each trial
including two experiments. In experiment 1, the dams were sacrificed on GD21 prior to
parturition, and in experiment 2, the dams were allowed to litter and were sacrificed on PND23;
offspring were sacrificed on PND35. In the first trial (both experiment), dams (n=12) were
exposed to 0, 0.1, 1, or 25 mg/m3. In trial 2, the high dose was reduced to 10 mg/m3. In
experiment 1 of trial 2, dams numbered 12-15/group and two additional groups (6 dams/group)
were added and were pair-fed to the 10 and 25 mg/m3 groups. In experiment 2 of trial 2, only 6
control and 6 dams dosed at 10 mg/m3 were allowed to litter.
In the first experiment, the dams were weighed on GD 1, 6, 9, 13, 16, and 21 and
observed daily for abnormal clinical signs. On GD 21, the dams were sacrificed by cervical
dislocation and examined for any gross abnormalities, liver weights were recorded and the
reproductive status of each animal was evaluated. The ovaries, uterus and contents were
examined for the number of corpora lutea, live and dead fetuses, resorptions and implantation
sites. Pups (live and dead) were counted, weighed and sexed and examined for external, visceral,
and skeletal alterations. The heads of all control and high-dosed group fetuses were examined for
visceral alterations as well as macro- and microscopic evaluation of the eyes.
Treatment-related clinical signs of maternal toxicity occurred at 10 and 25 mg/m3 and
consisted of wet abdomens, chromodacryorrhea, chromorhinorrhea, a general unkempt
appearance, and lethargy in four dams at the end of the exposure period (high-concentration
group only). Three out of 12 dams died during treatment at 25 mg/m3 (on GD 12, 13, and 17).
Food consumption was significantly reduced at 10 and 25 mg/m3; however, no significant
differences were noted between treated and pair-fed groups. Significant reductions in body
weight were also observed at these concentrations, with statistical significance at the high-
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concentration only. Likewise, statistically significant increases in mean liver weights (p < 0.05)
were seen at the high-concentration group. Under the conditions of the study, a NOAEL and
LOAEL for maternal toxicity of 1 and 10 mg/m3, respectively, were indicated.
No effects were observed on the maintenance of pregnancy or the incidence of
resorptions. Mean fetal body weights were significantly decreased in the 25 mg/m3 PFOA group
(p = 0.002) and in the pair-fed control group (p = 0.001). Interpretation of the decreased fetal
body weight is difficult given the high incidence of mortality in the dams. Under the conditions
of the study, a NOAEL and LOAEL for developmental toxicity of 10 and 25 mg/m3,
respectively, were indicated.
In the second experiment, in which the dams were allowed to litter, the procedure was the
same as that for the first experiment until GD 21. Two days before the expected day of
parturition, each dam was housed in an individual cage. The date of parturition was noted and
designated PND 1. Dams were weighed and examined for clinical signs on PNDs 1, 7, 14, and
22. On PND 23 all dams were sacrificed. Pups were counted, weighed, and examined for
external alterations. At birth, each pup was subsequently weighed and inspected for adverse
clinical signs on PNDs 4, 7, 14, and 22. The eyes of the pups were also examined on PNDs 15
and 17. Pups were sacrificed on PND 35 and examined for visceral and skeletal alterations.
In the second experiment, clinical signs of maternal toxicity seen at 10 and 25 mg/m3
were similar in type and incidence to those described for trial one. Maternal body weight gain
during treatment at 25 mg/m3 was less than controls, although the difference was not statistically
significant. In addition, 2 out of 12 dams died during treatment at 25 mg/m3. No other treatment-
related effects were reported, nor were any adverse effects noted for any of the measurements of
reproductive performance. Under the conditions of the study, a NOAEL and LOAEL for
maternal toxicity of 1 and 10 mg/m3, respectively, were indicated.
Signs of developmental toxicity in this group consisted of statistically significant
reductions in pup body weight on PND1 (6.1 g at 25 mg/m3 vs. 6.8 g in controls, p=0.02). On
PNDs 4 and 22, pup body weights continued to remain lower than controls, although the
difference was not statistically significant. No significant effects were reported following
external examination of the pups or with ophthalmoscopic examination of the eyes. Under the
conditions of the study, a NOAEL and LOAEL for developmental toxicity of 10 and 25 mg/m3,
respectively, were indicated.
Dermal Exposure
No data on the developmental effects of dermal exposures to PFOA were identified in the
literature.
4.2.6 Chronic Toxicity
Oral Exposure
Monkey. Male Cynomolgus monkeys (n=4 or 6/dose) were administered PFOA by oral capsule
containing 0, 3, 10, or 30/20 mg/kg/day for 26 weeks (Thomford, 2001b, Butenhoff et al., 2002).
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Dosing of animals in the 30 mg/kg-day dose group ceased after 12 days and decreased to 20
mg/kg-day when reinstated on Day 22 because of low food consumption, decreased body weight,
and decreased feces. Sacrifice of the surviving monkeys, except for two control monkeys and
two monkeys from the mid dose group (recovery animals) occurred at 26 weeks. The animals in
the recovery groups were sacrificed 13 weeks later.
Animals were observed twice daily for mortality and moribundity and were examined at
least once daily for signs of poor health or abnormal behavior. Ophthalmic examinations were
performed before treatment began and at weeks 26 and 40. Body weight, food consumption,
clinical hematology, clinical chemistry, urinalysis, serum hormone levels, and PFOA levels in
blood and tissue were assessed throughout the study. One animal from the 30/20 mg/kg-day
dose group was sacrificed in moribund condition on day 29 with signs of dosing injury and liver
lesions. One animal from the 3 mg/kg-day dose was sacrificed (day 137) with signs of hind limb
paralysis, ataxia and hypoactive behavior, few feces and no food consumption. Treatment of the
remaining three animals given 30/20 mg/kg-day was halted on days 43, 66, and 81, respectively,
because of thin appearance, few or no feces, low or no food consumption, and weight loss, but
the animals appeared to recover from compound-related effects within 3 weeks after cessation of
treatment. No significant changes in mean body weight were observed at doses of 3 or 10
mg/kg/day.
Serum hormone levels (estrone, estradiol, estriol, testosterone, TSH, free T4, total T4
and CCK) were not significantly altered throughout the study. However, free and total T3 levels
were significantly decreased (p<0.05) from weeks 5 to 10 and at week 27 in the 30/20 mg/kg/day
dose group compared to controls.
At terminal sacrifice (26 weeks), mean absolute liver weight was significantly increased
in all dose groups, and the relative liver-to-body weight ratio was significantly increased for the
high dose group. Final body weight and liver weight data are presented in Table 4-13. The
cause of the increase in liver weights was suggested to be due to hepatocellular hypertrophy
(indicated by decreased hepatic DNA content) which was hypothesized to result from
mitochondrial proliferation based on an increase in hepatic succinate dehydrogenase activity.
TABLE 4-13. Liver weight data in monkeys administered PFOA for 6 months
Dose
Omg/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*
From Butenhoff et al. 2002
* Significantly different from control, p<0.01.
Because administration of PFOA to rats has been shown to result in liver, Ley dig cell and
pancreatic acinar cell tumors, Butenhoff et al. (2002) analyzed markers of tumor formation in the
monkey study just described. In the liver, a two-fold increase in hepatic palmitoyl CoA oxidase
activity was observed in the 30/20 mg/kg-day group, consistent with reports for species that are
not particularly responsive to PPARa-agonists. Replicative DNA synthesis in the liver, an
indication of cell proliferation, was not altered in the treated animals. It has also been proposed
that changes associated with the pancreatic acinar cell tumors in rats include increased serum
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CCK concentrations and indications of cholestasis, including increases in alkaline phosphatase,
bilirubin, and bile acids. None of these changes were observed in the Cynomolgus monkeys.
There were also no significant changes in estradiol, estriol, or testosterone in the monkeys. Each
of these factors is associated with Ley dig cell tumors in rats. There were no changes in
replicative DNA synthesis in the pancreas or testes.
After a 13 week recovery period, there were no treatment-related effects on terminal body
weights or on absolute or relative organ weights, suggesting that the treatment-related liver
weight changes were reversible. There were no treatment-related macroscopic or microscopic
changes at the recovery sacrifice. Under the conditions of the study, the LOAEL was 3 mg/kg-
day based on increased liver weight and a NOAEL was not established.
Rat. The chronic toxicity of PFOA was investigated in a 2 year study in rats by Butenhoff et al.
(2012). Sprague-Dawley (Crl:CD BR) rats (50/sex) were fed diets containing 0, 30 or 300 ppm
PFOA (0, 1.3, and 14.2 mg/kg-day for males; 0, 1.6, and 16.1 mg/kg-day for females). Groups of
15 additional rats per sex were fed 0 or 300 ppm PFOA and evaluated at the one year interim
sacrifice. All animals were observed daily throughout the two year dosing period. Periodic
observations included body weights and feed consumption, hematology, serum chemistry,
urinalysis, gross pathology, organ weights, and histopathology. Animals were sacrificed after
one and two years of dosing. Organ weights were determined after each sacrifice and the tissues
subjected to histological examination.
There were dose-related decreases in body weight gains in male and female rats as
compared to the controls; statistical significance was attained for only the high-dose group (both
sexes). The body weight changes were considered treatment related since feed consumption
increased (rather than decreased) during the study. There were no significant differences in
mortality between the treated and untreated groups. Significant decreases in red blood cell
counts, hemoglobin concentrations and hematocrit values were observed in the high-dose male
and female rats as compared to control values. Clinical chemistry changes included slight (less
than 2-fold) but significant increases in ALT, AST, and AP in both treated male groups from 3-
18 months, but only in the high-dose males at 24 months. No dose- or treatment-related
differences in absolute and relative organ weights were found between the treated and control
groups at 2 years.
Significantly increased incidences of lesions in the liver were observed in the high-dose
male group. At 1 year, diffuse hepatomegalocytosis, portal mononuclear cell infiltration and
hepatocellular necrosis were seen. At 2-years, significant increases in megalocytosis were
observed in the males, and females in the high-dose group. Hepatic cystoid degeneration, a
condition characterized by areas of multilocular microcysts in the liver parenchyma, and
hyperplastic nodules, a localized proliferation of hepatic parenchymal cells, were also
significantly increased in high-dose males.
Among the high dose males histological changes were noted in tissues other than the
liver. A significant increase in aspermatogenesis and vascular mineralization of the testes and the
incidence of pulmonary hemorrhage (44%) and alveolar macrophates (62%) were reported.
The LOAEL for male rats is 300 ppm (14.2 mg/kg/day) based on a decrease in body
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weight gain and histological changes in the liver, testes, and lungs. The LOAEL for female rats
is 300 ppm (16.1 mg/kg/day) based on decreased body weight gain and hematologic effects. The
NOAEL for both sexes is 30 ppm (1.3 mg/kg/day and 1.6 mg/kg/kg for females).
Biegel et al. (2001) conducted a 2-year mechanistic study in which male Crl:CD BR
(CD) rats (n=156/group) were fed a diet containing 0 or 300 ppm PFOA (0 or 13.6 mg/kg/day).
Interim sacrifices were conducted every three months up to 21 months for measurements of liver
and testes weights, peroxisome proliferation, and cell replication. Serum samples were collected
and reproductive hormones measured.
Body weight was significantly decreased from days 8 to 630 in PFOA-exposed rats. In
the treated group, relative liver weights and hepatic 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% vs. 14% control group)
Pancreatic aceinar cell proliferation was significantly increased at 15, 18, and 21 months. The
incidence of acinar cell hyperplasia was 30/76 (39%) compared to the incidence in the control
group, 14/80 (18%). There were no significant differences in serum testosterone or prolactin in
the PFOA-treated rats when compared to the controls. Serum FSH was significantly increased at
6 months, and LH was significantly increased at 6 and 18 months. There were significant
increases in serum estradiol concentrations in the treated rats at 1, 3, 6, 9, and 12 months.
4.2.7 Carcinogenicity
Oral Exposure
Tissues from the animals in the Butenhoff et al. (2012) study were evaluated for
neoplastic and preneoplastic formations. Hepatocelluar carcinomas were observed at 6, 2, and
10% in the control, low-, and high-dose male rats, respectively. None were observed in females
in the control and low-dose groups, but a 2% incidence was observed for female rats in the high-
dose group. The differences between the treated and control groups were not significantly
different. No liver adenomas were observed.
At the one-year sacrifice, testicular masses were found in 6/50 high-dose and 1/50 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 Leydig cell tumor incidence in the control, low-, and high-dose groups was 0/50 (0%), 2/50
(4%), and 7/50 (14%), respectively. The increase was also statistically significant when
compared to the historical control incidence of 0.82% observed in 1,340 Sprague-Dawley control
male rats used in 17 carcinogenicity studies (Chandra et al., 1992). The spontaneous incidence
of Leydig cell tumors in 2-year old Sprague-Dawley rats in other studies was reported to be
approximately 5% (cited in Clegg et al., 1997).
A statistically significant, dose-related increase in the incidence of ovarian tubular
hyperplasia was found in female rats at the 2-year sacrifice. The incidence of this lesion in the
control, low-, and high-dose groups was 0%, 14%, and 32%, respectively. The biological
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significance of this effect at the time of the initial evaluation was unknown, as there was no
evidence of progression to tumors.
Slides of the ovaries from the Butenhoff et al. (2012) study were re-evaluated by Mann
and Frame (2004) with emphasis placed on the proliferative lesions of the ovary. Using more
recently published nomenclature, the ovarian lesions were diagnosed and graded as gonadal
stromal hyperplasia and/or adenomas, which corresponded to the diagnoses of tubular
hyperplasia or tubular adenoma by the original study pathologist. The data are summarized in
Table 4-14. No statistically significant increases in hyperplasia (total number), adenomas, or
hyperplasia/ adenoma combined were seen in treated groups compared to controls. There was
some evidence of an increase in size of stromal lesions observed at the 300 ppm group; however,
adenomas occurred in greater incidences in the control group than in either of the treated groups.
Results of this follow-up evaluation indicated that rats sacrificed at the one-year interim
sacrifice, as well as rats that died prior to the interim sacrifice were not considered at risk for
tumor development.
TABLE 4-14. Incidence of Ovarian Stromal Hyperplasia and Adenoma in Rats
Group
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
From Mann and Frame, 2004
A significant increase (p<0.05) in the incidence of mammary fibroadenomas in the low
and high dose groups of female rats was noted in the Butenhoff et al. (2012) study. The
incidence of mammary fibroadenomas was 22% (11/50), 38% (19/50) and 42% (21/50) in the
control, 30, and 300 ppm groups, respectively (Hardisty, 2010). The increase was also
statistically significant when compared to the historical control incidence of 19.0% observed in
1,329 Sprague-Dawley control female rats used in 17 carcinogenicity studies as reported by
Chandra et al. (1992). Butenhoff et al. (2012) did not consider the mammary fibroadenomas to
be treatment related on the basis of the historical control incidence (24%) from a study of 181
female rats terminally sacrificed at 18 months (Prejean et al., 1973). However, this is an
inappropriate historical reference because the study duration was 18 months instead of 2 years.
When the mammary fibroadenoma incidences were compared to the historical control incidence
(37%) in 947 female rats in the Haskell Laboratory, however, there did not appear to be any
compound related effect (Sykes, 1987). The incidence of other mammary gland tumor types is
provided in Table 4-15.
The mammary gland findings were re-examined by a Pathology Working Group
(Hardisty et al., 2005, 2010; Table 4-15) using the diagnostic criteria and nomenclature of the
Society of Toxicological Pathologists (Mann et al., 1996). The Pathology Working Group
(PWG) concluded that there were no statistically significant differences in the incidence of
fibroadenoma, adenocarcinoma, total benign neoplasms (lobular hyperplasia, fibroadenoma, and
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adenoma) or total malignant neoplasms (adenocarcinoma) of the mammary glands between
control and treated animals using Fischer's Exact Test for pair-wise comparision (see Table 4-
15). There was also no significant difference in combined benign and malignant neoplasms
between control and treated groups.
The primary difference between the original reported findings and the PWG results
involved findings initially reported as lobular hyperplasia being reclassified as fibroadenomas,
mostly in the control group. With the reclassification, the incidence of single mammary
fibroadenoma in the control, low- and high-dose groups were: 32% (16/50), 32% (16/50), and
40% (20/50), respectively. The evidence for mammary fibroadenomas in the female rat is
equivocal since the incidences were comparable to some historical background incidences, so the
PWG concluded that there were no statistically significant differences in the incidence of
fibroadenoma or other neoplasms of the mammary gland between control and treated animals
(Hardistyetal., 2005, 2010).
TABLE 4-15. Mammary Gland Tumor Incidence Comparison
Number
necropsied
Lobular
hyperplasia (%)
Adenocarcinoma
(%)
Fibroadenoma*
(%)
Adenoma (%)
0 ppm
Butenhoff
50
6
(12%)
8
(16%)
10
(20%)
o
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
Butenhoff
50
2
(4%)
5
(10%)
21
(42%)
0
(0%)
ppm
Hardisty
50
0
(0%)
5
(10%)
23
(46%)
0
(0%)
From Hardisty et al., 2005, 2010 (Compares Butenhoff et al., 2012 with Hardisty et al., 2005,2010)
*Includes fibroadenoma, multiple counts
A 2-year mechanistic study in male Crl:CD BR (CD) rats (Cook et al., 1994; Biegel et al.,
2001) resulted in Ley dig cell tumors supporting the observations from the Butenhoff et al. (2012)
study. The rats (n=156/group) were fed diets containing 0 ppm (ad libitum control and control
pair fed to the PFOA-exposed rats) or 300 ppm PFOA (13.6 mg/kg intake). Rats were
euthanized at interim time points of 1, 3, 6, 9, 12, 15, 18, and 21 months. All rats surviving the
24-month test period were necropsied for microscopic examination of various organs: e.g.,
kidneys, liver, testes, brain, heart, spleen.
There was a significant increase in the incidence of Ley dig cell adenomas in the treated
rats 11% (8/76) when compared to the pair-fed control rats (3%, 2/78). The incidence in ad
libitum control rats was 0% (0/80). In addition, the treated group had a significant increase in the
incidences of liver adenomas and pancreatic acinar cell adenomas when compared to the pair-fed
and ad libitum control groups. The incidence of liver adenomas in the ad libitum control, pair-fed
control, and treated groups was 3% (2/80), 1% (1/79), and 13% (10/76), respectively. The
incidence for liver carcinomas was 0% (0/80), 3% (2/79), and 0% (0/76) in the ad libitum
control, pair-fed control, and treated groups, respectively. The incidence for the pancreatic
acinar cell adenomas was 0% (0/80), 1% (1/79), and 9% (7/76). The incidence of pancreatic
acinar cell carcinoma was 1% (1/76) in the treated rats, 0% (0/79) in the pair-fed control rats, and
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0% (0/80) in control rats.
In the first carcinogenicity study (Butenhoff et al., 2012), there was no reported increase
in the incidence of pancreatic acinar cell tumors, and the incidence of pancreatic acinar
hyperplasia in the male rats was 0/33, 2/34, and 1/43 in the control, 30, and 300 ppm groups,
respectively. To resolve this discrepancy, the histological slides from both studies were reviewed
by independent pathologists. This review of the microscopic lesions of the pancreas in the two
studies indicated that PFOA produced increased incidences of proliferative acinar cell lesions of
the pancreas in the rats of both studies at the dietary concentration of 300 ppm. The differences
observed were quantitative rather than qualitiative; more and larger focal proliferative acinar cell
lesions and greater tendency for progression of lesions to adenoma of the pancreas were
observed in the Biegel et al. (2001) study compared to the Butenhoff et al. (2012) study. The
difference between pancreatic acinar hyperplasia (Butenhoff et al., 2012) and adenomas (Biegel
et al. 2001) in the rat was a reflection of arbitrary diagnostic criteria and nomenclature by
different pathologists. The basis for the quantitative difference in the lesions observed is not
known but was believed to be due most likely to difference in the diets used in the two
laboratories (Frame and McConnell, 2003).
Inhalation Exposure
No data on the tumorigenic effects of chronic inhalation exposures to PFOA were
identified in the literature.
Dermal Exposure
No data on the tumorigenic effects of chronic dermal exposures to PFOA were identified
in the literature.
4.3 Other Key Data
4.3.1 Mutagenicity and Genotoxicity
PFOA has been tested for genotoxicity in a variety of in vivo and in vitro assays. The
data from the in vitro studies are summarized in Table 4-16.
PFOA was tested in a cell transformation and cytotoxicity assay conducted in
mouse embryo fibroblasts. The cell transformation was determined as both colony
transformation and foci transformation. There was no evidence of transformation at any of the
dose levels tested in either the colony or foci assay methods (Garry & Nelson, 1981).
PFOA was tested twice (Lawlor, 1995; 1996) for its ability to induce mutation in the
Salmonella - E. 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, 1996c; NOTOX,
2000). Sadhu (2002) reported that PFOA did not induce gene mutation when tested with or
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without metabolic activation in the K-l line of Chinese hamster ovary (CHO) cells in culture.
Murli (1996b,d) tested PFOA twice for its ability to induce chromosomal aberrations in
CHO cells. In the first assay, PFOA induced both chromosomal aberrations and polyploidy in
both the presence and absence of metabolic activation. In the second assay, no significant
increases in chromosomal aberrations were observed without activation. However, when tested
with metabolic activation, PFOA induced significant increases in chromosomal aberrations and
in polyploidy (Murli, 1996b). The effects were observed only at toxic concentrations (EFSA,
2008).
PFOA did not display mutagenic activity with or without metabolic activation in S.
typhimurium strains TA98, TA100, TA102, and TA104 (Freire et al., 2008).
TABLE 4-16. Genotoxicity of PFOA In Vitro
Test System
C3H10T,/2 mouse
embryo fibroblasts
C3H 10T,/2 mouse
embryo fibroblasts
S. typhimurium
TA1537
E. coli
Chinese hamster
ovary cells
Chinese hamster
ovary cells
Human lymphocytes
K-l Chinese hamster
ovary 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 & Nelson,
1981
Garry & Nelson,
1981
Lawlor, 1995; 1996
Lawlor, 1995; 1996
Murli, 1996b, 1996d
Murli, 1996b, 1996d
Murli, 1996c;
NOTOX, 2000
Sadhu, 2002
Freire etal., 2008
NA= Not applicable
The data summarized above suggest that PFOA is not a mutagen. A single positive result
in S. typhimurium was not reproducible by the same authors and was not replicated in other
studies. Potential chromosomal effects were found in CHO cells at toxic concentrations but not
in human lymphocytes.
PFOA was tested twice in the in vivo mouse micronucleus assay. PFOA did not induce
any significant increases in micronuclei and was considered negative under the conditions of this
assay (Murli, 1995, 1996a).
Zhao et al. (2010) used AL cells to determine the mutagenicity of PFOA to mammalian cells. AL
cells are a human-hamster hybrid containing CHO-K1 chromosomes and a single copy of human
chromosome 11. The significance of human chromosome 11 is that is encodes for expression of
the human cell surface protein CD59. AL and mitochondria-deficient AL cells were incubated
with 0, 1, 10, 100, or 200 uM PFOA for up to 16 days and used in the 3-(4,5-dimethylthiazol-2-
yl)-2,5-diphenyltetrazolium bromide (MTT) viability, mutation, or caspace assays. Reactive
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oxygen species, nitric oxide, and superoxide anion production were measured in the cells, and
the effects of reactive oxygen species/reactive nitrogen species quenchers [0.5% dimethyl
sulfoxide (DMSO) and 0.2 mM NG-methyl-L-arginine, respectively] on mutagenicity and
caspace acitivities were determined. At 100 and 200 uM 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 uM PFOA. There was no increase
in mutations in mitochondria-deficient AL cells after incubation with 100 or 200 uM PFOA.
Production of reactive oxygen species, nitric oxide, and superoxide anion was
significantly increased at 100 and 200 uM PFOA after incubation for 1, 4, and 16 days.
Inhibition of reactive oxygen species through incubation with DMSO significantly decreased the
CD59 mutation frequency caused by 200 uM PFOA after a 16 day incubation. Incubation with
100 or 200 uM PFOA did not increase reactive oxygen species or superoxide anion production in
mitochondria-deficient AL cells. The highest dose significantly increased caspase 3/7 and 9
activities after 1 and 4 day incubations. Incubation with 0.5% DMSO and 0.2 mM NG-methyl-L-
arginine significantly decreased the increased caspace activity induced by 200 uM PFOA. The
results led the authors to suggest that mitochondrial-dependent reactive oxygen species may play
an important role in PFOA-induced mutagenicity and induction of caspase activities may be
mediated by reactive oxygen and nitrogen species.
4.3.2 Immunotoxicity
Yang et al. (2000, 2001, 2002a,b) completed a series of studies investigating the
immunotoxic effects of PFOA. In the first study, Yang et al. (2000) examined the liver, spleen,
and thymus effects of several known 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, or 10
days. At the end of the feeding period, mice were sacrificed and the liver, spleen, and thymus
were weighed. Administration of PFOA resulted in a significant increase in liver weight relative
to control even at day two. Following five days of administration, significant decreases in
thymus and spleen weight were noted.
A second component of the Yang et al. (2000) study examined the effect of 0.02% PFOA
in the diet on the cellularity, cell surface phenotype, and cell cycle of thymocytes and
splenocytes. After 7 days, significant decreases in the total number of thymocytes (|85%) and
splenocytes (|80%) were observed. There is a pattern to the development of thymocytes that
must be considered when evaluating the impact of chemicals on their differentiation. Early
thymocytes formed in the bone marrow do not express CD4 or CDS (CD4"CD8"). In the thymus
they differentiate and express both CD4 and CDS (CD4+CD8+). They also undergo proliferation
and down regulation 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 numbers 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+
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surface protein marking them as T-cells for exportation to the spleen. There are several classes
of the T-cells that are characterized by surface proteins. Yang et al. (2000) found significant
decreases in helper CD3+T-cells with CD4+ surface proteins (78%) and cytotoxic CD3+T-cells
with CD8+ surface proteins (74%). The authors suggested that, unlike the CD3+T-cells that
originate in the thymus, the decrease in CD19+B cells of the spleen reflects decreased
differentiation and maturation in the bone marrow where they are formed.
In the final phase of the Yang et al. (2000) study, the effects of in vitro exposure of
thymocytes and splenocytes to PFOA were examined. The in vitro studies showed spontaneous
apoptosis occurring in splenocytes and thymocytes after 8 or 24 hours of culturing in the
presence of varying concentrations (50, 100, or 200 uM) of PFOA. However, under the exposure
conditions PFOA did not appear to significantly alter the cell cycle.
In a later publication Yang et al. (2001) reported on their examination of the
immunosuppressive effects of PFOA. As was the case with their earlier publication (Yang et al.,
2000), the 2001 report includes several components. A diet of 0.02% PFOA (-40 mg/kg/day)
was fed to C57BL/6 mice for 2 to 10 days. One group of animals was exposed to PFOA each
day until day of sacrifice on days 2, 5, 7, or 10. At sacrifice body, liver, and spleen weights were
recorded. A second group of animals was dosed according to the same schedule, but dosing
ceased after day 7, and the animals were fed normal diets for 2 to 10 days in order to monitor
recovery from the effects of exposure. In the recovery group, animals were sacrificed after 2, 5,
or 10-day recovery periods.
The mice that received 0.02% PFOA for up to 10 days experienced significant increases
in liver weight compared to controls beginning at day 2. Significant decreases in thymus and
spleen weights were observed starting on day 5. Body weight increased for the first two days of
the study and decreased continuously for the remainder of the exposure period. The activity of
palmitoyl-CoA and lauroyl-CoA, biomarkers for PPARa activation and and peroxisome
proliferation, were also increased significantly and increasingly across the exposure period. The
impact of PFOA exposure was similar to that observed in the Yang et al. (2000) study. After
administration for 7 days, the number of thymocytes expressing neither CD4 nor 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
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the control values for the thymocyte (CD4+ and CD8+) or splenocyte (CD3+ and CD19+)
populations evaluated.
In the second component of the Yang et al. (2001) study, C57BL/6 mice were
administered diets consisting of 0.001%-0.05% PFOA (w/w) for 10 days. These doses are
equivalent to approximately 2.0-100 mg/kg/day. There was a dose-related decrease in spleen
and thymus weights and dose-related increase in liver weights accompanied by a corresponding
increase of pamitoyl Co-A and lauroyl Co-A activity. Enzyme activity was significantly
increased for all doses. Spleen and thymus weights were significantly decreased at doses >
0.01% and higher but not for the lower doses; the increases in liver weights were significantly
increased for the 0.02% and 0.05% doses.
A Yang et al. (2002a) study was designed to examine the possible involvement of 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.
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 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 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.
Yang and colleagues extended their studies of the immunotoxicity of PFOA in a feeding
study designed to examine the effects of PFOA on specific humoral immune responses in mice
(Yang et al. 2002b). For this study, 0.02 % PFOA was administered to male C57BL/6 mice for
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10 days. The the animals were then evaluated via plaque forming cell (PFC) and serum antibody
assays, for their ability to generate an immune response to horse red blood cells (HRBCs). Ex
vivo and in vitro splenic lymphocyte proliferation assays were also performed. The results
showed that mice fed normal chow responded to challenge with HRBCs with a strong humoral
response, as measured by the PFC assay. In contrast, mice fed PFOA responded to HRBC
immunization with no increase in HRBC-specific PFCs, relative to unimmunized controls.
However, in experiments where PFOA-treated mice received normal chow following HRBC
immunization, there was a significant recovery of the numbers of specific PFCs stimulated. The
suppression of the humoral immune response by PFOA was confirmed by analysis of the serum
anti-HRBC response.
In ex vivo experiments, splenocytes isolated from control mice responded to both ConA
and LPS with lymphocyte proliferation, as measured by thymidine incorporation. However,
treating mice with PFOA (0.02% for 7 days) attenuated the proliferation. In a set of in vitro
experiments, PFOA (1- 200 uM) added to the culture medium of splenocytes cultured from
untreated mice did not cause an alteration of lymphocyte proliferation in response to LPS or
ConA.
DeWitt et al. (2008) expanded the repertoire of studies of the immunological effects of
PFOA by examining various aspects of humoral and cellular immunity. The first component of
their publication had many similarities with the Yang et al. (2001) study. Adult female
C57BL/6J mice (n=40/endpoint and 8/group) were exposed to a single daily dose of 30 mg
PFOA/kg/day in distilled water by gavage for 10 continuous days. After 10 continuous days of
exposure, half of the mice continued receiving PFOA from Day 11 through Day 15 (constant
group) while the other half received distilled water from Day 11 through Day 15 (recovery
group). On Day 11, the mice were immunized with sheep red blood cells (SRBC). Sacrifices
took place on Day 16(1 day post-exposure period) or Day 31(15 days post-exposure period).
Vehicle and cage controls were also included in the study. All groups were monitored for the
following effects:
• Body weight and organ weights (Day 16, Day 31)
• Serum IgM levels (Day 16)
• Delayed-type hypersensitivity (DTK) foot-pad response to bovine serum album (BSA)
(Day 26)
• Serum IgG levels after booster immunization with SRBC on Day 20 (Day 31)
The results for body and organ weights were comparable to those in the Yang et al.
(2001) study. Body weight was significantly decreased from Days 8-11 for both PFOA-treated
groups and on Day 16 for mice in the constant exposure group. By Day 31, there were no body
weight differences between the groups. Relative liver weight was significantly elevated in both
PFOA-treated groups on Days 16 and 31. Absolute and relative spleen and thymus weights of
animals in both PFOA groups were significantly decreased compared to control groups on Day
16. By Day 31, thymus and spleen weights were not statistically different between control and
treated mice. IgM levels following immunization with SRBC were reduced by up to 20%
compared to controls on post exposure day 1 in both the recovery and constant exposure groups.
There were no significant differences from controls for IgG levels and DTH foot-pad responses
to the BSA challenge.
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The C57BL/6 mice used for the continuous-dosing vs. recovery component of the DeWitt
et al. (2008) study were found to develop ulcerative dermatitis following the PFOA exposure. It
was determined that this effect was a genetic susceptibility in the strain and they were not used
for the dose-response component of the study; the C57BL/6N strain was used in its place.
Two studies of dose-response were included in the DeWitt et al. (2008) publication.
Groups of 16 female C57BL/6N mice were given 0, 3.75, 7.5, 15, or 30 mg PFOA/kg/day in the
drinking water for 15 days during the first experiment. In the second experiment, the doses were
0, 0.94, 1.88, 3.75, or 7.5 mg PFOA/kg/day administered for 15 days in the drinking water. The
immunological sensitization and post-dose monitoring were identical to that used in the constant-
dosing vs. recovery experiment.
In the first experiment, body weight was significantly decreased from Day 8-16 at 30
mg/kg PFOA and on Day 16 at 15 mg/kg PFOA. As observed previously, liver weights were
significantly elevated at Day 16 and Day 31 at all doses. Absolute and relative spleen and
thymus weights were significantly decreased at >15 mg/kg PFOA on Day 16. With the
exception of the absolute thymus weight at 15 mg/kg PFOA, all spleen and thymus weights were
similar to weights in controls 15 days after dosing. The IgM response was significantly reduced
at >3.75 mg/kg PFOA in a direct dose-related manner. The IgG response was slightly but
significantly elevated at 3.75 and 7.5 mg/kg PFOA but similar to that of the control level at the
higher doses. Thus, there was a direct response of IgM but not IgG to dose across the dose
levels. There was no significant change in the DTH response at any dose. The LOAEL from the
first experiment was 3.75 mg/kg/day dose based on decreased IgM and increased IgG response
to SRBC immunization and increased liver weights (p<0.05).
The second dose-response experiment confirmed the 3.75 mg/kg/day dose as the
immunological LOAEL on the basis of significantly decreased spleen weight, decreased IgM
levels on Day 16, and increased IgG levels on Day 31. The immunological NOAEL was 1.88
mg/kg/day. Benchmark dose analysis of IgM serum titer data gave a lower bound 95%
confidence limit of 1.75 mg/kg/day on a benchmark dose (one standard deviation) of 3.06
mg/kg/day. Liver weight was significantly increased at all doses on Days 16 and 31. The
LOAEL for increased liver weight was 0.94 mg/kg PFOA.
Loveless et al. (2008) administered 0, 0.3, 1, 10, or 30 mg linear PFOA/kg by oral gavage
to groups of male CD rats (n=10/group) and CD-I mice (n=20/group) for 29 days. Both species
received a dose of SRBC on day 23 (rat) or 24 (mice). A separate group of high dose rats and
mice were injected with water instead of SRBC. Rat body weight was recorded on days 0, 3, and
6-28, while mice were weighed daily. At necropsy blood was collected for immune parameters.
Cell counts were determined for the thymus and spleen. Total spleen and thymocyte cell counts
and organ weights in exposed rats were comparable to control. Microscopic examination of the
thymus, mesenteric lymph nodes, and popliteal lymph nodes revealed no effects in treated rats
resulting from PFOA exposure. There was no difference observed in IgM liters between treated
and control rats. The immunological NOAEL was 30 mg/kg/day.
In mice, absolute and relative spleen and thymus weights were significantly decreased at
>10 mg/kg/day. The relative spleen weight of mice dosed with 1 mg/kg/day was also
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significantly decreased compared to control animals. Spleen and thymus cell counts were
significantly decreased and minimal to severe lymphoid depletion/atrophy of the thymus was
observed at >10 mg/kg/day. IgM liters were significantly decreased at >10 mg/kg/day. Serum
corticosterone was significantly increased at 10 mg/kg/day and elevated (not statistically
significant) at 30 mg/kg/day. When IgM and corticosterone were plotted against each other, a
negative correlation coefficient suggested that increasing corticosterone levels decreased the
ability to make SRBC antibody. The LOAEL was 1 mg/kg/day based on decreased spleen
weight, and the NOAEL was 0.3 mg/kg/day.
Loveless et al. (2008) hypothesized that at least a portion of the thymic response to PFOA
might be related to physiological stress and increased levels of corticosterone hormones. DeWitt
et al. (2009) investigated this hypothesis by comparing the immunological response of
adrenalectomized (ADX) C57BL/6N female mice to that of sham operated female mice from the
same strain. The animals were dosed with 0, 3.75, 7.5, or 15 mg PFOA/kg/day in the drinking
water for 10 days. Body weight was recorded on dosing days 0, 4, and 8, plus 2 and 5 days post
dosing. On exposure days 5 and 10, blood and serum were collected for analysis of a broad array
of clinical chemistry parameters including activity of liver enzymes indicative of cellular damage
(AP, AST, ALT, LDH, GOT, and SDH), serum lipids (cholesterol and triglycerides) and
corticosterone. A baseline measure of corticosterone was determined from serum samples
collected before dosing began. One day after cessation of exposure, the mice were immunized
with SRBC. Four days later, serum was collected and the levels of corticosterone and IgM were
determined.
Body weight in the sham-operated mice declined during dosing in the highest dose group
but recovered by 5 days post dosing. In the ADX mice, body weight declined during dosing at
7.5 and 15 mg PFOA/kg/day, but recovered in mice receiving 7.5 mg PFOA/kg/day by 5 days
post dosing. There were significant increases in ALT, AST, LDH, and SDH at the highest dose
for the ADX mice indicative of damage to hepatic cell membranes (Table 4-17).
TABLE 4-17. Selected Clinical Chemistry Parameters in Mice Treated with PFOA for 5 Days
Dose (mg/kg/day)
ALT
AST
LDH
SDH
Sham-Operated
0
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*
From DeWitt etal., 2009
*=p<0.05 vs corresponding sham control or ADX control group
Serum levels of triglycerides significantly decreased compared to controls with all doses
for the sham operated mice on day five of dosing but only for the 7.5 and 15 mg/kg/day doses in
the ADX mice. Cholesterol levels were significantly decreased (p<0.05) in the sham operated
mice at the highest dose, but no differences in cholesterol levels were observed in the ADX mice.
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After 10 days, there were no significant differences in liver enzymes for the ADX or
sham mice. However, there was a dose-related trend towards increased levels of liver enzymes
for the PFOA-exposed sham operated animals and for LDH in the PFOA-exposed ADX animals
(Table 4-18).
TABLE 4-18. Selected Clinical Chemistry Parameters in Mice Treated with PFOA for 10 Days
Dose (mg/kg/day)
ALT
AST
LDH
SDH
Sham-Operated
0
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
From DeWittetal, 2009
At the end of dosing, corticosteroid levels in the sham-operated animals were greatly
elevated compared to the levels in the control animals at all doses and the difference was
statistically significant at the highest dose. By 5-days post dosing the corticosterone levels had
declined for all doses but were still elevated compared to controls for the 7.5 and 15 mg/kg/day
groups. In the animals lacking their adrenal glands, there were no statistically significant
differences in the hormone levels. IgM levels were significantly lower than controls at the
highest dose for the sham-operated animals and at the two highest dose groups for the ADX
mice. However, when comparing the sham mice to the ADX mice the only significant difference
in IgM was found for the 7.5 mg/kg/day animals. On the basis of data it appears that stress-
related corticosterone production did not have a major impact on the IgM response to the SRBC
inoculation.
Son et al. (2009) administered 0, 2, 10, 50, or 250 ppm PFOA (0, 0.49, 2.64, 17.63, or
47.21 mg/kg PFOA) in the drinking water to 4 week old male ICR mice for 21 days to determine
if PFOA alters T lymphocyte phenotypes and cytokine expression in mice. The spleen, thymus,
and trunk blood were collected at sacrifice. Sections of the spleen and thymus were processed
for histological examination. Splenic and thymic expression of mRNA from proinflammatory
cytokines including tumor necrosis factor-a, interleukin-lp, and interleukin-6, and the proto-
oncogene c-myc were analyzed using RT-PCR. Flow cytometry was used to phenotype the
splenic and thymic lymphocyte populations.
Spleen and thymus weights were slightly decreased in mice treated with PFOA.
Enlargement with marked hyperplasia of the white pulp and increased cellular density of the
lymphoid follicles were observed in spleens at 250 ppm. In the thymus, decreased cortex and
medulla thickness and densely arranged cortex lymphoid cells were observed at 250 ppm.
Tumor necrosis factor-a, interleukin-lp, interleukin-6, and c-myc expression were significantly
elevated at 250 ppm in the spleen. Interleukin-lp expression was also elevated at 50 ppm in the
spleen. In the thymus, c-myc expression was significantly elevated by treatment with 50 and 25
ppm PFOA.
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The splenic and thymic lymphocyte population was altered by PFOA treatment as shown
in Table 4-19. A 50% decrease was observed in splenic CD8+ lymphocytes at all PFOA doses,
and an increase in splenic CD4+ lymphocytes of 43% and 106% at 50 and 250 ppm PFOA,
respectively, were observed. In the thymus, a 110% increase was observed in thymic CD8+
lymphocytes at 250 ppm, but thymic CD4+ lymphocyte populations were not affected by PFOA.
TABLE 4-19. Impact of PFOA on Splenic and Thymic Lymphocyte Populations
Dose (mg/kg/day)
Spleen
CD4'CD8-
CD4+CD8'
CD4"CD8+
CD4+CD8+
Thymus
CD4'CD8-
CD4+CD8'
CD4'CD8+
CD4+CD8+
0.49
t
-
1
1
-
-
-
-
2.64
-
-
1
1
-
-
-
-
17.63
-
t
1
-
-
-
-
1
47.21
1
t
1
-
t
-
t
1
From Son et al., 2009
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% or 0.02% PFOA (~2 or 40 mg/kg) in the
diet (w/w) for 10 days. After the last dose all mice were sacrificed. Sacrifice was delayed for a
subset of the animals until 2 hours after they had received a lipopolysaccharide (LPS) injection
to stimulate an immunological response. Blood, peritoneal exudate cells, liver, epididymal fat,
spleen, thymus, and bone marrow were recovered. The blood, peritoneal exudate, bone marrow,
and spleen were evaluated for total and differential white blood cell counts and concentrations of
tumor necrosis factor (TNF-a) and interleukin 6 (TL-6).
Consistent with other studies of the 0.02% dose, there was a significant increase in liver
weight after the 10-day exposure. Body weight, thymus weight, spleen weight, and epididymal
fat depots were decreased. Food consumption in these animals was reduced by 35% which may
have played a role in the reduced body, organ, and tissue weights. Compared to the controls
there was a significant decrease in total white cells, lymphocytes, and neutrophils at 0.02%
PFOA. This same dose was associated with a decrease in total white cell count in bone marrow
and spleen, and an increase in the proportion found as macrophages in the bone marrow, spleen
and peritoneal cavity. Although the total number of macrophages was not reduced in the
peritoneal cavity and spleen, it was reduced in the bone marrow. The increase in the proportion
of macrophages reflects a decrease in other white cell populations. There was significant increase
in the concentration of IL-6 in all of the 0.02 % dosed animals but only those receiving the LPS
injections showed a significant increase in TNF-a. The 0.001% dose (about 2 mg/kg/day) was a
NOAEL.
In a pilot study, Brieger et al. (2011) examined the effects of PFOA on human
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leukocytes. Peripheral blood mononuclear cells (PBMC) were obtained from the blood of 11
voluntary donors (n=6 females, 5 males). PBMC were incubated with varying concentrations of
PFOA followed by assays for cell viability, proliferation, and natural killer (NK) cell activity.
The human promyelocytic leukemia cell line, HL-60, was used in cell viability and monocyte
differentiation assays. The various components of the assays employed are identified as follows:
• In the cell viability assay, the PBMC were incubated with 0-500 ug/mL for 24, 48, or
72 hours, and HL-60 cells were incubated with 0-125 ug/mL PFOA for 24 hours.
• In the proliferation assay, the PBMC were incubated with 0-100 ug/mL PFOA for 24
hours, labeled with 6-carboxyfluorescein succinimidyl ester (CFSE), stimulated with
concanavalinA, a T-cell mitogen (ConA, 5 ug/mL to half of all samples), and
incubated for an additional 72 hours.
• For the NK assays, PBMC cells 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) tabled
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 nM 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. Lipopolysaccharide
(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
9 samples used, cells from 6 donors had slightly increased proliferation and 3 had no response.
In cells incubated with ConA and 100 ug/mL PFOA, a non-significant decrease in the number of
proliferating cells was observed. PFOA decreased NK cell activity approximately 16% (not
statistically significant). In the presence of 1,25D3, 100 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 priod to the end of incubation. The authors also looked at the
correlation between basal PFOA concentration and cytokine release. A significant association
was observed between PFOA concentration and the release of LPS-induced TNF-a and IL-6 by
peripheral monocytes. The authors suggested that the trends observed at the lower
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concentrations may show an impact on human immunity with a larger population.
Inhalation Exposure
No data on the effects of inhalation exposure on immunological endpoints were identified
in the literature.
Dermal Exposure
Fairley et al. (2007) carried out a complex study of toxicity and respiratory
hypersensitivity to ovalbumin (OVA) as impacted by dermal exposure to PFOA dissolved in
acetone compared to acetone alone. There were several phases to the study. In the first phase, a
range finding study, PFOA was applied to each ear of female BALB/c mice (n=5/group) at doses
of 0, 0.01%, 0.1%, 0.25%, 0.5%, 1.0%, or 1.5% PFOA (equivalent to 0, 0.25, 2.5, 6.25, 12.5, 25
or 50 mg/kg/day) for 4 days. Six days after last inoculation, the animals were sacrificed. The
liver, spleen, and thymus were recovered and weighed. A significant increase in liver weight
was observed at >6.25 mg/kd PFOA. Spleen weight was significantly decreased in mice dosed
with 25 mg/kg and 50 mg/kg PFOA, and thymus weight was significantly decreased in mice at
the highest dose (p<0.05). The cell counts in the spleen were significantly decreased compared
to control at all doses and for the highest two doses in the thymus. The LOAEL was 6.25
mg/kg/day based on a statistically-significant increase in liver weight (p<0.01) and the NOAEL
was 2.5 mg/kg/day.
In the second phase of the Fairley et al. (2007) study, groups of 5-15 animals were dosed
dermally on the ears for 4 days with doses of 0, 0.5%, 0.75%, 1.0%, or 1.5% PFOA (equivalent
to 0, 12.5, 18.75, 25, or 50 mg/kg/day). On days 1 and 10 they were injected ip with either 2.0
mg alum or 7.5 ug ovalbumin (OVA) and 2.0 mg alum in 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 (ELISA). IgE is the immunoglobulin that is best correlated
with respiratory allergic responses. It functions to stimulate mast cells and basophils to release
histamine and other mediators of inflammation (Saladin, 2004). The IgE response was increased
in a dose-related fashion compared to the OVA control for all the PFOA treated animals; the
increase was significant (p<0.05 or 0.01) at doses >18.75 mg/kg/day. The OVA-specific IgE
response did not demonstrate a direct response to dose but there was a significant increase
(p<0.05) for the 18.75 and 25 mg/kg/day groups. The OVA-specific response for the 37.5
mg/kg/day highest dose groups was practically indistinguishable from the OVA control.
The dermal LOAEL was 12.5 mg/kg/day based on increased liver weight and decreased
spleen weight and spleen cellularity. No NOAEL was established.
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In part three of the Fairley et al., (2007) study, mice (n=5) were dosed dermally via their
ears for 4 days as described above (0, 12.5, 18.75, 25, and 50 mg/kg/day). On days 19 and 26
after the start of dosing they were challenged by pharyngeal aspiration of 250 |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 spleenocytes carrying the B220+ marker (expressed on B
cells, activated B cells, subsets of T and NK cells) compared to the OVA controls. The changes
were significantly different for the 25 mg/kg/day (p<0.05) and 37.5 mg/kg/day (p<0.01) groups.
After the day 19 challenge the mice (n=5) were placed in a plethysmography chamber for
measurement of enhanced pause airway respiration (PenH) values. PenH values reflect volume
of air in the lungs. Once in the chamber they were challenged with nebuilized methacholine for
three minutes followed by two minutes of fresh air. The PenH measures were recorded every 30
seconds over the next 5 hours. The area under the curve (AUC) for the PenH measures was
determined after correction for baseline (acetone control, no OVA or PFOA). An AUC of 1.6
was considered to be a positive response. Twenty-four hours later blood was drawn from the
abdominal artery and the mice were sacrificed. The lungs were recovered for histological
analysis. An increase in antigen-specific hyperactivity response to PFOA, in both the PenH
values and the number of animals responding was observed at doses up to about 25 mg/kg/day.
The PenH AUC was significantly (p <0.05) increased in mice treated with 25 mg/kg/day PFOA
and OVA compared to the OVA control mice, but there was no significant difference between
the OVA control and the animals exposed to 50 mg/kg/day PFOA and OVA. The LOAEL for
the PenH response was 25 mg/kg/day, and the NOAEL was 18.75 mg/kg/day.
Histopathological evaluation of the lungs revealed macrophage and eosinophil infiltration
in response to the challenge with 250 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 were also observed. Increased secretory matter, sloughing of
epithelium, and secretory cell necrosis were observed in mice exposed to all concentrations of
PFOA and OVA. The response was not observed in the mice exposed to only PFOA. The
authors concluded that dermal exposure to PFOA was immunotoxic and enhanced the airway
hypersensitivity response to OVA suggesting that PFOA may augment the IgE response to
environmental allergens.
4.3.3 Hormone Disruption
Thyroid. Martin et al. (2007) administered 20 mg PFOA/kg to adult male Sprague-Dawley rats
(n = 4 or 5) for 1, 3, or 5 days by oral gavage and determined the impact of PFOA on hormone
levels. Blood was collected via cardiac puncture and the serum was analyzed for cholesterol,
testosterone, free and total T4, and total T3. RNA extracted from the livers was used for gene
expression profiling, genomic signatures, and pathway analyses to determine a mechanism of
toxicity.
Following a 1 day, 3 day, and 5 day dose, a significant decrease (p<0.05) was observed in
serum cholesterol (~|45-72%), total T4 (~|83%), free T4 (~|80%), and total T3 (~|25-48%).
Serum testosterone was significantly decreased (p<0.05, ~|70%) following a 3 day and 5 day
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PFOA dose. PFOA treatment was matched to hepatotoxicity-related genomic signatures, as well
as signatures for hepatocellular hypertrophy, hypocholesterolemia, hypolipidemia, and
peroxisome proliferation. 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 and Diol, which deiodinates prohormone T4 to bioactivate T3,
were affected by PFOA. Treatment with PFOA resulted in significantly upregulated expression
ofDio3 and downregulated expression of Diol (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.
Reproductive Hormones. Cook et al. (1992) gavage dosed male CD rats (n=15/group) for 14
days with 0, 1, 10, 25, or 50 mg PFOA/kg to examine the possibility than an endocrine-related
mechanism may explain Leydig cell adenomas observed in rats. A separate control group was
pair fed to the 50 mg/kg group. Blood and testicular interstitial fluid were collected at necropsy
for hormone analysis including testosterone, estradiol, and LH. A separate group of rats was
dosed with 0 or 50 mg PFOA/kg for 14 days and challenged with 100 IU of human corionic
gonadotropin (hCG) or 2 mg naloxone/kg 1 hour prior to necropsy to induce testosterone
concentrations. Blood was collected and analyzed for testosterone and LH. Serum from rats
challenged with 100 IU hCG was also anlayzed for progesterone, 17 a-hydroxyprogesterone, and
androstenedione.
The relative liver weight at 10 mg PFOA/kg was significantly increased (p<0.05). The
relative liver weight was significantly increased (p<0.05) and the accessory sex organ unit
relative weight was significantly decreased (p<0.05) at 25 and 50 mg PFOA/kg compared to
those weights in control rats. The results at the highest dose were also different when compared
to the pair fed controls. The relative weight of the ventral prostate was significantly decreased at
the highest dose compared to the pair fed control. Serum estradiol was significantly increased at
>10 mg PFOA/kg compared to the control. No differences were observed in testosterone and LH
between the treated rats and control. In the challenge experiment, serum testosterone was
significantly decreased (p<0.05) by treatment with 50 mg PFOA/kg after challenge with 100 IU
hCG. No differences in testosterone concentration were observed in the naloxone challenged
rats, and no differences in LH were observed after either challenge. In the hCG challenged rats,
androstenedione was significantly reduced at 50 mg PFOA/kg, but no differences in
concentrations were observed in progesterone or 17 a-hydroxyprogesterone between control and
treated rats. The authors suggested that the decreased serum testosterone levels observed may be
due to decreased conversion of 17 a-hydroxyprogesterone to androstenedione attributable to the
increased serum estradiol levels. The LOAEL was 10 mg/kg based on increased liver weight and
increased serum estradiol levels, and the NOAEL was 1 mg/kg.
Biegel et al. (1995) conducted in vitro, in vivo, and ex vivo studies to determine the
effects of PFOA on Leydig cell function. In the in vitro study, Leydig cells were cultured ± 2 IU
hCG (final 3 hours) and 0, 100, 200, 250, 500, 700, or 1000 |iM PFOA for a total of 5 hours and
analyzed for testosterone concentration. Leydig cells were also incubated ±500 jiM PFOA and
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anlayzed for testosterone and estradiol at 0, 0.5, 1, 3, 6, 12, 24, and 48 hours. Male CD rats were
gavage dosed for 14 days with 0, 0 pair fed, or 25 mg PFOA/kg and necropsied on day 15.
Blood and testicular interstitial fluid were collected for hormone analysis. Liver samples were
collected for analysis of peroxisomal p-oxidation and microsomal aromatase activities. 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.
In the in vitro studies, there was no effect of PFOA treatment on testosterone in Ley dig
cells cultured without hCG. In cells cultured with hCG, PFOA caused a dose dependent
decrease in testosterone production. At 100 jiM PFOA + hCG, the testosterone concentration
was significantly increased compared to cells treated with only 100 jiM PFOA. Cytotoxicity
occurred at > 750 jiM PFOA. In the time course experiment, 500 jiM PFOA significantly
inhibited hCG-stimulated release of testosterone at time points of at least 3 hours compared to
control. Estradiol levels of PFOA-treated Leydig cells at 48 hours were statistically greater than
the control.
In the in vivo study, serum estradiol was significantly increased (p<0.05) by 25 mg
PFOA/kg when compared to the ad libitum and pair fed control rats. Testicular interstitial fluid
testosterone concentration was significantly decreased (p<0.05) and microsomal aromatase
activity, and peroxisomal P-oxidation activity were significantly increased (p<0.05) in PFOA-
treated rats compared to the pair fed control rats. In the ex vivo study, an increase of 8.6-fold in
testosterone production (p<0.05) was observed in Leydig cells isolated from PFOA-treated rats.
The authors suggested that the increased serum estradiol levels resulted from liver aromatase
induction by PFOA, and that PFOA may directly affect Leydig cell function.
Liu et al. (1996) treated adult male Crl:CD BR (CD) rats (n=15/group) with 0, 0 pair fed,
0.2, 2, 20, or 40 mg PFOA/kg for 14 days by oral gavage to determine the impact of PFOA on
aromatase activity. A separate study examined the effect of PFOA on aromatase activity in
cultured hepatocytes and is discussed below. Aromatase is a cytochrome P450 enzyme localized
to the endoplasmic reticulum that catalyzes the conversion of androgens to estrogens. At
necropsy on day 15, blood was collected for serum estradiol determination. Liver samples were
collected for determination of microsomal aromatase activity and total P450 concentration. The
testes were collected and testicular aromatase was determined.
In the in vitro study, hepatocytes isolated from control male CD rats were incubated with
0-1000 jiM PFOA and the aromatase activity was evaluated after 18, 42, and 66 hours (Liu et al.
1996). Compared to aromatase activity in time-matched control cultures, PFOA caused a
decrease in aromatase activity after 18 and 42 hours incubation with hepatocytes and an increase
after the 66 hour incubation period.
In the in vivo study, the body weight of rats treated with > 20 mg PFOA/kg was
significantly decreased (p<0.5) compared to the control rats. Pair fed control rats also had
significantly decreased body weight compared to the control rats. Body weight was not different
between the pair-fed control rats and rat dosed with 40 mg/kg PFOA. Absolute and relative
liver weight were significantly increased (p<0.5) at > 2 mg PFOA/kg. Relative testes weight was
significantly increased at >20 mg PFOA/kg, but the differences were due to decreased body
weight. There were no differences observed in testicular aromatase activity. In the remaning
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analysis, the pair-fed control group were similar to the ab libitum control group. The protein
yield of hepatic microsomes was significantly increased at >0.2 mg PFOA/kg and hepatic
aromatase activity, total hepatic aromatase activity adjusted for liver and body weight effects,
and serum estradiol were significantly increased (p<0.05) at >2 mg PFOA/kg. The maximum
increase in total hepatic aromatase activity was 16-fold and the increase was 2-fold for serum
estradiol. A significant correlation (p<0.0001) was observed between total hepatic aromatase
activity and serum estradiol. The aromatase activity in liver microsomes isolated from control
rats and incubated for 2 hours with PFOA was significantly decreased at > 100 jiM. The authors
estimated the ECso values for the outcomes and they are shown in Table 4-20. Liu et al. (1996)
concluded that the PFOA-increased protein yields suggested induction of the endoplasmic
reticulum resulting in aromatase induction which led to increased serum estradiol. However,
PFOA also inhibited aromatase activity which would explain why serum estradiol was only
increased up to 2-fold.
TABLE 4-20. Estimated EC50 Values
Parameters
Hepatic microsome protein yield
Hepatic microsomal aromatase activity
Absolute liver weight
Relatvie liver weight
Serum estradiol
Terminal body weight
EC50 (mg PFOA/kg)
0.53
0.76
1.07
1.56
3.24
11.65
From Liu et al. 1996
EC50= half-maximum response
In a study examining the impact of PFOA on aromatase activity, Liu et al. (1996) also
examined the impact of PFOA on peroxisome p-oxidation and cytochrome P450 activities. Male
Crl:CD BR (CD) rats (n=15/group) were orally dosed with 0, 0 pair fed, 0.2, 2, 20, or 40 mg
PFOA/kg for 14 days. Liver samples were collected for determination of microsomal total
cytochrome P450 concentration and peroxisome P-oxidation activity. Total cytochrome P450
was significantly increased (p<0.05) at >20 mg PFOA/kg and P-oxidation activity were 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 may
play in animals exposed during gestation compared to during their adult years. Timed-pregnant
CD-I mice were gavage dosed in two blocks on gestation days 1-17 but not thereafter. Block 1
animals were dosed with 0, 1, 3, or 5 mg PFOA/kg, and Block 2 animals were dosed with 0,
0.01, 0.1, 0.3, 1, or 5 mg PFOA/kg/day. At birth, pups were pooled within each block and dose
group and randomly redistributed among the dams (10 pups/litter). Offspring were weaned at 3
weeks, and a subset of females from each dose group (0, 0.01, 0.1, 0.3, 1, and 5 mg
PFOA/kg/day) was ovariectomized (OVX) at weaning or the day after weaning. All animals
were observed until they reached 18 months of age.
Body weight was recorded weekly for the first 9 months of age, followed by monthly
body weight recordings over the next 9 months. As the animals matured, they were evaluated for
the endpoints listed in Table 4-21. A group of naive 8-week old adult mice were dosed for 17
days with 0, 1, or 5 mg PFOA/kg/day to compare the impact of exposure in adult animals to
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those occurring during gestation. At 18 months of age, the mice were sacrificed. Blood,
retroperitoneal abdominal fat, interscapular brown fat, organs, and abnormal growths were
collected at necropsy.
TABLE 4-21. Data Collection for Female Mice Gestationally-exposed to PFOA
Test
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
From Hinesetal., 2009
Body weight of offspring born to dams exposed to 5 mg PFOA/kg was significantly
decreased (p<0.05) on PND1 and through 18 months of age compared to control offspring body
weight. At weaning, the body weight of offspring born to dams exposed to 1 mg PFOA/kg/day
was significantly decreased (p<0.05) compared to control offspring body weight. A significant
increase (p<0.05) in body weight, due to more rapid weight gain after week 10, compared to
intact control body weight was observed in intact mice exposed to 0.01-0.3 mg PFOA/kg/day
during gestation. Body weight of intact mice gestationally exposed to 0.01-0.3 mg PFOA/kg/day
was comparable to body weight of control mice at 18 months.
Due to the increased weight gain observed in intact mice exposed to PFOA during
gestation, glucose tolerance tests were carried out along with determination of serum insulin
concentration. In cases of insulin resistance, plasma glucose and insulin levels are elevated and
the insulin response is lessened. Insulin resistance has also been associated with excess
abdominal fat. Serum leptin levels were also determined as increased leptin levels have been
associated with a leptin-resistance mechanism of action for increased weight gain in humans.
Body mass composition was used to determine if there were differences in body fat between the
intact groups, and feed consumption was recorded to determine if consumption played a role in
body weight differences in intact control and intact gestionally-exposed mice. Serum estradiol
was measured to determine if PFOA impacted hormone levels at 18 months in intact control and
intact gestionally-exposed mice.
Glucose tolerance testing showed no statistically significant differences in baseline
glucose or response to glucose challenge at 15-16 weeks or at 17 months. At 21 and 31 weeks of
age, a significant increase in serum leptin and insulin levels was observed in intact mice exposed
to 0.01 and 0.1 mg PFOA/kg/day. No statistically significant difference was observed between
the fat to lean ratio of intact control and intact gestationally-exposed animals at 42 weeks of age.
No significant difference was observed in food consumption between intact control and intact
gestationally-exposed animals at 42 weeks of age. Serum estradiol levels were not different
consumption between intact control and intact gestionally-exposed animals at 18 months.
Exposure to PFOA as an adult did not result in body weight differences among the
groups at 18 months of age. The body weight of intact mice gestationally exposed to 1 mg
PFOA/kd/day was significantly increased (p<0.05) compared to adult mice exposed to 1 mg
PFOA/kg/day. No other differences in body weight among the groups were observed.
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No significant differences among the groups were observed in survival during the 18
month study. At necropsy, abdominal white fat was significantly decreased (p<0.05) at 1 and 5
mg PFOA/kg/day in gestationally-exposed intact mice compared to intact control mice.
Interscapular brown fat was significantly increased (p<0.05) at 1 and 3 mg PFOA/kg/day in
gestationally-exposed intact mice and in gestationally-exposed ovx mice at 1 mg PFOA/kg/day.
Relative spleen weight was significantly decreased (p<0.05) at 3 mg PFOA/kg/day in
gestationally-exposed intact mice and at 1 and 5 mg PFOA/mg (p=0.05-0.07) in 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, and at high doses, decreased
weight in early and late life, decreased white fat, increased brown fat, and decreased spleen
weight were observed. Under the conditions of the study, the developmental LOAEL was 0.01
mg PFOA/kg based on increased weight gain and increased serum insulin and leptin levels. No
developmental NOAEL was established. The adult NOAEL was 5 mg PFOA/kg, and no
LOAEL was established.
Adrenal Hormones. Thottassery et al. (1992) exposed intact or adrenalectomized (ADX) male
Sprague-Dawley rats to a single dose of 150 mg/kg PFOA in corn oil to determine the role of
adrenal hormones on liver enlargement and peroxisomal proliferation. ADX rats were dosed 2
days after surgery with PFOA (ADX PFOA), corticosterone (ADX CORT), or both (ADX
CORT PFOA). A group of intact and ADX rats received only the vehicle and served as controls.
The animals were sacrificed 48 hours after dosing with PFOA or vehicle. Assays were
conducted to determine DNA levels and changes in peroxisomal 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, vs. 30.5
pmole CO2/hr/mg protein, p <0.001). Ornithine decarboxylase activity was not different
between intact rats treated with PFOA and intact rats treated with the vehicle.
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PFOA increased whole liver peroxisomal 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 activity was not increased in rats
of the ADX PFOA group. Based on the results, the authors concluded that adrenal hormones
were not required to induce peroxisomal 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.
4.3.4 Physiological or Mechanistic Studies
Gene Expression. Rosen et al. (2007) examined the gene expression profile in the lung and liver
of mouse fetuses exposed to PFOA. Pregnant CD-I mice were gavage dosed with 0, 1, 3, 5, or
10 mg PFOA/kg/day on GD1-17. Dams were sacrificed on GDIS, and three fetuses per litter
were processed for total RNA from portions of the liver and lung. Global gene expression was
analyzed using Affymetrix gene chips.
A dose-related increase was observed in the number of genes altered by PFOA exposure
in both the liver and lung. A greater number of genes in the liver were altered compared the
number of genes altered in the lung. Analysis of the genes by canonical pathway or biological
function showed that most of the altered genes in both the liver and lung were associated with
lipid homeostasis. In the fetal lung, the two highest doses of PFOA altered genes associated with
fatty acid catabolism. In the fetal liver, all doses of PFOA were associated with genes involved
in fatty acid catabolism, lipid transport, cholesterol biosynthesis, bile acid biosynthesis,
lipoprotein metabolism, steroid metabolism, retinol metabolism, inflammation, phospholipid
metabolism, glucose metabolism, proteosome activation, and ketogenesis. Although PPARa is
known to regulate (at least partly) the expression of genes for the pathways or biological
functions involved in lipid homeostasis, PFOA may also activate other nuclear receptors
influencing the metabolic responses observed.
Rosen et al. (2008a) described the gene profiles in liver tissue from wild-type
12981/SvlmJ mice (7-8/group) and PPARa-null mice (129S4/SvJae-PPARatmlGonz/J, 6-8/group)
dosed with 0, 1, or 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. Real-time (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 was less at 1 mg/kg than at 3 mg/kg PFOA. However, 85% of the altered genes at 1
mg/kg PFOA were alterd at 3 mg/kg PFOA.
PPARa target genes including 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
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mice exposed to Wyeth 14,643. Principle components analysis showed that PFOA-treated
PPARa-null mice shared similarity with PFOA-treated wild-type mice at 3 mg PFOA/kg.
In wild-type PFOA and Wyeth 14,643 treated mice, alterations were observed in genes
associated with fatty acid metabolism (mostly upregulated), inflammatory response (mostly
downregulated), cell cycle control (mostly upregulated), peroxisome biogenesis (mostly
upregulated), and proteasome structure and organization (mostly upregulated). In genes
associated with xenobiotic metabolism, the response was different between PFOA and Wyeth
14,643 wild-type mice. Many of the Cyp2 genes were upregulated by PFOA and downregulated
by Wyeth 14,643. In PPARa-null PFOA-treated mice, genes associated with fatty acid
metabolism, inflammation, xenobiotic metabolism, and cell cycle control were altered similar to
the changes observed in PFOA-treated wild-type mice.
RT-PCR generally revealed good agreement with microarray analysis. However,
expression of Ehhadh, a PPARa regulated gene, was up-regulated in PFOA-treated wild-type
mice but not in PFOA-treated PPARa-null mice in microarray analysis. Expression of Ehhadh
was up-regulated in all PFOA-treated mice in RT-PCR analysis. The authors concluded that
PFOA induces transcriptional changes mediated through 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 129Sl/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 (WY)-14,643
(PPARa agonist); 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 WY-14,643 for 12 hours, or 3 or 7 days were compared. Four groups of altered genes
were identified based on their expression behavior in wild-type and PPARa-null PFOA-exposed
mice compared to genes from WY-14,643 treated mice. The first group consisted of genes (397)
regulated by both PFOA and WY-14,643 in wild-type mice. They had a common direction and
magnitude of change and were characterized as being lipid homeostasis, inflammation, cell
proliferation, or proteome maintenance genes. Group II consisted of genes in wild-type mice
(51) regulated solely by PFOA, and most were involved in amino acid metabolism. Of the 81
genes altered by PFOA exposure in PPARa-null mice (Group III), 62 had similar expression in
wild-type mice, and many were involved in lipid metabolism. Regulation of these genes was
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also observed in WY-14,643 wild type mice. Group IV genes (19) were altered by PFOA only
in PPARa-null mice, and most were xenobiotic metabolizing enzymes.
By comparing the gene expression patterns between PFOA and WY-14,643, the authors
concluded that:
• PPARa is required for a majority of the transcriptional changes observed in the
mouse liver following PFOA or WY-14,643 exposure
• PFOA regulates genes that are also regulated by other peroxisome proliferators in a
PPARa-dependent manner, and
• PFOA also regulates PPARa-independent genes.
The PPARa-independent genes included those involved in lipid homeostasis (upregulated),
amino acid metabolism (down regulated), and xenobiotic metabolism (upregulated).
The transcription profiles of PFOA exposed wild-type and 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
suggested that some genes altered by PFOA exposure in PPARa-null mice are regulated by CAR
butnotNrfZ.
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, or 200 uM PFOA for 24
hours. Expression of acyl-CoA oxidase (Acox), Cyp4al (rat), Cyp4all (human), acyl-CoA
thioesterase (Cte-rat, Acotl-human), and DNA damage inducible transcript (Ddit3) were
determined by quantitative RT-PCR. The aforementioned genes are inducible by peroxisome
proliferators except Ddit3 which is induced in the presence of direct or indirect DNA damage.
Exposure to > 100 uM 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 uM
PFOA, and Cte/Acotl was significantly induced at >20 uM PFOA in rat hepatocytes only.
Expression of Cyp4al/l 1 was significantly induced in rat hepatocytes at 5-50 uM and in human
hepatocytes at 20-50 uM. 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 4.2.2. Male, 8-week old wild-type
(mPPARa) mice, PPARa-null mice, and hPPARa mice were gavage dosed with 0, 0.1, or 0.3
mg/kg PFOA (n=4-6/group) for 2 weeks and sacrificed 18-20 hours following the last dose.
Livers were collected and analyzed for mRNA (RT-QPCR) and protein levels (Western blot
analysis) of PPARa and related genes (retinoid X receptor alpha [RXRa], peroxisomal
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bifunctional protein [PH], peroxisomal thiolase [PT], very long chain acyl-CoA dehydrogenase
[VLCAD], medium chain acyl-CoA dehydrogenase [MCAD], and CYP4A10). RXRa forms a
heterodimer with PPARa controls transcription of genes affecting lipid metabolism. CYP4A10
also plays a role in lipid metabolism. Treatment with peroxisome proliferators caused an
increase in both PH and PT. MCAD and VLCAD are mitochondrial fatty acid metabolizing
enzymes whose gene expression is mediated by PPARa (Aoyama et al., 1998). The results of
mRNA expression impacted by PFOA exposure are shown in Table 4-22.
TABLE 4-22. mRNA Expression of Hepatic 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
-
-
-
-*l
-
-
0.1
mg/kg
NA
-
-*l
-*l
-*l
-*l
-*l
0.3
mg/kg
NA
-
-*l
-*l
-*l
-
-*l
hPPARa
0
mg/kg
_*t
-
-
-*t
-
-
-*t
0.1
mg/kg
_*t
-
-*l
-
-*l
-*l
-n
0.3
mg/kg
.*t
-
-*l
-
-*l
+ t
-*l
From Nakamura et al., 2009
- Not different from respective control
+Significantly different from respective control
* Significantly different from mPPARa mice treated with same concentration
4-Decreased expression relative to respective control or mPPARa mice at same concentration
f-Increased expression relative to respective control or mPPARa mice at same concentration
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.
As expected, mRNA and protein expression of PPARa was absent in PPARa-null mice.
Treatment with 0.1 or 0.3 mg/kg PFOA did not alter mRNA or protein expression for any genes
investigated compared to control 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 of 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 was also 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
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significantly higher (p<0.05) in all hPPARa mice compared to mPPARa mice given the same
concentration of PFOA. Treatment of hPPARa mice with 0.1 or 0.3 mg/kg PFOA caused a
decrease (p<0.05) in mRNA expression of PH, VLCAD, and CYP4A10 compared to 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.
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
may 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 may be needed to cause
activiation 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 or 3 mg PFOA/kg on
GDs 1-17 by oral gavage (Albrecht et al., 2013). The study design and developmental toxicity
data are described in section 4.2.5. Females were either sacrificed on GDIS (n = 5-8/group) or
allowed to give birth and then sacrificed, along with their litters (n = 8-14), on PND20. Livers
from dams, fetuses, and pups were collected for measurement of mRNAs encoding the PPARa
target genes cytochrome P4504alO (Cyp4alO) and acyl-CoA oxidase 1 (Acoxl), the constitutive
androstane receptor (CAR) target gene (Cyp2blO), and the pregnane X receptor (PXR) target
gene (CypSall).
On GDIS, 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 Cyp3all were increased following PFOA administration in all three genotypes.
On PND20, maternal liver samples from treated groups showed increased expression of Acoxl in
wild-type mice; expression ofCyp2bJO was unchanged in all groups; and expression ofCypSall
was increased following PFOA administration in all three genotypes.
For fetuses on GDIS, liver samples from treated groups showed increased expression of
Acoxl and Cyp4alO in wild-type and humanized mice. Expression ofCyp2blO was
unchangeded following maternal PFOA administration in all three genotypes, while expression
ofCypSall was increased in humanized fetal liver. On PND20, 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 ofCypSall 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 Pgc-la -mediated pathway of mitochondrial
biogenesis: peroxisome proliferator-activated receptor gamma coactivator la (Pgc-la), estrogen-
related receptor a (Erra), nuclear respiratory factors 1 and 2 (Nrfl, Nrf2), transcription factor A
(Tfam), cytochrome c oxidase subunit II and IV (Cox II, Cox IV), NADH dehydrogenase 2
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(Nd2), and NADH dehydrogenase iron-sulfur protein 8 (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 were significantly increased 2 to 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 transcriptioinal level by
activation of the Pgc-la pathway.
Elcombe et al. (2010) examined the expression of some cytochrome P450 isoforms in the
livers of male Sprague-Dawley rats fed diets containing 300 ppm PFOA or 50 ppm Wyeth
14,643 for 1, 7, or 28 days. The isoforms included those involved in activation of 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 to PFOA. Treatment with Wyeth
14,643 caused the induction of CYP4A1 only.
PPAR Receptor Activation. Takacs and Abbott (2007) evaluated the potential for PFOA to
activate PPARs, using a transient transfection cell assay. Cos-1 cells, derived from the kidney
cells of the African green monkey, were transfected with mouse or human PPARa, P/5, or y
reporter plasmids and exposed to 0.5-100 |iM PFOA or 0.5-100 |iM PFOA and MK-886
(PPARa antagonist) or GW9662 (PPARy antagonist). An antagonist for PPARp/6 was not
available. The three types, PPARa, P/5 or y, are encoded by different genes, expressed in many
tissues and have specific roles during development as well as in the adult. The results are shown
in Table 4-23. PFOA activated mouse and human PPARa in a dose-dependent manner with a
significant increase in activity observed at 10, 20, 30, and 40 jiM in the mouse and 30 and 40 jiM
in the human 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. Activity of mouse PPAR.p/5 was significantly increased after exposure to 40-
80 jiM 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 P/5 under the
conditions in this study.
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TABLE 4-23. Activation of Mouse and Human PPAR by PFOA
PPARa
PFOA
(urn)
0
0.5
1
3
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
(um)
0
1
5
10
20
30
40
50
75
100
Mouse
-
-
-
-
-
-
-
-
-
-
Human
-
-
-
-
-
-
-
-
-
-
From Takacs and Abbott, 2007
+ Significant increase in activity between treated and control
- No difference in activity between treated and control
Biomarkers for Peroxisome Proliferation. Pastoor et al. (1987) dosed male Crl:CD (SD) BR
rats for 1, 3, or 7 days with 0 or 50 mg PFOA/kg. Hepatic DNA content, cytochrome P450
content, UDP-glucuronyltransferase, glutathione S-transferase, benzphetamine N-demethylase
activity (marker for smooth endoplasmic reticulum proliferation), and 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 carnitine aceyltransferase (CAT) and carnitine palmitoyltransferase (CPT) activity
assays. CAT served as a marker for peroxisome proliferation and CPT was a marker for
mitochondrial proliferation. Incorporation of [14C]acetate into hepatic lipids was used to
determine the effect of PFOA on hepatic lipid metabolism. Plasma total cholesterol and
triacylglycerides was determined from rats dosed for 7 days.
Hepatic DNA content was not increased in treated rats when compared to content in
control rats. Cytochrome P450 was significantly increased (p<0.05) and 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 2-fold (p<0.05)
after a 3-day treatment with 50 mg PFOA/kg. No differences were observed among the groups
for the other enzymes. No differences were observed between rats treated for 7 days and control
rats in plasma total cholesterol or triacylglycerol. Although a significant increase (p<0.05) was
observed for [14C]acetate incorporation into triacylglycerols, cholesteryl esters, and polar lipids,
there was no difference in the distribution of the incorporated label between control and treated
rats. The authors concluded that the lack of increased DNA content, proliferation of smooth
endoplasmic reticulum, and peroxisome proliferation pointed to increased liver weight due to
hepatocyte hypertrophy.
Cook et al. (1992) examined peroxisome p-oxidation activity in male CD rats gavage
dosed with for 14 days with 0, 1, 10, 25, or 50 mg PFOA/kg. Peroxisome P-oxidation activity
was significantly increased at >10 mg PFOA/kg compared to the control. At the highest dose, P-
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oxidation activity was significantly different from the pair fed controls.
Sohlenius et al. (1992) fed male and female Wistar rats diets containing 0 or 0.02%
PFOA for 7 days and examined hepatic peroxisome proliferation and related parameters. At
necropsy, liver samples were collected for analysis of catalase, peroxisomal lauroyl-CoA oxidase
activity, palmitoyl-CoA oxidation, DT-diaphorase, cytosolic epoxide, glutathione transferase
activity, co-hydroxylation of lauric acid, and protein concentration. In males, an increase was
observed in liver weight, liver somatic index, mitochondrial protein content, and palmitoyl CoA-
oxidation. No effects were observed in the females.
Sohlenius et al. (1992) fed male and female C57B1/6 mice diets containing 0 or 0.02%
PFOA for 7 days to examine possible sex differences in hepatic response to PFOA. Two
additional groups of female mice were fed a diet containing 0 or 0.05% PFOA for 5 or 10 days.
At necropsy, liver samples were collected for analysis of catalase, peroxisomal lauroyl-CoA
oxidase activity, palmitoyl-CoA oxidation, DT-diaphorase, cytosolic epoxide hydrolase,
glutathione transferase activity, co-hydroxylation of lauric acid, and protein concentration.
Liver weight, liver somatic index, mitochondrial protein, palmitoyl-CoA oxidation, and
lauroyl-CoA oxidase activity were significantly increased (p<0.05) in all treated mice.
Microsomal protein was significantly increased (p<0.05) in male mice fed diets of 0.02% PFOA
for 7 days. In male and female mice fed diets of 0.02% PFOA, catalase, DT-diaphorase, co-
hydroxylation of lauric acid, cytosolic epoxide hydrolase, and glutathione transferase activities
were significantly increased (p<0.05) compared to controls. The authors concluded that no sex
differences were apparent hepatic peroxisome proliferation and related parameters in mice fed
diets containing PFOA.
Elcombe et al. (2010) examined peroxisome proliferation in the livers of male Sprague-
Dawley rats exposed to 300 ppm PFOA in the diet for 1, 7, or 28 days (described in 7.2.2).
Analysis of palmitoyl-CoA oxidation showed significant increases (p<0.05) in /^-oxidation at all
time points in PFOA diet-fed rats compared to /^-oxidation in control rats.
Cellular Differentiation. Slotkin et al. (2008) characterized the neurotoxicity of PFOA using
PC12 cells. The cells were derived from a neuroendocrine tumor of the rat adrenal medulla and
serve as a model for neuronal development and differentiation. Exposure to nerve growth factors
causes PC12 cells to differentiate into cells expressing either dopamine or acetylcholine
phenotypes. The cells were incubated with 10, 50, 100, or 250 jiM PFOA. Synthesis of DNA,
cell viability, cell growth, and lipid peroxidation were measured to determine if PFOA targets
specific events in neural cell differentiation. Differentiation shifts towards or away from the
dopamine and acetylcholine phenotypes were measured by assessing the activities of tyrosine
hydroxylase (TH, dopamine) and choline acetyltransferase (ChAT, acetylcholine). The
undifferentiated cells were evaluated after a 24 hour exposure, and differentiating cells were
evaluated after 4-6 days of exposure to PFOA. The results are shown in Table 4-24.
Significant inhibition (p<0.0001) of DNA synthesis occurred in the undifferentiated cells
after exposure to 250 jiM PFOA with no change in DNA content. Lipid peroxidation was
significantly increased (p<0.02) after exposure to 10 jiM PFOA, and cell viability was
significantly decreased (p<0.03) after 24 hour exposure to 100 or 250 jiM PFOA.
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In differentiating PC12 cells, exposure to 250 jiM PFOA caused decreased DNA content
with no change in total protein/DNA content ratio or the membrane/total protein ratio. The
lowest and highest PFOA concentrations caused a significant increase (p<0.007) in lipid
peroxidation, but no effect was observed in cell viability. Tyrosine hydroxylase activity was
decreased (p<0.05) after exposure to 10 and 250 jiM PFOA, and the TH/ChAT ratio was
decreased (p<0.05) at 10 jiM PFOA. The results suggest that PFOA exposure caused the
differentiating cells to shift (slightly) to favor the acetylcholine phenotype.
TABLE 4-24. Impact of PFOA exposure on PC12 cells
Undifferentiated PC12
Cells
DNA Synthesis
DNA Content
Lipid Peroxidation
Cell Viability
Differentiating PC12
Cells
Lipid Peroxidation
Cell Viability
DNA Content
Total protein/ DNA content
ratio
Membrane/Total protein
TH activity
ChAT activity
TH/ChAT activity
PFOA (uM)
0
-
-
-
-
-
-
-
-
-
-
-
-
10
-
-
t
-
t
-
-
-
-
4
-
4
50
-
-
-
-
-
-
-
-
-
-
-
-
100
-
-
-
4
-
-
-
-
-
-
-
-
250
4
-
-
4
t
-
4
-
-
4
-
-
From Slotkin et al., 2008
t,| Significantly different from control (p<0.05)
- Not different from control
Ahuja et al. (2009) examined the effects of PFOA on the production and activation of
human monocyte-derived dendritic cells. These cells are responsible for a primary immune
system response by activating lymphocytes and secreting cytokines. Peripheral monocytes and
immature dendritic cells were incubated with 200 jiM PFOA from day 0 to day 6 or 8 to
determine the impact on phenotype and cytokine secretion. Maturation stimulus (prostaglandin
E2, tumor necrosis factor, interleukin ip, and interleukin 6) was added during the last 48 hours
of incubation to induce dendritic cell maturation. Mixed leukocyte reaction assays were
conducted to determine if immature dendritic cells could stimulate T cells. Cytokine (FTLA-DR,
CD25, CD80, CD83, and CD86) expression was measured as a determination of maturity. FTLA-
DR is a cytokine which presents antigens to elicit T-cell response. CD25, 80, 83, and 86 are cell
receptors that act as co-receptors in T-cell activation; and interleukin 12p40 and 10 stimulate T-
cells. Mature cytokine-activated dendritic cells secrete interleukin 12p40 and interleukin 10 as
anti-inflammatory cytokines.
In peripheral monocytes incubated with only PFOA from day 0-6 or day 0-8, expression
of FTLA-DR and CD86 was increased compared to expression in control cells indicating that
immature dendritic cells were present. In the mixed leukocyte reaction assay, the ability to
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stimulate T cells was not different between immature dendritic cells generated in the absence or
presence of PFOA.
To determine if PFOA impacted the differentiation of immature dendritic cells to mature
dendritic cells, immature dendritic cells were incubated with 200 jiM of PFOA for 6 days and the
maturation stimulus was added for the final 2 days of incubation. There were no differences in
cytokine (CD25, CD80, CD83, and CD86) expression between cells incubated with PFOA and
control cells. Expression of interleukin 12p40 and interleukin 10 was significantly inhibited by
PFOA in mature cytokine-activated dendritic cells, even in the presence of maturation stimulus
during the last 48 hours of incubation. The resulted suggested that exposure to PFOA during
generation of dendritic cells affected the phenotype and cytokine production of human dendritic
cells and could lead to immunomodulation in the development of the immune response.
Intracellular Communication. Upham et al. (1998, 2009) examined the effects of
perfluorinated fatty acids on gap junctional intercellular communication in male Fischer 344 rats
fed diets containing 0 or 0.02% PFOA (intake 37.9 mg/kg/day) for 1 week and in WB-F344 rat
liver epithelial cells. The chain lengths of the perfluorinated fatty acids ranged from 2-10, 16,
and 18 carbons. Liver weight in the rats fed diets containing 0.02% PFOA was significantly
increased compared to control rat liver weight. No differences were observed in serum AST,
ALT, and ALP. PFOA significantly inhibited gap junctional intercellular communication in the
livers of rats after treatment for 1 week. In WB-F344 cells, gap junctional intercellular
communication was inhibited by perfluorinated fatty acids with 7 to 10 carbons within 15
minutes of incubation. The inhibition was reversible with full recovery occurring within 30
minutes of PFOA removal from media. Extracellular receptor kinase was activated by PFOA
within 5 minutes of incubation in the cells. Preincubation of cells with the phosphatidylcholine-
specific phospholipase C inhibitor D609 partially prevented gap junction intercellular
communication inhibition by PFOA. The authors concluded that PFOA, having an 8 carbon
chain, inhibited gap junction intercellular communication by activation of extracellular receptor
kinase and phosphatidylcholine-specific phospholipase C, but noted that other mechanisms may
be involved.
Production of Reactive Oxygen Species. Takagi et al. (1991) fed male Fischer 344 rats diets
containing 0, 10 or 20 mg PFOA/kg for two weeks to determine the formation of 8-
hydroxydeoxyguanosine (8-OH-dG, marker of oxidative DNA damage). Livers and kidneys
were removed at necropsy and DNA was isolated from each organ and analyzed. The relative
liver and kidney weights were significantly increased (p<0.05) in the treated rats compared to the
control. A significant increase in 8-OH-dG liver levels was observed at > 10 mg PFOA/kg.
There were no significant differences in 8-OH-dG kidney levels between PFOA-treated and
control rats. The authors concluded that PFOA could cause organ specific oxidative DNA
damage.
Hu and Hu (2009) exposed human hepatoma cells, HepG2, to PFOA to evaluate
cytotoxic effects. Cells were also exposed to a mixture of PFOA and PFOS to determine
antagonistic or synergic effects. The cells were exposed to 0, 50,100, 150, or 200 jiM PFOA or
to 0, 50, 150, or 200 jiM each of PFOA and PFOS. A group of cells were also exposed to 0,
50,100, 150, or 200 |iM PFOS. The cells were cultured for 24, 48, or 72 hours. Cell viability,
apoptosis, reactive oxygen species, mitochondrial membrane potential, antioxidant enzymes,
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glutathione content, and differential expression of apoptosis gene regulators p53, Bax, Bcl-2,
caspace-3, and caspace-9 genes were evaluated.
Exposure to PFOA or PFOS caused a dose-dependent decrease in viability of HepG2
cells. A non-significant dose-dependent increase in apoptosis was observed in the cells cultured
with PFOA. However, the combination of both PFOA and PFOS showed a significant dose-
dependent increase (p<0.05) in apoptosis. Intracellular reactive oxygen species were
significantly increased (p<0.05) in cells cultured with 100, 150, or 200 |iM PFOA or PFOS.
HepG2 cells exposed to the mixture of 100 or 200 jiM 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 jiM PFOA
significantly decreased (p<0.05) glutathione concentration and glutathione peroxidase activity,
and 150 jiM PFOA significantly increased (p<0.05) the activities of superoxide dismutase,
catalase, and glutathione reductase, and significantly decreased (p<0.05) glutathione peroxidase
activity, and glutathione concentration in HepG2 cells. The trend was the same at 200 jiMol
PFOA with the exception of glutathione-^-transferase activity being significantly decreased
(p<0.05).
Exposure to PFOA did not change p53, Bax, or caspace-3 expression in HepG2 cells.
Expression of Bcl-2 was down-regulated and caspace-9 was up-regulated in a dose-dependent
manner in HepG2 cells following exposure to 50-200 u/Mol PFOA. The authors proposed that
PFOA and PFOS induced cell apoptosis by overwhelming the homeostasis of antioxidative
systems, increasing reactive oxygen species, impacting mitochondria, and changing gene
expression of apoptosis gene regulators.
Eriksen et al. (2010) examined ability of PFOA to generate reactive oxygen species
(ROS) and induce oxidative DNA damage in human HepG2 cells. Cell were incubated with 0,
0.4, 4, 40, 200, 400, 1000, or 2000 uM PFOA and 2',7'-dichlorofluorescein diacetate (DCFH-
DA). Hydrogen peroxide, H2O2, was used as a positive control. A fluorescence
spectrophotometer was used to measure ROS production every 15 minutes during the 3 hour
incubation period in all cultures. The comet assay was used to measure DNA damage in cells
exposed to 0, 100, and 400 uM PFOA for 24 hours. Cytoxicity was determined by measuring
the level of lactate dehydrogenase activity in the cell medium. Exposure to PFOA caused a
dose-independent increase (all doses p<0.05) in ROS production in HepG2 cells. Compared to
ROS production in negative control cells, PFOA induced a 1.52-fold increase in production.
There was no difference in oxidative DNA damage and lactase dehydrogenase activity between
PFOA-treated cells and negative control cells. The authors concluded that oxidative stress and
DNA damage were probably not relevant for potential adverse effects of PFOA.
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4.3.5 Structure-Activity Relationship
Bjork and Wallace (2009) compared the PPARa-dependent transcript!onal activation
potential of linear perfluorocarboxylid and sulfonic acids in rodent and human hepatic liver cells.
The PFAAs tested included perfluorinated carboxylic acids with carbon chain lengths of 2 to 8
and perfluorinated sulfonic acids with chain lengths of 4 to 8. Primary rat and human
hepatocytes and HEPG2/C3 A cells were incubated with 0 or 25 uM perfluorinated compounds
for 24 hours. Expression of acyl-CoA oxidase (Acox), Cyp4al (rat), Cyp4all (human), acyl-
CoA thioesterase (Cte-rat, Acotl-human), and DNA damage inducible (Ddit3; GADD153)
transcripts were determined by quantitative RT-PCR. All the genes are inducible by peroxisome
proliferators except Ddit3 which is induced in the presence of direct or indirect DNA damage.
Perfluorinated compounds induced mRNA expression of either Acox or Cte/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 suggested that the PPARa related
gene expression by perfluorinated compounds was induced in primary rat hepatocytes, increased
with carbon chain length and appeared to be greater in the carboxylic acids (i.e. PFOA) when
compared to the sulfonates (i.e. PFOS). There was no induction in expression of Acox and
Cte/Acot 1 in either primary or transformed human liver cells in culture. The authors suggested
that the PPARa mediated peroxisome proliferation observed in rodent liver may not be relevant
as an indicator to human risk.
Wolf et al. (2008b) tested PFAAs, including PFOA, to determine if mouse and human
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% dimethyl sulfoxide [DMSO]),
a positive control (WY14,643) or PFOS at 0.5-100 |iM. Other concentrations of PFAAs were
used but not provided in this report. At the end of 24-hours of exposure, the luciferase activity
was measured. The positive and negative controls had the expected results. A lowest observed
effect concentration (LOEC) and no observed effect concentration (NOEC) was determined for
each PFAA. In the study, the mouse 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 jiM in the mouse and 5 jiM in humans; the LOEC was 1 |iM
(0.43 |ig/mL) in the mouse and 10 jiM (4.3 jig/mL) in humans.
A similar study as conducted above (Wolf et al., 2008b) was repeated but included
additional PFAAs (Wolf et al., 2012). Transfected cells were incubated with PFAAs at
concentrations of 0.5 to 100 jiM, vehicle (water or 0.1% DMSO as negative control) or with 10
jiM WY14,643 (WY, positive control) on each plate. Assays were performed with 3 identical
plates per compound per species with 9 concentrations/plate and 8 wells/concentration. Cell
viability was assessed using the Cell Titer Blue cell viability kit and read in a fluorometer. The
positive and negative controls had the expected results. All human and mouse PPARa was
induced and was significant in all the PFAAs. Again, similar findings were identified with the
carboxylates and the mouse PPARa being more reactive. The study also provided the
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which was the concentration at which the PFAA produced 20% of the maximal response elicited
by the most active PFAA. For PFOA, this was 6 jiM in mouse PPARa and 7 jiM in human
PPARa. For comparison, PFOS was 94 jiM and 262 |iM, respectively.
4.4 Hazard Characterization
4.4.1 Synthesis and Evaluation of Major Noncancer Effects
Serum Lipids. Some epidemiological studies have noted an association between PFOA
exposure and increased serum lipid levels, a risk factor for cardiovascular disease in humans.
Because of the structural similarities between linear perfluorinated acids and the short and
medium chain fatty acids, the potential for these chemicals to cause elevated serum lipids has
been an area of considerable interest. Steenland et al. (2009) and Frisbee et al. (2010) examined
the relationship between serum PFOA and total cholesterol, high density lipoprotein (HDL)
cholesterol, and low density lipoprotein (LDL) cholesterol in participants of the C8 Health
project. Both studies reported positive associations between PFOA and total cholesterol.
Steenland et al. (2009) found significant increasing trends with increasing serum PFOA decile
for LDL and triglycerides but not HDL. Frisbee et al. (2010) found an association between
serum PFOA and total cholesterol and LDL in children and adolescents in the C8 study. Nelson
et al. (2010) found no associations between PFOA and LDL or VLDL in the general population.
A study of residents in the Little Hocking water district by Emmett et al. (2006) found no
association between PFOA and cholesterol. Occupational studies (Olsen et al., 2001a,b, 2003;
Sakr et al., 2007a,b; Costa et al., 2009) found positive significant associations between serum
PFOA and total cholesterol. These studies also found positive associations between serum
PFOA and VLDL and LDL (Sakr et al., 2007a), and triglycerides (Olsen et al., 2001 a,b, 2003;
Olsen and Zobel, 2007).
Taken together, the above studies show a positive association between serum PFOA and
cholesterol in both worker and general populations. Although not all studies showed an
association, positive results were reported in both cross-sectional and longitudinal studies.
Limited evidence also showed an increase in VLDL and LDL with higher serum PFOA
concentrations. The C8 Science Panel (2012) concluded that there is a probable link between
exposure to PFOA and diagnosed high cholesterol which is consistent with the data reviewed by
EPA.
High levels of serum lipids are a risk factor for cardiovascular disease in humans
including ischemic heart disease (IHD), a condition in which blood flow to the heart is decreased
through the development of atherosclerotic plaque or clots in the cardiac arteries. Sakr et al.
(2009) conducted a mortality study examining workers at the Washington Works plant and IHD.
The amount of time spent in relative job categories was used to estimate cumulative exposure.
There were over 4,000 employees in the mostly male cohort and 239 deaths from IHD. Mortality
from IHD was not associated with PFOA exposure. These results were similar to Leonard et al.
(2008) who found a non-significant elevation for IHD mortality when compared to the regional
worker population. Leonard et al. (2008) used 3 reference populations to derive his mortality
rates for comparison; the US population, the West Virginia state population, and an 8-state
regional employee population from the same company. Lundin et al. (2009) found no
association between serum PFOA and risk of dying from IHD in workers of the Cottage Grove,
Minnesota, plant compared to the general population of Minnesota.
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Information on serum lipids from animal studies have received less attention than in the
human population because of the fact that decreases in triglycerides, cholesterol, and lipoprotein
complexes are an expected consequence of PPARa activation in rodents. Activation of the
PPARa receptor leads to lowered serum lipids because of its induction of peroxisome
proliferation. Peroxisomes are subcellular organelles that increased beta oxidation of fatty acids.
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.
Cholesterol and/or triglycerides have been monitored in only a few of the animal studies.
Nakamura et al. (2009) and Minata et al. (2010) examined the lipid endpoints relative to the
mouse strain's 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
total cholesterol was significantly decreased (LOAEL= 10.8 mg/kg/day) and total triglycerides
significantly increased (LOAEL= 5.4 mg/kg/day) in wild type mice. In the PPARa null mice,
the total cholesterol was significantly decreased for the 5.4 and 10.8 mg/kg/day doses but
significantly increased for a 21.6 mg/kg//day dose while total triglycerides were significantly
increased at all doses.
Martin et al. (2007) identified a 45-72% decrease in serum cholesterol after treatment of
male Sprague Dawley rats with 20 mg PFOA/kg/day for up to 5 days, and Loveless et al. (2008)
reported decreased total cholesterol, HDL, and non-HDL in male CD rats after doses of 0.3 and 1
mg/kg/day for 28 days. Triglycerides were decreased in the rats at >0.3 mg/kg/day. In male CD-
1 mice, total cholesterol, HDL, and triglycerides were decreased at 10 and 30 mg/kg/day
(Loveless et al., 2008). In pregnant female ICR mice, triglyceride, total cholesterol, and free
fatty acids were significantly decreased at 10 mg/kg (Yahia et al., 2010). Elcombe et al. (2010)
found a significant decrease in cholesterol in male Sprague-Dawley rats following a 7 or 28 day
exposure to 300 ppm PFOA in the diet. Accordingly, there is not a high degree of concordance
between the lipidemic effects of PFOA as noted in human epidemiology studies and those seen
in animals.
Clinical Chemistry. Epidemiological studies examining clinical chemistry parameters have
reported an association between PFOA exposure and increased liver enzymes and decreased
bilirubin. With the exception of Olsen et al. (2000, 2001a,b, 2003), the occupational studies
reported positive associations between PFOA and increased GOT (Sakr et al., 2007a; Decatur
workers only, Olsen and Zobel, 2007; Costa et al., 2009), AST (Sakr et al., 2007b), ALP
(Decatur workers only, Olsen and Zobel, 2007), or ALT (Decatur workers only, Olsen and
Zobel, 2007, Costa et al., 2009). A negative association between serum PFOA and bilirubin was
also observed in some workers (Sakr et al., 2007b; Olsen and Zobel, 2007; Costa et al., 2009).
Examining the general adult population in the US, Lin et al. (2010) and Gallo et al. (2012) found
a slight, but significant association with ALT and GGT in some population groups and in
individuals with a higher BMI.
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Changes in renal enzymes and hematology were not associated with PFOA exposure in
worker populations (Olsen et al., 2000, 2001a,b, 2003; Sakr et al., 2007; Costa et al., 2009).
However, one study found the odds ratio for chronic kidney disease was positively correlated
with serum PFOA levels in the general population (Shankar et al., 2011). It is noted that a low
glomerular filtration rate would diminish the removal of PFOA from serum for excretion by the
kidney.
In animal studies, serum levels of ALT and/or AST were significantly increased
indicating apoptotic or necrotic damage to liver cells (Butenhoff et al., 2012; Son et al., 2008;
Minata et al., 2010). Increased levels of ALT were observed at a LOAEL of 2.65 mg/kg/day in
ICR mice by Son et al. (2008). Yahia et al. (2010) reported significantly increased ALT, GOT,
AST, and ALP in PFOA-exposed (10 mg/kg) pregnant ICR mice. Total protein, albumin, and
globulin were significantly decreased in the same mice.
Both the human and animal studies suggest effects on the liver as indicated by increases
in liver enzymes. The data from animal studies for increases in ALTand AST are more
definitive than the findings in human epidemiology studies. Concurrent with the evidence in
animals of damage to liver cells, levels of certain transport proteins have been shown to be
altered. In mice, increased expression of MRP3 and MRP4 (Maher et al., 2008) and decreased
expression of OATps (Cheng and Klaassen, 2008) favors excretion of PFOA into the bile but
may interfere with normal transport of biliary components.
Hepatic Hypertrophy. In most PFOA animal studies (monkeys, rats, mice), short term
and chronic exposure caused an increase in liver weight as at least one of the critical effects
(Christopher and Marias, 1977; Merrick and Marias, 1977; Goldenthal, 1978a; Pastoor et al.,
1987; Butenhoff et al., 2012; Palazzolo, 1993; Butenhoff et al., 2002, 2004a; York, 2002;
Perkins et al., 2004; Son et al., 2008; Wolf et al., 2008a; DeWitt et al., 2009; Elcombe et al.,
2010; Minata et al., 2010). Increased liver weights were observed in mice that are both active
and null for PPARa activation (Wolf et al., 2008a; Minata et al., 2010; Albrecht et al., 2013).
The histological characteristics of the liver differed in the mice with and without the 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 (Wolf et al., 2008a; Minata et
al., 2010). In the case of the wild-type PFOA-exposed mice, there was less endoplasmic
reticulum than in controls and more lipid-like vacuoles scattered throughout the cytoplasm. The
PFOA-exposed PPARa null mice had limited endoplasmic reticulum 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 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
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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.
Minata et al. (2010) reported degenerative histological changes in the bile ducts of
PPARa null mice at doses >10.6 mg/kg/day. PPARa null mice had an increased hepatocyte
proliferating cell nuclear antigen labeling index at a dose of 10 mg/kg/day (Wolf et al., 2008a).
When considering the studies in animals with and without the active PPARa receptor, it is clear
that PFOA has effects on the liver that appear to be independent of PPARa activation.
Other Organs. In general, effects on organs other than the liver tend to occur at doses higher
than those that affect the liver. Some studies have observed kidney and lung effects including
tubular epithelium swelling and pulmonary congestion in male Sprague Dawley rats (LOAEL =
20 mg/kg/day; Cui et al., 2009). In a chronic rat study, increased relative kidney weight was
observed at 5.64 mg/kg/day (Goldenthanl, 1978a). In a chronic monkey study (Goldenthal,
1978b), the critical effects were increased pituitary weight in males (LOAEL= 3 mg/kg/day) and
increased brain and decreased heart weight in females (LOAEL=10 mg/kg/day). The chronic
study by Butenhoff et al. (2012) noted hematological effects (decreased red blood cells counts
and hemoglobin concentrations) in treated male and female Sprague Dawley rats at a LOAEL of
24 mg/kg/day. Increased thickness and prominence of the zona glomerulosa and vacuolation
were observed in the cells of the adrenal cortex of male rats fed 10 mg/kg/day for approximately
56 days (York, 2002; Butenhoff et al., 2004a).
Hyperglycemia. Several human epidemiology studies have examined the association between
PFOA and hyperglycemia both in the general population or a subset of individuals with type II
diabetes. Leonard et al. (2008) examined mortality associated with diabetes in workers of the
Washington Works plant and found an association between the Washington Works employees
and the regional plant worker population (SMR=1.97, 95% CI 1.23, 2.98). Lundin et al. (2009)
found diabetes mortality to be associated with moderate PFOA exposure (Hazard Ratio 3.7, 95%
CI: 1.4-10.1) among Cottage Grove workers whose cause of death had been identified as
diabetes. No deaths from diabetes had occurred among workers with high exposure, based on job
classification. MacNeil et al. (2009) examined type 2 diabetes among adults from the C8 health
project and found no association found between serum PFOA concentration and elevated fasting
serum glucose levels or type II diabetes among members of the community in the vicinity of the
Washington Works plant. There are limitations in mortality analyses for disorders such as
diabetes where adherence to diet and other treatment modalities has a major impact on longevity.
Occupational studies examining glucose levels or related endpoints found no relationship to
PFOA concentration (Olsen et al., 2000, 2001a,b, 2003; Sakr et al., 2007a; Costa et al., 2009).
Lin et al. (2009) and Nelson et al. (2010) found no association between PFOA and metabolic
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syndrome, glucose concentration, insulin levels, and/or insulin resistance in the general US
population.
Hines et al. (2009) found no differences in glucose tolerance tests at 15-16 weeks and at
17 months of age in CD-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.
Hypertension. Serum uric acid levels were measured in two occupational populations and in the
C8 Health Project population as a biomarker for hypertension. Hypertension is a complex
condition related to a number of physiological conditions. Cholesterol buildup in the arteries is
one factor that can lead to elevated blood pressure as are kidney function and dietary sodium.
Impaired circulation can impact glomerular and tubular removal of PFOA and other substrates
excreted utilizing the common transporters in the renal tubules. Both occupational studies (Sakr
et al., 2007a, Costa et al., 2009) found an increased association with serum PFOA and uric acid
levels. The odds ratio for high serum uric acid levels increased from the lowest to the highest
quintile of PFOA serum concentrations in the C8 Health Project population (Steenland et al.,
2010). The study showed that higher PFOA concentrations were associated with higher
incidence of high serum uric acid, but uric acid levels remained within the normal range. A
positive association between chronic kidney disease and serum PFOA levels found in the general
population (Shankar et al., 2011) was independent of body mass index, hypertension and serum
cholesterol levels. No epidemiology data were identified that have examined the relationship
between serum PFOA levels and resting blood pressure, and none of the animal studies have
examined this end point.
Neurotoxicity. There are limited data on the neurotoxicity of PFOA. Fei et al. (2008b) found no
association between maternal PFOA concentration and fine motor skills, gross motor skills, and
cognitive abilities at 6 and 18 months of age. An epidemiology study of children derived from
the NHANES population found a weak statistical association between serum PFOA with parental
reports of ADHD (OR=1.12, 95% CI 1.01-1.23). However, one animal study (Johansson et al.,
2009) suggests a potential effect on habituation and activity patterns in NMRI mice treated on
PND10 with a single dose of PFOA and evaluated at and 2 and 4 months of age (LOAEL=0.58
mg/kg). The in vivo observations were supported by changes in the expression of a variety of
neurologically active brain proteins in the treated pups (Johansson et al., 2009). The offspring of
C57BL/6/Bkl dams fed 0.3 mg PFOA/kg/day throughout gestation had detectable levels of
PFOA in their brains at birth (Onishchenko et al., 2011). Behavioral assessments of the
offspring starting at 5 weeks of age revealed sex-related exploratory behavior differences. In the
social group setting, the PFOA-exposed males were more active and PFOA-exposed females
were less active than their respective controls. The PFOA-exposed males also had increased
activity counts than control males in circadian activity experiments. The results of an in vitro
study of hippocampal synaptic transmission and neurite growth in the presence of long chain
perfluorinated compounds showed that 50 or 100 uM PFOA increased spontaneous synaptic
current and had an equivocal impact on neurite growth (Liao et al., 2009a, b). These data
suggest a need for additional studies on the effects of PFAs, including PFOA, on the brain.
Thyroid Effects. A few epidemiological studies have examined the association between
presence of PFOA and altered thyroid homeostasis in humans. Thyroid-related measures among
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fluorochemical production workers were assessed by Olsen et al. (2001a, 2003), Sakr et al.
(2007a), Olsen and Zobel (2007), and Costa et al. (2009) relative to serum PFOA levels. Most of
these studies showed no associations between PFOA and any thyroid hormones. Olsen and
Zobel (2007) reported a negative association for free T4 and a positive association for T3;
however, the hormone levels were well within the normal reference range. The authors
concluded that there was no consistent relationship between altered thyroid hormone status and
serum PFOA.
Emmett et al. (2006) found no association between serum PFOA and thyroid hormones
and/or disease in residents served by the Little Hocking water district. Results from a study by
Bloom et al. (2010) did not identify an association between thyroid function and non-
occupational PFOA exposure (mean serum concentration of 1.3 ng/ml; 95% CI: 1.15-1.53).
TSH and free T4 levels were used as markers for thyroid function. Only one subject (3.2%)
exceeded the laboratory reference range for normal (0.40-5.00 jiIU/mL) and no subjects
exceeded the laboratory reference range for normal free T4 (0.80-1.80 ng/dL). Pirali et al.
(2009) reported no relationship between intrathyroidal concentrations of PFOA and underlying
thyroid disease.
Chan et al. (2011) found no association between serum PFOA concentrations and
hypothyroxinemia in pregnant women. More recently, Melzer et al. (2010) examined the records
of the NHANES data for thyroid disease and serum PFOA levels in the U.S. general population
and found that women with PFOA > 5.7 ng/ml were more likely to report current thyroid disease
compared to women with PFOA < 4.0 ng/ml. A near significant and similar trend was observed
in men. However, lack of segregating the thyroid disorders into those associated with
hyperactivity versus hypoactivity was a limitation in their analysis. The physiological
significance of these cross-sectional analyses remains to be determined.
Effects of PFOA on thyroid hormones in animals are generally not as well characterized
as those of PFOS. Butenhoff et al. (2002) evaluated the toxicity of PFOA in male Cynomolgus
monkeys during 6 months of oral administration and reported that levels of total T3 and free T3
in circulation were reduced significantly in the 30/20 mg/kg/day treatment group, an effect seen
as early as 5 weeks after initiation of treatment. Recovery of T3 deficits was noted upon
cessation of chemical treatment after serum level of PFOA returned to baseline 90 days later.
Serum total T4, free T4 and TSH were not altered throughout the study. The preferential effects
of PFOA on serum T3 and a lack of TSH compensatory response are similar to those observed
with PFOS.
Martin et al. (2007) showed that serum total T4 and free T4 were markedly and abruptly
(1 day after oral gavage treatment) depressed (~ 80%) by PFOA (20 mg/kg) in adult male rats;
while serum T3 was also reduced (25%) though to a lesser extent. These findings were
confirmed when both male and female rats were given PFOA (10 mg/kg) daily for 3 weeks and
serum thyroid hormones were monitored (Lau, personal communication). Serum total T4 and
free T4 were profoundly depressed (>85%) and T3 less so (~ 25%) in male rats, but serum TSH
levels were not altered significantly. These hormonal changes were noted when serum PFOA
level reached about 67 ug/ml. The dose-response relationship of serum total T4 and PFOA has
yet to be evaluated and the lowest effective dose of the chemical remains unknown.
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None of the thyroid hormones were affected by PFOA in mature female rats, primarily
because these animals were able to clear the chemical effectively (with half-life estimate of 2-4
hours compared to that of 6-7 days for male rats). This suggests that the thyroid disrupting
effects of PFOA are directly related to endogenous accumulation of the chemical and may be
relevant to humans because of the long PFOA human half-life.
Displacement of T4 from binding to the hormone transport protein transthyretin (TTR)
has been proposed as a possible mechanism to account for the hypothyroxinemia in rats.
However, although PFOA binds to human TTR, its binding affinity is only one-fifteenth of that
of the natural ligand T4 (Weiss et al., 2009). Based on a toxicogenomic analysis of rat liver after
an acute exposure to PFOA, Martin et al. (2007) suggested a possible role of peroxisome
proliferators in the thyroid hormone imbalance, although this hypothesis has yet to be explored
in detail.
Reproductive and Developmental Effects. There is a wealth of data on the impact of PFOA
exposure on reproductive and/or developmental effects in both humans and animals. A series of
epidemiological studies have been published examining a variety of endpoint: fertility
(fecundity), growth, and developmental biomarkers or milestones. The animal data have focused
more extensively on developmental endpoints using traditional and non-traditional approaches.
Joensen et al. (2009) found that there was no association between serum levels of PFOA
and testicular hormones among 105 Danish men who provided blood and semen for the military
draft in 2003. A nonsignificant negative association was observed between serum PFOA
concentration and semen volume, sperm concentration, sperm count, sperm motility, or sperm
morphology. Fei et al. (2009) aimed to measure fecundity by using longer time to pregnancy as
the outcome. More than 1200 women were included in a phone survey. Infertility was defined
as having > 12months time to pregnancy (TTP); pregnancies were separated into: planned,
partially planned, and not planned. The fecundity odds ratios for the three highest PFOA
concentration quartiles were 0.72, 0.73, and 0.60. The likelihood ratio test for trend was
significant (p< .0001). The results provide suggestive evidence that PFOA concentration may
reduce fecundity, however, the results should be viewed with caution, due to potential selection
bias and/or confounding biological and behavioral factors.
Several studies have looked at the association between PFOA concentration and fetal
growth (birth weight, birth length, head/chest circumference, etc); whereas some researchers
found no significant associations, the results for others were positive. Monroy et al. (2008)
found no association between umbilical cord or serum PFOA concentration and birth weight in
105 mothers in the Family Study conducted in Canada. These findings were supported by Nolan
et al. (2009) who examined mothers and babies who were serviced by the Little Hocking Water
Association (LHWA). Many researchers used log-transformed values and adjusted for possible
confounders, but still found no significant association between PFOA and birth weight, birth
length, head, or chest circumference (Hamm et al., 2009, Washino et al., 2009). Conflicting
results were seen in the early studies from Fei et al (2007, 2008a) where there was a statistically
significant inverse association between birth weight and plasma PFOA concentration among
"normal-weight" women. A statistically significant decrease in birth length was also observed
when the second and fourth quartile measures of PFOA exposure were compared to those for the
first quartile.
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In studies that examined pregnancy outcome among women believed to be exposed to
PFOA, there were no association seen between congenital anomalies and complications of labor
and delivery such as meconium, prolonged labor, breach, cord prolapse, or fetal distress (Nolan
et al., 2010). There was also no association between PFOA and birth outcomes including
miscarriages and livebirths (Stein et al., 2009). Knox et al. (2011) suggested that PFOA may be
associated with endocrine disruption in women after finding that the odds of having experience
menopause were significantly higher in the highest PFOA quintile group relative to the lowest
PFOA group.
Fei et al. (2008b) found no association between maternal PFOA concentration and fine
motor skills, gross motor skills, cognitive skills, at 6 and 18 months of age. The age at which
children reached developmental milestones did not show any relationship to maternal plasma
PFOA concentration. Maternal plasma PFOA concentration was also not associated with child
risk of hospitalization for infectious diseases (Fei et al., 2010), but was associated with decreased
antibody liters in children following vaccination (Grandjean et al., 2012). In the general
adolescent population in the US, a positive association was observed between serum PFOA and
parent-reported ADHD (Hoffman et al., 2010). Lopez-Espinosa et al. (2011) reported reduced
odds of having reached puberty in girls included in the C8 Health project. Christensen et al.
(2011) found no association between puberty and PFOA exposure in children of the Avon
Longitudinal Study of Parents and Children in the United Kingdom. Halldorsson et al. (2012)
found that low dose developmental exposures to PFOA resulted in obesogenic effects in female
offspring at 20 years.
Among the animal studies there was no effect of PFOA on reproductive or fertility
parameters in rats, but some effects were observed in mice. Both species, however, showed
some indications of the potential developmental toxicity of PFOA. Doses which elicited a
response were much higher in rats compared with those in mice. The species differences in
response and dose rate are likely related to half-life differences of hours for the female rat and
days-to-weeks for the female mouse.
Body weight gain was decreased in parental male animals in a two-generation study after
dietary exposure to 30 mg/kg/day PFOA (York, 2002; Butenhoff et al., 2004a, York et al., 2010).
Maternal body weight gain was reduced at >30 mg/kg/day and death occurred at > 100 mg/kg/day
in rats given PFOA via gavage during gestation (Staples et al., 1984; Hinderliter et al., 2005;
Mylchreest, 2003). In mouse gavage studies, increased time to parturition was observed at >10
mg/kg/day, and decreased maternal weight gain was observed at >20 mg/kg/day (Lau et al.,
2006). Mouse maternal mammary glands displayed stunted alveolar development (GDIS) and
delayed epithelial differentiation (PND10) (White et al., 2009).
Decreased fetal and offspring body weight was observed in rats following an inhalation
exposure to 25 mg/m3 for 6 hours/day on GD6-15 (Staples et al., 1984). A two-generation diet
study in rats resulted in significantly decreased body weight gain prior to weaning and delayed
sexual maturity in the first generation males and females at 30 mg/kg/day (York, 2002;
Butenhoff et al., 2004a). In mouse gavage studies, decreased body weight and decreased
neonatal survival were observed at > 1 mg/kg/day and increased full litter resorptions and
increased stillbirths were observed at > 5 mg/kg/day for exposures lasting from GDI to 17 (Lau
et al., 2006; Abbott et al., 2007; White et al., 2007; Wolf et al., 2007).
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Delayed mammary gland development of female pups was also observed following
maternal doses > 0.01 mg PFOA/kg in one study (Macon et al., 2011) and >3 mg/kg in another
study (White et al., 2009). However, Albrecht et al. (2013) found no differences in the average
length of mammary gland ducts and the average number of terminal end buds per mammary
gland per litter in female pups following a maternal dose of 3 mg/kg.
Two developmental studies with compared wild-type mice with PPARa null mice but
results are inconclsive. 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 is involved in the
etiology of PFOA-induced neonatal mortality.
Cross-fostering studies showed that gestational PFOA exposure was required to produce
decreased body weight, delayed eye opening, delayed body hair growth, and decreased survival
in the offspring. However, persistent (PND1-PND63) mammary gland development delays were
observed in offspring that had been gestationally or lactationally exposed as well as in offspring
who had been gestationally and lactationally exposed (White et al., 2009). Restricted exposure
studies showed that gestational exposure to PFOA during a specific time period led to differeing
offspring effects (Wolf et al., 2007). Delayed eye opening and body hair growth were observed
at 5 mg/kg/day in offspring exposed GD7-17 or 10-17, but decreased postnatal survival was
observed at the same dose in offspring exposed GD15-17. Delayed mammary gland
development was observed irrespective of exposure duration, and delayed lactational
differentiation was seen in the dams (White et al., 2009). In a three generation study, PFOA
induced delays in mammary gland development and/or lactational differentiation. A chronic
dose of 5 ppm PFOA in the drinking water also altered mammary morphology in mice (White et
al., 2011). The mode of action for PFOA-induced delayed mammary gland development is
unknown.
Delayed vaginal opening was observed in BALB/c mice who received 1 mg/kg/day by
gavage for four weeks starting on PND21 (Yang et al., 2009). At higher doses (5 and 10
mg/kg/day), vaginal opening did not occur and mammary gland development was delayed. In
C57BL/6 mice, vaginal opening was delayed at 5 mg/kg/day and did not occur at 10 mg/kg/day.
Mammary gland development was accelerated at 5 mg/kg/day, but delayed at 10 mg/kg/day
indicating strain differences in pubertal mammary gland development following a dose of 5
mg/kg/day. Zhao et al. (2010) showed that 5 mg PFOA/kg/day stimulates mammary gland
development in C57BL/6 mice by promoting steroid hormone production in the ovaries and
increasing mammary gland growth factor levels.
Immunotoxicity. An occupational study of the immunotoxicity of PFOA found no association
between the immunoglobins IgA, IgG, and IgM and PFOA in male workers (Costa et al., 2009).
No associations were found between maternal PFOA concentrations and hospitalization for
infectious diseases in early childhood (Fei et al., 2010b) or infant allergies (Okada et al., 2012).
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In contrast, higher PFOA concentrations in children were associated with lower antibody levels
to diphtheria and tetnus (Grandjean et al., 2012) and a greater risk of having asthma (Dong et al.,
2013). In a pilot study using volunteer donated blood, Brieger et al. (2011) found that PFOA
altered peripheral blood mononuclear cell viability and reported a significant association was
observed between PFOA concentration and the release of LPS-induced TNF-a and IL-6 by
peripheral monocytes.
Several animal studies demonstrate effects on the spleen and thymus as well as their
cellular products (B lymphocytes and T-helper cells) in several strains of mice. Studies by Yang
et al. (2000, 2001, 2002b) and DeWitt et al. (2008) were conducted using relatively high PFOA
doses (-30-40 mg/kg/day). In each study the PFOA treated animals exhibited significant
decreases in spleen and thymus weights as well as splenocyte and thymocyte populations at
various stages of differentiation. Recovery usually occurred within several days of cessation of
PFOA dosing.
Activation of the PPARa receptor appears to be related to at least a part of the thymus
response to PFOA. When Yang (2002a) dosed PPARa null mice with the same dose that caused
an effect in the wild-type mice, no significant changes in spleen weight or cellularity were
observed, but there was a small and significant decrease in thymus weight and cellularity
compared to controls.
In one component of the Yang et al. (2002b) study, the functional impact of changes in
spleen and thymus were evaluated through the response of treated mice to horse red blood cells
(HRBC). The control mice responded to the HRBC exposure with an increased plaque forming
response; however, the PFOA treated mice did not have an increased plaque forming response
when tested (Yang et al., 2002b). In addition, when serum from PFOA treated mice was
evaluated post treatment, there was no increase in lymphocyte proliferation in response to the
addition of Con-A and LPS to the test media. The control mice responded with the expected
lymphocyte proliferation after the addition of Con-A and LPS antigens.
DeWitt et al. (2008) used different functionality assays in their study in C57B1/6 mice.
The IgM response to sheep red blood cells (SRBC) was suppressed by 20% when mice were
immunized immediately after exposure to the initial dose of 30 mg PFOA/kg/day ceased.
However, there was no significant increase in the response to BSA four days post PFOA
exposure, or in the IgG response to SRBC 15 days post PFOA exposure. These results are
indicative of recovery once PFOA exposures have ceased.
DeWitt et al. (2008) followed their initial study of PFOA with one designed to examine
the dose response for a 15 day drinking water exposure in a slightly different mouse strain,
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 at the NOAEL of 1.88 mg/kg/day.
Loveless et al. (2008) looked at the IgM response to SRBC in male CD rats and CD-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,
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however, thymus and spleen weights, thymus and spleen cell counts, and IgM liters were
decreased at >10 mg PFOA/kg/day. Corticosterone was also increased in mice at the same
doses.
The data collected from the immunotoxicity studies support a mode of action through
which PFOA interferes with splenocyte and thymocyte precursor cells in the bone marrow as
well as maturation of the cells once they have been transported to their respective organs.
Examination of cell populations at different stages of development reveals lower numbers of the
CD4"CD8" cells formed in bone marrow as well as decreased populations of splenocyte and
thymocyte cells at different stages of expressing the surface proteins that mark them as
functional beta lymphocytes (thymus) or T-helper cells (spleen) (Son et al., 2009). Although the
studies that measured the splenocyte and thymocyte populations were carried out at doses higher
than the 3.75 mg/kg/day LOAEL observed by DeWitt et al. (2008), the fact that the IgM
response to an antigenic material was decreased at that dose indicates an inability to produce
antibodies at adequate levels when exposed to a challenge.
Loveless et al. (2008) hypothesized that the observed effects on serum lymphocytes could
be the result of adenocorticotropic steroids in a response to stress. A study by DeWitt et al.
(2009) in which the immunological response of adrenalectomized mice treated with PFOA were
compared to sham operated controls did not support the Loveless et al. (2008) hypothesis.
Data from PPARa null mice suggest that rodents may be more susceptible to the
immunosuppressive impacts of PFOA than humans. However, the fact that there were still
effects on the thymus weight and cellularity even in the null mouse strain indicate the potential
for an inadequate humoral response in exposed populations.
4.4.2 Synthesis and Evaluation of Carcinogenic Effects
A number of human epidemiology studies were conducted on worker populations to
examine an association between PFOA exposure and cancer. Lundin et al. (2009) saw no
association between cancer mortality and exposure to PFOA when comparing workers of the
Cottage Grove facility to the Minnesota general population. The mortality data for liver cancer,
pancreatic cancer, prostate cancer, and testicular cancer all showed no association to exposure.
The Dupont (2003) report on the West Virginia Washington Works Plant found statistically
significant elevations in cancer of the bladder (SIR 1.9; 95% CI: 1.15-3.07) and kidney and
urinary organs (SIR 2.3; 95%CI: 1.36-3.65). This same report found non-statistically significant
elevations in leukemia incidence.
Two analyses of leukemia incidence were conducted prior to the Dupont (2003) report.
Walrath and Burke (1989) reported statistically significant elevated Odds Ratio of 2.1 for
leukemia incidence in male employees. Karns and Fayerweather (1991) conducted a follow up
case-control study at the same Washington Works facility where they matched employees (four
controls for each case). The matched Odds Ratios were significantly elevated for employees
who had previously worked as custodians and engineers, 8.0 (90% CI, 1.1-60.0) and 7.9 (90%
CI, 1.0-76.0), respectively. The level of significance chosen by the authors for each p=. 10 were
not comparable in stringency to those used in many of the other papers; also the CIs were quite
wide even at the 90%level.
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Two human epidemiology studies focused on general populations, one of which was
exposed to PFOA through contaminatd water supplies. First, Eriksen et al. (2009) looked at risk
of cancer in the Danish general population and plasma PFOA concentration. The Incidence Rate
Ratios (IRRs) crude and adjusted showed no association between plasma PFOA concentration
and prostate, bladder, pancreatic, and liver cancer. Vieira et al. (2013) showed a positive
association between PFOA levels and kidney and testicular cancers in residents living in an area
with contaminated water. Positive associations were found between kidney cancer and the very
high and high serum exposure categories (OR: 2.0, 95% CI: 1.0, 3.9; n=9 and OR: 2.0, 95% CI:
1.3, 3.2; n=22, respectively) and between for testicular cancer with the very high exposure
category (OR: 2.8, 95% CI: 0.8, 9.2; n=6). No associations for either cancer were found in the
lower exposure categories. Ovarian cancer, non-Hodgkin's lymphoma, and prostrate cancer
were also positively associated with the very high exposure category, but showed weaker or
negative associations for the other exposure categories.
Two animal carcinogenicity studies have shown that PFOA exposure can lead to liver
adenomas or carcinomas (Butenhoff et al., 2012; Biegel et al., 2001), Leydig cell adenomas
(Butenhoff et al., 2012; Biegel et al., 2001), and pancreatic acinar cell tumors (PACT) (Biegel et
al., 2001) in male Sprague-Dawley rats. Liver adenomas were observed in the Biegel et al.
(2001) study at an incidence of 10/76 (13%) at 20 mg/kg/day. The incidence in the control group
was 2/80 (3%). Although no liver adenomas were observed in Butenhoff et al. (2012),
carcinomas were identified in both the male controls and exposed male and female rats in the
high dose group and exposed males in the low-dose group. The differences from control were
not significant in either case but the incidence among the high dose males (10/50) was similar to
that for the adenomas in the Biegel et al. (2001) study. Liver lesions were identified in the males
and females at the one- and two-year sacrifices (Butenhoff et al., 2012). Increased incidence of
diffuse hepatomegalocytosis and hepatocellular necrosis were occurred at 20 mg/kg/day. At the
2-year sacrifice, hepatic cystoid degeneration was observed at 8, 14, and 56% in males of the
control, 2, and 20 mg/kg/day dose groups, respectively. Hyperplastic nodules in male livers
were increased in the high dose group (6% vs. 0% in control rats).
Testicular Leydig cell tumors (LCT) were identified in both the Butenhoff et al. (2012)
and Biegel et al. (2001) studies. The tumor incidence was 0/50 (0%), 2/50 (4%), and 7/50 (14%)
for the control, 2.0, and 20 mg/kg/day dose groups, respectively (Butenhoff et al., 2012). The
Biegel et al. (2001) study included one dose group (20 mg/kg/day), and the tumor incidence was
8/76 (11%) compared to 0/80 (0%) in the control group. LCT incidence at 20 mg/kg/day was
comparable between the two studies. PACTs were only observed in the Biegel et al. (2001)
study at an incidence of 8/76 (11%; 7 adenoma, 1 carcinoma) at 20 mg/kg/day while none were
observed in the control animals. Although no PACTs were observed by Butenhoff et al. (2012),
pancreatic acinar hyperplasia was observed at 2 and 20 mg/kg/day; 2/34 (6%) and 1/43 (2%),
respectively. Re-examination of the pancreatic lesions in Butenhoff et al. (2012) and Biegel et
al. (2001) resulted in the conclusion that 20 mg/kg/day increased the incidence of proliferative
acinar cell lesions in both studies, and the lesions in the Biegel et al. (2001) study had progressed
to adenomas.
The initial findings from the Butenhoff et al. (2012) study were equivocal for mammary
fibroadenomas in female rats. However, a re-examination of the tissues by a PWG found no
statistically significant differences in the incidence of fibroadenomas or other neoplasms of the
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mammary gland between control and treated animals (Hardisty et al., 2010). The PWG used the
diagnostic criteria and nomenclature of the Society of Toxicological Pathologists for the re-
examination. Under those criteria, there was an increase in the number of tumors documented in
the control group, especially fibroadenomas originally classified as lobular hyperplasia. The
reclassification led to a loss of significance when the tumors in the treated animals were
compared to tumors in the control animals.
Ovarian tubular hyperplasia and adenomas were also observed in female rats (Butenhoff
et al., 2012). Mann and Frame (2004) re-examined the ovarian lesions using an updated
nomenclature system which resulted in some of the hyperplastic lesions being reclassified. The
ovarian lesions originally described as tubular hyperplasia or tubular adenomas were regarded as
gonadal stromal hyperplasia and/or adenomas. After the reclassification there were no
statistically significant increases in hyperplasia (total number), adenomas, or
hyperplasia/adenoma combined in treated groups compared to controls.
Mutagenicity studies of PFOA using the S. typhimurium (Lawlor, 1995, 1996; Friere et
al., 2008) and E. coli (Lawlor 1995, 1996) system have resulted in negative results in the
presence and absence of activation. One mutagenicity study (Lawlor, 1995, 1996) in S.
typhimurium gave a positive result but it was not reproducible. In clastogenicity studies in CHO
by Murli (1996b, 1996d), the results were positive with activation for chromosomal
abnormalities and polyploidy and equivocal in the absence of activation. Micronucleus assays
by Murli (1995, 1996a) were negative.
A significant increase in 8-OH-dG liver levels, a biomarker for oxidative stress, was
observed at > 10 mg PFOA/kg in the liver but not the kidney of Fischer 344 male rats by Takagi
et al. (1991). Work with HepG2 cells by Hu and Hu (2009) suggested that PFOA could induce
apoptosis by overwhelming the homeostasis of antioxidative systems, increasing reactive oxygen
species, impacting mitochondria, and changing expression of apoptosis gene regulators. Erikson
et al. (2010) observed a PFOA-induced increase in reactive oxygen species production in HepG2
cells, but no PFOA-induced oxidative DNA damage or cytotoxicity.
4.4.3 Mode of Action and Implications in Cancer Assessment
The modes of lexicological/carcinogenic action of PFOA are not clearly understood.
However, available data suggest that the induction of tumors is likely due to non-genotoxic
mechanism involving receptors and perturbations of the endocrine system.
Rat Liver Tumors. PPARa agonism is the proposed mechanism of action (MOA) for the liver
carcinomas and adenomas in rats following chronic PFOA exposure (Maloney and Waxman,
1999; Klaunig et al., 2003). In the PPARa agonism MOA (Figure 4-1), binding of PFOA to the
PPARa receptor results in increased peroxisome proliferation and cell replication. Peroxisome
are single-membrane organelles found in a number of plant and animal cells that have the
capacity to carry out oxidation of long chain fatty acids and other substrates through hydrogen
peroxide generating pathways (Goodrich and Sul, 2000). Peroxisome beta oxidation produces
acetyl CoA that can be utilized for anabolic reactions but is not linked to ATP production.
Peroxisomes metabolize the long chain fatty acids via both beta and omega oxidation pathways
(Fielding, 2000).
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When a chemical binds to and activates the 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).
The key events in the MOA of PPARa-agonist induced liver tumors involve four causal
key events (Klaunig, 2003). The first key event is activation of PPARa. Other key events are
cell proliferation, apoptosis, and clonal expansion. Of these key events, PPARa activation is
highly specific for this MOA whereas cell proliferation, apoptosis, and clonal expansion are
common for other MO As.
Key Events in the Mode of Action for PPARa-
Agonist Induced Rodent Liver Tumors
PPARa Ago net
Causative Events
ActrvattonofPPAR a
Cell Proliferation
Decreased Apoptosis
Associative Events*
•Expression of Peroxisomal Genes
•Increase in Peroxisomes (number
&size)
Preneoplastic Foci
Clonal Expansion
Liver Tumors
Although thee are olher biological events (eg, Kupffer cell mediated eve nts^ inNbion of gap jure tone), the
rreasuiemenB of peroxiaome prdifeiationandpeioaaomalefE/ne activity (n particular acyl-CoAJare widdyusedas
reliable markers of PF^Ra acfivatioi.
From U.S. EPA, 2005b
FIGURE 4-1. PPARa-Agonism Mode of Action for Liver Tumors
The primary data that demonstrate PFOA activation of the PPARa receptor are those
from Rosen et al (2008a,b) 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, and evidence that PFOA also acts through PPARa independent
mechanisms associated with CAR and PXR receptors. Martin et al. (2007) examined the
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genomic signature from PFOA-treated Sprague-Dawley rats (up to 5-day exposure) using
microarray expression profiling, pathway analysis, and quantitative-PCR. The animal responses
were consistent with PPARa agonism, but there was also evidence of PPARy agonism (down-
regulation of cholesterol synthesis) and activation of CAR and PXR-related genes.
Few studies examined the occurrence of steps in the PPARa agonism MOA beyond
activation of PPARa. The one study that looked at apoptosis from a MOA perspective (Elcombe
et al, 2010) failed to see a significant decrease in male Sprague-Dawley rats with a 28 day
exposure to a diet containing 300 ppm PFOA (comparable to the high dose in both cancer
studies). The data in mice were not consistent. Minata et al. (2010) observed increased apoptosis
in hepatocytes, arterial walls, and bile duct epithelium of wild type 129S4/SvlmJ mice and in the
bile duct epithelium of PPARa-null mice at 10.8 or 21.6 mg/kg PFOA. On the other hand, Son
et al. (2008) saw evidence of decreased apoptosis in liver and kidney cells stained for caspaceS
in 4 week old male ICR mice treated for 21 days.
There are studies in male rats that examined cell proliferation. Using a BrdU labeling
technique, Elcombe et al. (2010) observed significant increases in cell proliferation in male
Sprague-Dawley rats after 1, 7, and 28 days of exposure to a 300 ppm dietary dose. The highest
increase was observed after 7 days of treatment (5-fold increase) which had declined to a 2-fold
increase after 28 days of dosing. The liver results from the Biegel et al. (2001) mechanistic
study were negative for cell proliferation in male Sprague-Dawley rats exposed to the same
dietary concentration (20 mg/kg/day) and sacrificed at 1, 3, 6, 9, 12, or 15 months. However,
based on the Elcombe et al. (2010) observations, the timing of the interim sacrifice would have
missed the peak of the proliferative response. The Butenhoff et al. (2012) study identified
hyperplastic nodules in 6/50 high dose males and 4/50 high dose females at 20 mg/kg/day; 10/50
males and 2/50 females had hepatocellular carcinomas.
The study by Wolf et al. (2008) looked at the labeling index in 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 but only at 10 mg/kg/day, the highest
dose, in PPARa-null mice. This suggests PPARa involvement in increased cell division of wild-
type mice at lower doses.
There were no studies identified that focused specifically on preneoplastic foci and clonal
expansion of altered cells after PPAR exposures. Minata et al (2010) observed a dose dependent
increase in eosinophilic cytoplasmic changes consistent with peroxisome proliferation in liver
parenchyma, but found no focal necrosis at doses < 21.6 mg/kg/day in wild-type 129S4/SvlmJ
mice.
There are several "associative" events that are markers of PPARa agonism but are not
directly involved in the etiology of liver tumors. These include peroxisome proliferation and
expression of genes for enzymes associated with the peroxisomes. These associative events
include (Klaunig et al., 2003):
• In vivo evidence of an increase in number and size of peroxisomes,
• Increases in the activity of acyl CoA oxidases such as palmitoyl CoA oxidase activity,
• Hepatocyte hypertrophy, increase in liver weights,
• Hepatic CYP4A1 (Laurie acids monooxygenase) induction, and
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• Peroxisomal gene expression.
There is ample evidence that PFOA is a potent peroxisome proliferator inducing peroxisome
formation in the livers of rats and mice (Ikeda et al, 1985; Pastoor et al., 1987; Sohlenius et al.,
1992; Yang et al., 2001; Wof et al., 2008; Minata et al., 2010; Elcombe et al., 2010). PFOA was
also found to activate mouse and human PPARa using a transient transfection cell assay (Takacs
and Abbott (2007). Maloney and Waxman (1999) demonstrated that 5-10 uM PFOA activated
mouse PPARa using COS1 cells (kidney fibroblast derived cells) transfected with a luciferase
reporter gene.
The temporal and dose-response relationship of measures of peroxisome proliferation,
hepatocellular hypertrophy, liver weight, and liver histopathology have been examined in male
Sprague-Dawley rats following 4, 7 and 13 weeks of administration of dietary PFOA at doses
ranging from 0-6.5 mg/kg-day (Palazzolo, 1993; Perkins et al., 2004). There was no evidence of
peroxisome proliferation, hepatocellular hypertrophy or liver weight increases at 0.06 mg/kg-
day. At doses ranging from 0.64 to 6.5 mg/kg-day there is a clear relationship between
peroxisome proliferation (indicated by increased palmitoyl CoA oxidase activity), hepatocellular
hypertrophy and increases in liver weight at all time points.
One can conclude, based on the available data, that the liver tumors in the PFOA
Sprague-Dawley cancer bioassays can be attributed to its peroxisome proliferative impact since
there are data that supports each of the steps in the proposed mode of action and thus are not
relevant to humans (Klaunig et al., 2003).
Leydig Cell Tumors (LCT). The mode of carcinogenic action for PFOA-induced LCTs
identified in Biegel et al. (2001) has not been fully elucidated. A large number of non-genotoxic
compounds of diverse chemical structures have been reported to induce LCTs in rats, mice, or
dogs. LCTs also occur in humans but are relatively rare (Carpino et al., 2007). The consensus
conclusions of a workshop on the MOA for LCT were reported by Clegg et al. (1997) and
classified the chemicals that caused LCT in animal studies into seven MOA categories. The
postulated MO As support the following hormonal steps to the process:
• A xenobiotic chemical inhibits the production of testosterone leading to low serum levels
• Low serum testosterone levels signal the hypothalamus to produce gonadotropin
releasing hormone (GnRH)
• GnRH signals the pituitary to release luteinizing hormone (LH)
• LH signals the Leydig cells to produce testosterone
• LH causes Leydig cell proliferation in testosterone production
The sustained increase in circulating LH and chronic stimulation of Leydig cells by
growth-stimulating mediators including IGF-1, TGF-a, leukotrienes and various free radicals
could lead to DNA replication errors and LCT development (Clegg et al., 1997). TGF-a, which
binds to the epidermal growth factor receptor and stimulated cell proliferation has been detected
in Leydig cells (Teerds et al., 1990).
A series of mechanistic studies have been conducted by Cook and colleagues (Cook et
al., 1992; Biegel et al., 1995; Liu et al., 1996) to investigate the mechanism of LCT formation in
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male Sprague-Dawley rats exposed to PFOA. Administration of PFOA to adult male rats by
gavage for 14 days was shown to decrease testosterone levels and increase serum estradiol levels
(Cook et al., 2002). These endocrine changes correlated with its potency to induce LCTs in rats
and were hypothesized to play a role in the PFOA-induction of LCTs (Biegel et al., 2001).
Subsequent experiments have shown that PFOA increased levels of estradiol by inducing
cytochrome P450 CYP19 (aromatase). Aromatase converts androgens to estrogens including the
conversion of testosterone to estradiol. PFOA has also been shown to directly inhibit
testosterone production when incubated with isolated Ley dig cells and ex vivo studies
demonstrated that this inhibition was reversible (Biegel et al., 1995). This inhibition of
testosterone biosynthesis appeared to be mediated by PPARa (Gazouli et al., 2002) and may
contribute to the development of LCTs through disruption of the hypothalamus-pituitary-
testicular axis. However, in the mechanistic bioassay by Biegel et al. (2001), serum testosterone
and LH levels were not significantly altered at the levels of PFOA that were tested.
Support for 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 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 shown to occur in wild-type and humanized
PPARa mice. These effects were not seen in PPARa-null mice.
The induction of LCTs by PFOA can be attributed to a hormonal mechanism whereby
PFOA either inhibits testosterone biosynthesis and/or lowers testosterone by increasing its
conversion to estradiol by upregulating aromatase activity in the liver. Both of these
mechanisms appear to be mediated by PPARa. However, the data are not currently sufficient to
demonstrate that the other key steps in the postulated MOA are present in PFOA-treated animals
following exposures that lead to tumor formation. Studies are needed to demonstrate the
increase of GnRH and LH in concert with the changes in aromatase and estradiol.
Pancreatic Acinar Cell Tumors (PACT). As with LCT, the MOA for PACTs is not
understood. These tumors are most commonly identified in rats but do occur in other animal
species (mice, hamsters) and in humans (Wisnoski et al, 2008). Males are more susceptible to
pancreatic tumors than females. Two hypothetical MO As have been proposed and are as follows
(Obourn et al., 1997; Klaunig et al., 2003):
• Events that stimulate increases of growth factors such as cholecystokinin (CCK)
and/or gastrin activate a feedback loop resulting in proliferation of the secretory
pancreatic acinar cells. A high fat diet, trypsin inhibition, cholestasis, or changes in
bile composition are proposed initiators of this sequence of events.
• Increased levels of testosterone support the growth of acinar cell preneoplastic foci
and carcinomas and estrogens inhibit the cellular proliferation.
There is minimal information on the relationship of PFOA exposure to either of the
proposed MOAs. Obourn et al. (1997) studied the impact of PFOA on CCK and trypsin using in
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vitro assays and found that PFOA was not an agonist for the CCKA receptor that activated CCK
release. PFOA also had no inhibitory action on trypsin at levels 1000 times greater (0.31 ug/mL)
than the positive control.
The Obourn et al. (1997) study also looked at Wyeth-14,643, a peroxisome proliferator, in
these same assays and found results similar to those for PFOA. When they conducted an in vivo
study with 100 ppmWyeth-14,643, they found a small but significant increase (p<0.05) in bile
flow per unit body weight, a decrease (p<0.05) in bile flow per unit liver weight, and a small
decrease (p<0.05) in the total bile acid concentration following a 6 month dietary exposure.
Plummer et al. (2007) reported on gene expression changes induced in pancreatic acinar cells
isolated from Sprague-Dawley rats fed diets containing 300 ppm (-20 mg/kg/day) PFOA for 28
days. Expression of genes regulated by PPARa, y, 5 in pancreatic acinar cells was directly
opposite of the expression of those same genes in liver tissue.
As indicated in the discussion of the LCTs above, PFOA appears to suppress testosterone
production through the induction of aromatase. Accordingly, the second proposed MOA for
PACTs is not applicable to PFOA. At the present time, there are insufficient data to demonstrate
a MOA that can account for the PACTs identified in the chronic study by Biegel et al. (2001).
Other Potential Modes of Action. There are other potential MO As of actions that may apply to
PFOA. They include interruption of intercellular communication, mitochondrial effects, and
hormonal effects. None of these mechanisms are considered to be key steps in the MO As
discussed above.
Gap junction intercellular communication (GJIC), a process by which cells exchange
ions, second messages, and other small molecules, is important for normal growth, development,
and differentiation, as well as maintenance of homeostasis in muticellular organisms. Because
tumor formation requires loss of homeostasis and many tumor promoters inhibit GJIC, it has
been hypothesized that GJIC may play a role in carcinogenesis (Trosko et al, 1998). PFOA has
been demonstrated to inhibit GJIC in liver cells in vitro and in vivo (Upham et al., 1998. 2009).
However, inhibition of GJIC is a widespread phenomenon, and the effect by PFOA was neither
species nor tissue specific. In addition it was reversible. Thus, the significance of GJIC inhibition
in regard to the mode of carcinogenic action of PFOA is unknown.
Several chemicals structurally related to PFOA have been shown to manifest their
toxicity by interfering with mitochondria biogenesis and bioenergetics. Berthiaume and Wallace
(2002) noted an increase in mitochondrial biogenesis in liver following treatment of rats with
PFOA. PFOA has also been demonstrated to uncouple oxidative phosphorylation in
mitochondria of the liver from rats exposed via their diet (Keller et al., 1992). At high
concentrations, PFOA caused a small increase in resting respiration rate and slightly decreases
the membrane potential. The observed effects are believed to be attributed to a slight increase in
nonselective permeability of the mitochondria membranes caused by PFOA's surface-active
properties (Starkov and Wallace, 2002).
Estrogen has been shown to promote hepatocarcinogenesis in rats (Yager and Yager,
1980; Cameron et al., 1982). Thus, the increase in estrogen levels after PFOA exposure may
play a role in hepatocarcinogenesis in rats as well as influencing LCTs. More research is needed
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to support the involvement of this MOA to the hepatocarcinogenesis of PFOA.
4.4.4 Weight of Evidence Evaluation for Carcinogenicity
The findings for cancer in humans are equivocal; some epidemiological studies show a
positive association with exposures and others are negative. The Dupont (2003b) report on the
West Virginia Washington Works Plant found statistically significant elevations in cancer of the
bladder, kidney, and urinary organs and a nonstatistically significant elevation in leukemia
incidence. Two earlier worker studies had statistically significant findings for leukemia but both
had methodological limitations (Walrath and Burke, 1989; Karns and Fayerweather, 1991). The
mortality data for liver cancer, pancreatic cancer, prostate cancer, and testicular cancer all
showed no association to exposure. In contrast, the highest serum PFOA levels were associated
with testicular, kidney, prostate, and ovarian cancers and non-Hodgkin's lymphoma among the
general population exposed through contaminated water (Vieira et al., 2013). These findings are
limited by small numbers of cases or a positive association with the very high exposure category,
but weaker or negative associations for the other exposure categories.
The only chronic bioassays of PFOA were conducted in rats (Butenhoff et al., 2012;
Biegel et al, 2001). The two studies support a positive finding for the ability of PFOA to be
tumorigenic in one or more organs of male but not female rats. There are no carcinogenicity data
from a second animal species. There are some data that provide support for the hypothesis that
the 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 may also play a role in the other tumor types observed, but more
data that will 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 fact that PFOA is not metabolized and would not be activated by the microsomal enzymes,
the clastogenic effects were likely the result of an indirect mechanism. Involvement of ROS in
the MOA as a result of PFOA alone is unlikely because of its metabolic stability. Conditions
leading to ROS would be a function of metabolic conditions perturbed by PFOA rather than
PFOA alone. A compound that is not metabolized will not be able to covalently introduce any of
its atoms into the structure of DNA.
Despite the limitations in the data for the Leydig cell and pancreatic tumors, a finding of
Suggested evidence of carcinogenicity is justified and quantification of the dose response is not
recommended. The tumor dose-response data are not indicative of a high potency, and the dose-
response information from the noncancer studies indicate that protecting for several of the
noncancer endpoints (e.g. increases in liver weights, hepatocyte cellular abnormalities,
immunotoxicity, developmental effects and delays) will result in an RfD that will be protective
for any tumorigenic effects.
4.4.5 Potentially Sensitive Populations
Human biomonitoring studies do not suggest major differences between serum PFOA
levels in males and females. However, the worker populations that would be those most likely to
demonstrate such differences because of their higher exposures were predominantly male.
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Some animal models have gender differences which affect toxicity of PFOA. Sexually
mature female rats excreted almost all of a 10 mg/kg dose of PFOA within 48 hours compared to
only 19% excreted by male rats. Male hamsters excrete PFOA faster than female hamsters, and
female rabbits excrete PFOA slightly faster than male rabbits. Male and female mice excrete
PFOA at approximately the same rate (Hundly et al., 2006). Studies of the transporters involved
in the toxicokentics of PFOA demonstrate that they are differentially impacted by the presence of
male and female sex hormones (Kudo et al., 2002; Cheng et al., 2006) influencing tissue
persistence. As studied in rats (Kudo et al., 2002), the male sex hormones increased half-life
(decreased excretion) of PFOA while the female hormones were associated with shorter half
lives (increased excretion). The gender differences in mice are not as pronounced as those in
rats. Work by Cheng et al. (2002) and Cheng and Klaassen et al. (2009) demonstrated that
hormones impact transporters in liver and kidney.
In studies in which male and female rats were used, the males were more sensitive to
toxicity than female rats (Goldenthal, 1978a; York, 2002; Butenhoff et al., 2004a; Sohlenius et
al., 1992). Monkeys and mice displayed similar sensitivities following PFOA exposure
(Goldenthal, 1978b; Christopher and Marias, 1977; Sohlenius et al, 1992). In the monkey
studies the number of animas/sex/dose group was too small reveal a difference related to gender.
Unfortunately much work remains to be done in order to determine whether the gender
difference seen in rats is relevant to humans. Similarities are possible because the long half-life
in humans suggests that they may be more like the male rat rather than the female rat. There is a
broad range of half-lives in human epidemiology studies suggesting a variability in human
transport capabilities resulting from genetic variations in structures and consequently in function.
Genetic variation in human OATs and OATps has been identified as described in a review by
Zair et al. (2008).
Neonates, Infants, and Fetuses.
An epidemiological study examining postnatal outcomes including developmental
milestones at 6 and 18 months of age did not find any differences in fine motor, gross motor, or
cognitive skills relative to maternal PFOA concentration (Fei et al., 2008b).
Several animal studies have examined potential modes of action for developmental
effects following prenatal exposure to PFOA. PFOA-exposure during development in rabbits,
rats, and mice resulted in increased resorptions (mouse) increased fetal skeletal variation (rabbits,
rats, mouse), decreased neonatal survival (rat, mouse), decreased postnatal body weight (mouse),
delayed eye opening and body hair growth (mouse), delayed vaginal opening (mouse),
accelerated preputial separation (mouse), delayed mammary gland development (mouse dam and
offspring) (York, 2002; Butenhoff et al., 2004a; Lau et al., 2006; Abbott et al., 2007; Wolf et al.,
2007; White et al., 2007, 2009, 2011; Macon et al., 2011). Other long term effects observed in
the surviving offspring included increased body weight gain, serum leptin, and serum insulin
levels along with changes in adipose tissue (Hines et al., 2009). The mechanisms of action for
these developmental effects are unknown, but several potential modes of action have been
investigated.
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Wolf et al. (2007) restricted mouse prenatal PFOA-exposures to 3-11 day periods during
gestation to determine if PFOA was affecting a certain stage of organogenesis that resulted in the
observed developmental effects. Decreased postnatal survival was observed at the highest dose
used (20 mg/kg/day). Eye opening and body hair growth were delayed in offspring exposed for
the longest periods of time (GD7-17 and GD10-17) and may have been the result of a higher
cumulative dose or greater sensitivity during early gestation. A cross-fostering paradigm was
used to determine if the developmental effects were the result of gestational exposure, lactational
exposure, or a combination of both. Postnatal survival was decreased in offspring exposed
through gestation and lactation (5 mg/kg/day). Eye opening and body hair growth were delayed
and body weight was reduced in offspring exposed during gestation (5 mg/kg/day), and gestation
and lactation (3 and 5 mg/kg/day). No developmental delays were observed in offspring exposed
via lactation only indicating that PFOA alters growth regulation in the developing fetus which
persist as growth continues postnatally.
Abbott et al. (2007) examined activation of PPARa as a factor in the developmental
toxicity of PFOA. Wild-type and PPARa null mice experienced full litter resorptions following
gestational (GD1-17) PFOA exposure (>5 mg/kg/day) indicating the mechanism of PFOA-
induced resorptions was independent of PPARa expression. These resorptions may be due to
insufficient trophoblast cell type differentiation and/or increased trophoblast cell necrosis (Suh et
al., 2011). Postnatal survival was significantly decreased in wild-type offspring but not in
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 may also play a role in its mechanism. Takacs and Abbott (2007) showed
that PFOA can activate mouse PPARp/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 PFOA GDI-17 and
GD8-17 had significantly delayed mammary gland development (full of alveoli, visible adipose
tissue, not well differentiated) at PND10 which is the peak of lactation in rodents. The delayed
dam mammary gland development may play a role in the observed reduced offspring body
weight gain if the quantity or quality of the milk is altered by PFOA (Lau et al., 2006; Abbott et
al., 2007; Wolf et al., 2007; White et al., 2007). Restricted gestational exposure and cross-
fostering studies showed that delayed offspring mammary gland development, observed PND1-
63, occurred regardless of exposure duration or timing (gestation vs. lactation exposure). The
developmental delays persisted even as the internal PFOA dose decreased (White et al., 2007,
2009, 2011; Macon et al., 2011). More studies need to be conducted to elucidate the MOA for
dam and offspring mammary gland effects.
Mammary gland development was also affect by peripubertal exposure to PFOA (Yang et
al., 2009, Zhao et al., 2010). Low doses (5 mg/kg/day) of PFOA from 3 to 7 weeks of age
caused accelerated mammary gland development in C57BL/6 mice but delayed mammary gland
development in BALB/c mice. Experiments examining the mechanism for accelerated
mammary gland development showed that PFOA promotes steroid hormone production in the
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ovaries and increases the levels of several mammary gland growth factors in C57BL/6 wild-type
and 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, may place PFOA
into the environmental endocrine disrupter obesogen category similar to diethylstilbestrol (DES)
(Newbold et al., 2007). In humans, increased leptin levels are associated with increased body fat
and suggestive of a leptin-resistance mechanism of action for overweight (Considine et al.,
1996). A similar relationship may occur in prenatally PFOA-exposed mice. Studies determining
MO As should be conducted to determine relevance to human health.
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5.0 DOSE-RESPONSE ASSESSMENT
5.1 Dose-Response for Noncancer Effects
A Reference Dose (RfD) or Reference Concentration (RfC) is used as a benchmark for
the prevention of long-term toxic effects other than carcinogen!city. RfD/RfC determination
assumes that thresholds exist for toxic effects, such as cellular necrosis, significant body or organ
weight changes, blood disorders, etc. The RfD is expressed in terms of milligrams per kilogram
per day (mg/kg-day) and the RfC is expressed in milligrams per cubic meter (mg/m3). The RfD
and RfC are estimates (with uncertainties spanning perhaps an order of magnitude) of the daily
exposure to the human population (including sensitive subgroups) that is likely to be without an
appreciable risk of deleterious effects during a lifetime.
5.1.1 RfD Determination
The risk assessment for PFOA is fraught with a number of challenges including:
• Its toxicokinetic complexity
• The variability in half-lives between and within species
• The wealth of inconsistent data from epidemiological studies
• The robust but diverse database of animal studies
• Inconsistencies between some of the associations from the epidemiology studies
and the effects from the animal studies
• The metabolic inertness of PFOA in living systems
• The number of animal studies that lack a NOAEL.
A subset of the challenges has an impact on dose-response, especially the toxicokinetic
features that lead to differences in half-lives across species and in the case of rats, genders.
Toxicokinetics also influence intraindividual and lifestage variability. Each of these
toxicokinetic features highlights the importance of internal dose, and supports the utilization of a
pharmakolinetic model as a component of the dose-response assessment.
Human Data. Studies have examined occupational and residential populations at or near 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. Endpoints monitored included standard clinical chemistry
parameters, measures of cardiovascular risk, signs of organ damage, standard haematological
endpoints, and diabetic or prediabetic conditions. Reporductive and developmental parameters
were also evaluated in some epidemiology studies.
In reviewing the epidemiological monitoring data for PFOA it is important to remember
that many individuals were coexposed to other fluorocarbons, and, although only serum PFOA
measurements are often given, many individuals had other PFCs present in their serum. In no
study was the actual exposure concentration or dose to the individuals known; thus, a
quantitative exposure-response assessment based on the human data is not feasible.
In many C8 project studies, drinking water was identified as the primary contributor to
exposure. However, indoor air and dietary exposures were possible and quantitative exposure
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data are lacking for all routes. Similarly, occupational epidemiology studies cannot quantify
exposures occurring away from the workplace. In all of the epidemiology studies the wide
ranges of serum levels are likely a reflection, not only of intrahuman toxicokinetic variability,
but also of diversity in external exposure sources. This is especially evident in the units used to
express PFOA serum levels in occupational populations, |ig/mL, versus in the general
population, ng/mL. From the studies discussed in Section 4.1.1.1, mean serum levels in the
occupational cohorts ranged from 1.02 - 4.63 |ig/mL while those for the general population
ranged from 1.05 - 354 ng/mL. The highest level seen in the general population is from
residents serviced by a contaminated water district. For both the positive and negative findings
summarized in the paragraphs below, mean or median serum PFOA levels fall within these
ranges for worker and general populations.
The epidemiological studies have found a positive association between serum PFOA
concentration and total cholesterol in the general population (Steenland et al., 2009; Frisbee et
al., 2010; Nelson et al., 2010) and in worker populations (Olsen et al., 2001a,b, 2003; Sakr et al.,
2007a,b; Costa et al., 2009) but no consistent trends in LDL, VLDL, or HDL levels. A positive
association has been shown between serum PFOA concentration and increased liver enzymes
and/or decreased bilirubin in both worker and general populations (Sakr et al., 2007,a,b; Olsen
and Zobel, 2007; Costa et al., 2009; Lin et al., 2010; Gallo et al., 2010), chronic kidney disease
in the general population (Shankar et al., 2011), and the odds of experiencing early menopause
(Knox et al., 2011). Maternal or child plasma levels of PFOA were positively associated with
decreased antibody liters in children after vaccination (Grandjean et al., 2012), obesogenic
effects in female children at 20 years of age (Halldorsson et al., 2012), delayed puberty in girls
(Lopez-Espinosa et al., 2011), and parent reported ADFID (Hoffman et al., 2010).
No association between serum PFOA and risk of dying from ischemic heart disease has
been found in several worker populations (Sakr et al., 2009; Leonard et al., 2008; Lundin et al.,
2009). No consistent associations have been found between hyperglycemia, type II diabetes
(MacNeil et al., 2009; Lin et al., 2009; Nelson et al., 2010), thyroid homeostasis (Olsen et al.,
2001a,2003; Sakr et al., 2007a; Olsen and Zobel, 2007; Costa et al., 2009; Emmett et al., 2006;
Bloom et al., 2010), fertility, fecundity, and birth outcome (Fei et al., 2009; Monroy et al., 2008;
Hamm et al., 2009; Washino et al., 2009; Nolan et al., 2010; Stein et al., 2009). Maternal or
child plasma PFOA levels were not associated with developmental milestones in children (Fei et
al., 2008b), child risk of hospitalization for infectious diseases (Fei et al., 2010), or attainment of
puberty (Christensen et al., 2011).
Animal Data - Long Term Studies. Effects in animal studies are most commonly those
associated with activation of the PPARa receptor leading to peroxisome proliferation. These
include increased liver weight, decreases in triglycerides, cholesterol, and lipoprotein, plus
increases in ALT and/or AST. However, the mechanisms for other effects, such as decreased
body weight, immunological effects, and developmental delays, are unknown, yet may be
relevant to human health risk assessment.
As an initial step in the dose-response assessment, U.S. EPA identified a suite of animal
studies with low dose NOAELs and/or LOAELs as potential candidates for development of a
chronic Reference Dose (RfD). These studies are listed in Table 5-1. The candidate studies
were selected based on their low dose NOAEL and/or LOAEL, a duration of >7 weeks, use of a
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control, and two or more doses. Table 5-1 does not include the data from human epidemiology
studies because, although they include information on serum levels, they do not identify
exposure sources or external doses.
TABLE 5-1. NOAEL/LOAEL data for Oral Subchronic and Chronic Studies of PFOA
Species
Monkey
Male
Monkey
Female
Monkey
Male
Rat
Male
Rat
Male
Rat
Male
Rat Male
FO
generation
Rat Female
FO
generation
Rat Male
Fl
generation
Rat
Female
Rat
Male and
Female
Rat Female
Fl
Generation
Study
Duration
90 days
90 days
26 weeks
90 days
13 weeks
2 years
84 days
127
10 weeks
90 days
2 years
10 weeks
NOAEL
mg/kg/day
none
o
J
none
0.56
0.06
1.3
none
30
none
22.4
1.3 (m)
1.6 (f)
10
LOAEL
mg/kg/day
3
10
3
1.72
0.64
14.2
1
none
1
76.5
14.2 (m)
16.1 (f)
30
Critical Effects (s)
Increased relative pituitary weight
Decreased absolute and relative heart
weight
Increased absolute liver weight
(hepatocellular hypertrophy) and mean
liver-to-body weight percentages
Hepatocyte necrosis, increased absolute
liver weight
Increased absolute and relative liver weight
with hepatocellular hypertrophy
Decreased body weight gain, increased
kidney weight, and hepatic cystoid
degeneration, megalocytosis, decreased red
blood cell count, hemoglobin
concentration, and hematocrit values,
testicular vascular mineralization,
pulmonary hemorrhage
Increased absolute liver and kidney weight
Decreased body weights and weight gains
increased absolute and relative liver
weights
Increased absolute and relative liver
weight
M: Decreased body weight gain; histopath.
lesions in liver, testes, and lungs.
F: Decreased body weight gain and
decreased red blood cell count,
hemoglobin, concentration, and hematocrit
values
Female: Delay in sexual maturity,
decreased body weight and weight gain
Reference
Goldenthal,
1978b
Goldenthal,
1978b
Thomford,
2001b,
Butenhoff etal.,
2002
Goldenthal,
1978a
Palazzolo, 1993;
Perkins et al.,
2004
Butenhoff etal.,
2012
York, 2002;
Butenhoff etal.,
2004a; York et
al., 2010
York, 2002;
Butenhoff etal.,
2004a; York et
al., 2010
York, 2002;
Butenhoff etal.,
2004a; York et
al., 2010
Goldenthal,
1978a
Butenhoff etal.,
2012
York, 2002;
Butenhoff etal.,
2004a; York et
al., 2010
When examining the effects associated with the low dose exposures summarized in Table
5-1, changes in relative and/or absolute liver weight appear to be the common denominator for
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monkeys and rats (York, 2002; Butenhoff et al., 2004a; Goldenthal, 1978a; Palazzolo, 1993;
Perkin et al., 2004; Thomford, 2001b, Butenhoff et al., 2002) with or without other hepatic
indicators of adversity (Palazzolo, 1993; Perkin et al., 2004; Thomford, 2001b, Butenhoff et al.,
2002). Serum PFOA levels, where available, associated with increased liver weight were 81 and
41 |ig/mL for the male monkey and rat, respectively. Body weight effects were seen in several
studies (York, 2002; Butenhoff et al., 2004a; Butenhoff et al., 2012). Testicular effects were
observed by Butenhoff et al. (2012) and in the chronic one-dose study by Beigel et al. (2001).
There were developmental delays for males and females in the two-generation study published
by York (2002) and Butenhoff et al. (2004a).
Four of the studies in Table 5-1 lack a NOAEL and have LOAELs that range from 1 to 3
mg/kg/day. The NOAELs for the remaining 7 studies range from 0.06 to 22.4 mg/kg/day. Male
monkeys and rats appear to respond at doses that are lower than their female counterparts. No
long-term studies in mice were identified. Since NOAELs and LOAELs are to some extent the
product of concentration or dose level selection, examination of the dose information in Table 5-
1 suggests that several data sets have the potential to be co-critical in the dose-response
evaluation.
Animal Data -Short Term Studies. A number of studies in mice have also identified adverse
effects following low dose exposures. The studies fall into two clusters, those evaluating
developmental or reproductive effects and those with a focus on immunological effects. The
critical shorter-term studies in rats and mice are summarized in Table 5-2.
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TABLE 5-2. Shorter term and Developmental Oral Exposure Studies
Species
Study
Duration
NOAEL
(mg/kg/day)
LOAEL
(mg/kg/day)
Critical Effect(s)
Reference
Rat
Rat
Male
29 days
1
10
Increased absolute and relative
liver weight, focal liver necrosis
Loveless et
al., 2008
Mouse
Mouse
Male
Mouse
Male
Mouse
Female
Mouse
Female
Mouse
Female
Mouse
Female
Mouse
Female
29 days
29 days
15 days
15 days
17 (pups)
718 (dams)
days
17 days
GD1-17
1 1 days
GD7-17
0.3
1
none
1.88
none
none
none
1
0.3
0.94
3.75
1
3
5
Increased absolute and relative
liver weight, decreased relative
spleen weight, moderate-severe
liver hypertrophy with single
cell and focal necrosis
Decreased spleen weight
Increased absolute and relative
liver weight (1 and 15 days post
dose)
Decreased IgM (1 days post
dose), increased IgG (15 days
post dose), decreased absolute
and relative spleen weight (1 day
post dose)
Increased absolute maternal liver
weight, reduced ossification
(calvarin, enlarged fontanel),
accelerated onset of puberty in
male offspring
Increased absolute and relative
maternal liver weight, delayed
offspring eye opening and body
hair growth, increased offspring
relative liver weight, decreased
offspring body weight, delayed
mammary gland development
(female offspring)
Increased maternal and pup
relative liver weight, delayed
offspring eye opening and hair
growth, decreased male
offspring body weight, delayed
mammary gland development
(female offspring)
Loveless et
al., 2008
Loveless et
al., 2008
DeWittetal.,
2008
DeWittetal.,
2008
Lau et al.,
2006
Wolfetal.,
2007; White
etal.,2009
Wolfetal.,
2007; White
etal.,2009
As was the case with the longer-term studies, increased liver weight was a common effect
among the shorter-term studies summarized in Table 5-2. Increases in absolute or relative liver
weights were reported in six of the eight studies (Lau et al., 2006; Wolf et al., 2007; DeWitt et
al., 2008; Loveless et al, 2008; Table 5-2). The co-occurring effects at the LOAEL were effects
on spleen, thymus, liver, and/or developmental endpoints. Four of the eight studies lacked a
NOAEL. For those studies with a NOAEL, the value ranged from 0.3 to 1.88 mg/kg/day while
the LOAELs ranged from 0.3 to 10 mg/kg/day. Serum PFOA levels associated with effects in
female mice ranged from 21.9 - 36.3 jig/mL. In both cases, the range of values across studies is
narrow with overlap between the NOAELs and LOAELs. In all instances the durations of
exposure in shorter term studies were less than 29 days, suggesting that physiological responses
Perfluorooctanoic Acid - February 2014
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5-5
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to PFOA occur early in the exposure continuum and at doses comparable to those observed in the
long term studies.
5.1.1.1 Benchmark Dose Approach
As a second step in the dose-response analysis, benchmark modeling of dose-
response for the liver weight changes from the studies wherein this was a critical effect was
conducted. The data were modeled using benchmark dose software (BMDS) v. 2.1.2 for a
continuous dataset with modeled variance and restricted slope where applicable. A 10% increase
in absolute liver weight was chosen as the benchmark response (BMR) for the initial analysis.
Although it is a biomarker for systemic exposure in rodents rather than a biomarker of adversity,
it serves as a lowest common denominator for loss of homeostasis in a sensitive species.
PFOA is not metabolically altered during absorption, distribution and excretion.
Systemic homeostasis is lost when renal excretion is no longer balanced with absorption. There
is clear evidence that PFOA has a high affinity for the proteins that compose the PPARa receptor
of rodents resulting in hypertrophy and increased liver weight as an early signature of exposure.
Thus, both liver responses serve as biomarkers for the loss of homeostasis and can be
extrapolated to species other than PPARa sensitive-rodents. The studies of Nakamura et al.
(2009) and Albrecht et al. (2013) of mice with the humanized PPAR-a receptor did not observe
significant changes in liver weight. However, there were changes in humanized mice that
occurred at doses that exhibited the liver weight effects in wild type mice and support its use as a
biomarker for loss of homeostasis. Specifically, humanized mice exhibited changes in gene
products that modulate lipid metabolism (Albrecht et al., 2013; Nanamura et al., 2009) at doses
comparable to those associated with liver weight effects in wild-type mice.
Results of the benchmark analyses are shown in Table 5-3 where goodness-of-fit
information [p value and Akaike's Information Criterion (AIC)] was used to choose the best
model for each data set. Liver weight data were available in the male rat (Butenhoff et al.,
2004a; Palazzolo, 1993; Perkins et al., 2004; York, 2002; and York et al., 2010), male mouse
(Loveless et al., 2008), and female mouse (Lau et al., 2006). For each data set, the BMD for a
10% increase in liver weight and the lower-bound confidence limit on the BMD (BMDL) are
provided.
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TABLE 5-3. Benchmark Dose Modeling for a 10% Increase in Liver Weight
Species
Best fit
Model
BMD10
mg/kg/day
BMDL10
mg/kg/day
Notes
Reference
Rat
Rat
Male
Rat Male
Rat Male
FO
generation
Rat Male
Fl
generation
Exponential:
Model 4
Exponential
Model 4
Exponential:
Model 4
Exponential:
Model 4
0.798012
0.241376
0.399087
0.599409
0.456224
0.151833
0.274007
0.3655
Increased absolute and
relative liver weight
Increased absolute and
relative liver weight
Increased absolute and
relative liver weights
Significant increases in
absolute and relative liver
weights
Palazzolo, 1993;
Perkins et al.,
2004
Loveless et al.,
2008
York, 2002;
Butenhoff etal.,
2004a; York et
al., 2010
York, 2002;
Butenhoff etal.,
2004a; York et
al., 2010
Mouse
Mouse
Male
Mouse
Female
Exponential:
Model 4
Exponential:
Model 4
0.099268
0.292244
0.0680743
0.246874
Increased absolute and
relative liver weight
Increased maternal absolute
liver weight
Loveless et al.,
2008
Lau et al., 2006
Data from Thomford (2001b) for monkeys and Goldenthal (1978) for the male rat were
not presented because the low number of animals (n= 2 to 6) in each dose group resulted in
highly variable lower bound estimates. An attempt was also made to use the BMDS to model
liver-to-body weight ratio data from the studies where this information was available. However,
all models failed with p<0.1 indicating that the fit was not reliable
Taken together, the analyses presented in Table 5-3 identify BMD values that vary by
less than an order of magnitude (about 0.1 to 0.8 mg/kg/day) for a 10% increase in liver weight
despite differences in species. Likewise, the values for the calculated lower-bound estimates did
not vary greatly (about 0.1 to 0.5 mg/kg/day) indicating similarity in response between species.
It is interesting to compare the empirical results presented in Table 5-1 with the BMD model
results in Table 5-3. For the six data sets modeled, a NOAEL was not identified in four and was
0.06, 0.3, and 0.56 mg/kg/day in the remaining three studies. The LOAEL was 1 mg/kg/day in
four studies and 0.64, 1.72, and 3.0 mg/kg/day in the others. The BMDLio values all fall below
the experimental LOAELS.
The model results shown in Table 5-3 were chosen based on goodness-of-fit criteria as
described above. The modeling output for all studies is included in Appendix 1. A subset of the
modeled data is used in the RfD derivation (section 3.1.1.3). Figure 5.1 provides the graphic
results from those datasets which were selected as critical for the RfD derivation. Data from the
FI generation male rats (York, 2002; Butenhoff et al., 2004a; York et al., 2010) were not used
because of uncertainties about the timing for the developmental changes in the renal tubular
transporters that differentiate male rats from female rats (see Section 3.4). The developmental
toxicity study in female mice (Lau et al., 2006) supports the endpoint of liver weight effects and
is included in Figure 5-1 despite the limitation posed by its 17-day, relatively short exposure
duration.
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5-7
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Exponential Model 4 with 0.95 Confidence Level
Exponential -
BMDL BMD
07:3708/182011
Male Rat (Palazzolo, 1993; Perkins et al., 2004)
Exponential Model 4 with 0.95 Confidence Level
Exponential
15
dose
20 25
13:1212/222011
Male Rat (Loveless et al., 2008)
Exponential Model 4 with 0.95 Confidence Level
Exponential Model 4 with 0.95 Confidence Level
8 26
10:57 10/262011
0 5 10 15 20 25 30
dose
F0 Male Rat (Butenhoff et al., 2004a; York, 2002; York et
al., 2010)
10:4808/242011
15
dose
20 25
Male Mouse (Loveless et al., 2008)
Exponential Model 4 with 0.95 Confidence Level
5.5
5
3.5
3
2.5
2
Exponential
10
dose
15:1210/252011
Female mouse (Lau et al., 2006)
FIGURE 5-1. BMDS graphic output from selected model runs.
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5.1.1.2 Pharmacokinetic Model approach
In linking chemical exposure to toxic endpoints, careful consideration of the
pharmacokinetics is crucial. This is especially true for PFOA, where inter-species and gender
variation in clearance half-life can vary by several orders of magnitude. If the toxicological
endpoints are assumed to be driven by internal concentrations, then it is the internal exposure
that must be calculated and considered across species. Differences in pharmacokinetics (e.g.,
male rats excrete PFOA less quickly than females) and differences across species produce
differences in the external dose needed to achieve the same internal dose.
PFOA has been found to have dose-dependent kinetics. Although repeated doses rapidly
result in quasi-equilibrium blood concentrations, a single dose results in a much longer half-life
than would be consistent with a rapid approach to quasi-equilibrium (Andersen et al. 2006, Lou
et al. 2009). Using a simple, linear pharmacokinetic model (e.g. a one- or two-compartment
model) to predict internal dose resulted in estimated exposures due to repeated doses that were
greater than actually seem to occur (Butenhoff et al. 2004, Lou et al. 2009).
The saturable renal resorption model (Andersen et al. 2006: Section 3.5.1) has been used
to describe how PFOA can reach steady state faster than the elimination half-life would indicate.
The mechanism for this phenomenon is suggested to be the result of saturable resorption from
the kidney filtrate: Unbound compound in plasma is filtered at the kidney glomerulus and passes
to the proximal tubules where it is reabsorbed by transporters and re-enters systemic circulation.
If the amount of compound in the filtrate exceeds the ability of the transporters to reabsorb it, it
is rapidly excreted so that additional doses do not affect the steady state serum levels.
In the Lou et al. (2009) report, the one and two compartment models were able to
reasonably predict the serum levels for single 1 and 10 mg/kg doses, but not at the high dose of
60 mg/kg/day where excretion exceeded transporter resorption capability and serum values
declined more rapidly than predicted based on the responses at 1 and 10 mg/kg/day (Section
3.5.1). Elimination from the body was much faster at the high dose than it was for the lower
doses. This behavior is consistent with the saturable resorption model: at low doses excretion
becomes balanced with uptake. Thus, plasma levels remain at steady state as long there is no
increase in transporter density in the membrane of the renal tubule and dosing is constant. When
a dose was large enough to exceed resorption capacity, the excretion rate increases but simple
one- and two-compartment models do not allow for excretion rate changes. In Lou et al. (2009)
the addition of a saturable resorption component to the model greatly improved the fit for the 60
mg/kg single dose and 20 mg/kg/day repeat doses, while still describing the 1 and 10 mg/kg
doses.
The biological basis for the saturable resorption model parameters is still uncertain (Lou
et al., 2009), particularly with respect to the idea that glomerular filtration alone is responsible
for the movement of PFOA into the renal filtrate given that PFOA is highly protein bound
(Kerstner-Wood et al., 2004) and export transporters are located along the apical and basal
membranes of the renale tubular cells. Protein binding kinetics will impact the levels available
for loss through glomerular filtration because the amount of free PFOA in serum will vary with
the saturation of serum protein-binding sites. Some loss of PFOA to the renal filtrate will also
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occur by way of tubular export transporters. However, the qualitative behavior of the saturable
resorption model is excellent (Andersen et al., 2006; Lou et al., 2009).
The saturable resorption model is an empirical model similar to the commonly used one-
and two-compartment models. Due to the empirical nature of the saturable resorption
mechanism present in the Tan et al. (2008) and Loccisano et al. (2011, 2012a,b) models, no
existing model for PFOA PK allows extrapolation between species. Resorption parameters
cannot be directly linked to the properties of the relevant serum transport proteins or membrane
transporters. Therefore the saturable resorption model parameters must be estimated using
species-specific pharmacokinetic data.
Pharmacokinetic data (serial blood concentrations following treatment with known
quantities of PFOA) were collected for three species: cynomolgus monkey (Butenhoff et al.,
2004), Sprague Dawley rat (Kemper et al., 2003), and mouse. Data were available for two
strains of mouse and were analyzed separately: CD1 (Lou et al., 2009) and C57BL6/N (Dewitt et
al., unpublished). Due to the pronounced difference in the pharmacokinetics of male and female
rats, the two genders were fit separately. For mice, only female data were used. For monkeys a
limited amount of female data was used jointly with male data, assuming the only difference
between the genders for monkey was bodyweight.
The available pharmacokinetic studies were not designed with the expectation of non-
linear pharmacokinetics. Accordingly, many parameters associated with saturable resorption
may be uncertain. For example, the Kemper et al. (2003) studies (Section 3.3) focused upon
single doses, although the relatively large highest dose (25 mg/kg) may be sufficient to cause
non-linearities such as that observed for a 60 mg/kg dose in CD1 mice (Lou et al., 2009)
described above. In the absence of ideal data, the question of parameter values becomes one of
what range of values is consistent with the data, rather than which single value is most
consistent. For the monkey, some of the data came from an intravenous administration study in
males and females (Butanoff et al., 2004b). Considering a range of possible values that might
equally well explain the data can be addressed with Bayesian statistical analysis. In this case the
Bayesian analysis was used to determine parameter distributions for all five data sets considered.
A non-hierarchical model for parameter values has been assumed; i.e. there is a single
value shared by all individuals of the same species/strain and gender (except for rat, where there
is a clear gender difference in renal excretion of PFOA). Body-weight and treatment (number
and magnitude of doses) are the only parameters that may vary between individuals. However,
with the exception of the Kemper et al. (2003) data, individual bodyweights were not available
so that a default bodyweight was used for each species and gender. In model development it was
assumed that animals were sacrificed at the end of the external exposure; i.e. there was no
continued internal exposure after dosing had stopped.
Bayesian analysis allows formal inclusion of prior knowledge in the form of set
distributions on the parameters being estimated (Gelman et al., 2004). Given that empirical
pharmacokinetic parameters can have a wide range of values, vague, bounded prior distributions
are appropriate (Wambaugh et al., 2008). For all estimated parameters the prior knowledge
values were assumed to be distributed log-normally. This constrained the parameters to positive
values. The mean and variances assumed for the CD1 mice, male and female Sprague-Dawley
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rats, and Cynomolgus monkeys are given in Table 5-4. For the C57BL/6N mice the available
data were insufficient to achieve a converged statistical analysis, so the posterior parameter
distributions for the CD1 mouse were used, with the variances increased ten-fold.
TABLE 5-4. Description of prior distributions used to analyze all but the C57BL/6N mice
Parameter
ka
vcc
ki2
Rv2:Vl
T
^maxc
kT
free
Pec
v^
Mean
1
1
1
1
1
1
0.01
1
1
Variance
1000
1000
1000
0.5
1000
1000
0.5
1000
1000
Bounds
Lower
1/6
10-io
10-io
10-io
10-io
10-io
10-io
0.01
10-io
Upper
105
105
105
100
105
105
1
105
10
Parameters were log-normally distributed with the mean and variance listed. Bounds were used to reduce the time spent sampling
in areas thought to have low probability, and were expanded when large posterior mass at the bounds indicated that the bounds
were too narrow. For the C57BL/6N the posterior parameter distributions for the GDI mouse was used, with the variances
increased ten-fold.
The deep tissue compartment of the Andersen etal. (2006) model is characterized in
terms of the rates to and from that compartment (ki2 and k2i, respectively). This corresponds to a
volume of distribution V2 = ki2*Vi/k2i, so that the ratio of the volume of the second
compartment to the first is equal to the ratio of ki2 to k2i (Rv2:vi = ki2/k2i). In order to enforce the
assumption that the primary (serum) compartment contains a significant portion of the PFOA, it
was assumed that the volume of the deep tissue compartment was constrained to be no more than
100 times greater than the volume of distribution of the serum. For this reason the ratio of the
two volumes is estimated, rather than the rate from the second compartment to the first
compartment. The rate of flow from the deep tissue back to the serum was calculated as k2i =
kl2/Rv2:vl-
Bayesian analysis was performed using Markov Chain Monte Carlo (Gelman et al. 2004).
The crux of a Bayesian analysis is the assessment of whether the distributions of values in the
Markov Chain reflect the posterior distribution of the statistical model given the priors and the
data. The distribution of parameter values were considered to "burned in" (i.e., true draws from
the posterior combination of the prior assumptions and the available data which were
independent of the assumed starting values) when they passed the Heidelberger and Welch
Stationarity test (Heidelberger and Welch 1983) as implemented in the Coda Package (Best et al.
1995).
For all pharmacokinetic parameter values (Table 5-5) a mean and a 95% interval are
reported, summarizing the distribution of parameter values found to be consistent with the
available data. Parameters such as the flow to and volume of the filtrate compartment were found
to be very uncertain as in Anderson et al. (2006) and Lou et al (2009). Between male and female
rats parameters were largely similar; including the affinity of the putative resorption transporters
but the maximum resorption transport rate was nearly 200 times greater for the males. The
median fraction of blood flow to the filtrate (Qmc) was physiologically reasonable (less than or
equal to the fraction of blood flow to the kidney) in all cases. In contrast the Lou et al. (2009)
model allowed an unreasonably high portion of cardiac output to pass through the kidney in
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order to optimize the fit to the data. Although not totally based on biology, the form of the
saturable resorption model seems to be an appropriate framework for the projection of internal
PFOA serum dose.
TABLE 5-5. Estimated and assumed pharmacokinetic parameters used in model development
Parameter
Body
Weight3
Cardiac
outputb
ka
vcc
kiz
Rv2:vl
T
Amaxc
kT
Free
Pec
vfllc
Units
kg
L/h/kg07
4
1/h
L/kg
1/h
Unitless
M/h
M
Unitless
L/h
L/kg
CD1 Mouse
Female
(Lau et al.,
2009)
0.025
8.68
290 (0.6 -
73000)
0.18(0.16-
2.0)
0.021
(S.lxlO"10-
3.8xl04)
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)
9.7xlO'4
(3.34xlO'9-
7.21)
C56BL/6N
Mouse
Female
(DeWitt et al.
unpublished)
0.025
8.68
340 (0.53 -
69000)
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(0.010-
0.081)
7.6xlO'5
(2.7X10'10-
6.4)
Sprague
Dawley Rat
Female
(Kemper et
al., 2003)
0.20(0.16-
0.23)a
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)
2.6xlO'5
(2.9X10'10-
28)
Sprague
Dawley Rat
Male
(Kemper et
al., 2003)
0.24 (0.21 -
0.28)a
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-
50000)
0.092
(3.4xlO'4-
1.6)
0.08 (0.03 -
0.22)
0.22(0.011
-58)
0.0082
(l.SxlO'8-
7.6)
Cynomolgus
Monkey male
and female
(Butenhoff et
al., 2004)
7 (m) 4.5 (f)
19.8
230 (0.27 -
73000)
0.4 (0.29 -
0.55)
0.0011
(2.4xlO"10-
3.5xl04)
0.98 (0.25 -
3.8)
3.9 (0.65 -
9700)
0.043 (4.3x10-
5 -0.29)
0.01(0.0026-
0.038)
0.15(0.02-
24)
0.0021
(3.3xlO'9-
6.9)
Estimated and assumed pharmacokinetic parameters for the Andersen et al. (2006) model. Means and 95% interval from Baysian
analysis are reported. For some parameters the distributions are quite wide, indicating uncertainty in that parameter (i.e., the
predictions match the data equally well for a wide range of values). As part of the Bayesian formalism, it is the posterior
distribution of parameter values, as opposed to the mean values, that is used to make the predictions.
a Estimated average bodyweight for species used except with Kemper study where individual rat weights were available and
assumed to be constant.
b Cardiac outputs obtained from Davies and Morris (1993)
If there is large uncertainty for parameters that greatly influence the predicted value, then
the distribution of predicted values will be wide. If the quantity being predicted is not sensitive
to the uncertain parameter, then the distribution of the predicted value will be unaffected
(Wambaugh et al. 2008). Assessing parameter uncertainty allows appropriate confidence in
model predictions.
For each study with a toxicological endpoint and LOAEL, the AUC and final serum
concentrations were determined for the exposure duration investigated in that study. These
values are summarized in 8-6 for rats, 8-7 for mice, and 8-8 for monkeys. In order to make a
rough assessment of the validity of the model predictions, a final serum concentration was
predicted for each treatment so that it could be compared to measured serum values. The
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predicted final serum concentration is the estimate for serum concentration at the time of
sacrifice. Generally these values were similar. They differed by a factor of four when strains
were different and closer to a factor of two when predicting using parameters from the same
strain. Because these predictions do not perfectly match the measured serum concentrations,
there remains uncertainty about the exposure estimates that has not been fully characterized.
TABLE 5-6. Predicted Final Serum Concentration and Time-integrated Serum Concentration (AUC) for
Studies in Rats
Study
Palazzolo,
1993;
Perkins et
al., 2004
York,
2002;
Butenhoff
et al.,
2004a;
York et
al.,2010
Species/
Strain
Rat (M)
ChR-CD
Rat (M)
Sprague-
Dawley
Exposure
Duration
13 weeks
Diet
FOM: 10
wkpre
mating-
mating
Oral
gavage
Oral Doses
mg/kg/day
0.06
0.64
1.94
6.50
1
3
10
30
Measured
Final Serum
value
ug/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
ug/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
7310(188)
69900 (1640)
170000 (6770)
331000 (27100)
93500 (2460)
207000(11000)
351000(35100)
422000 (74600)
Numbers in parentheses indicate standard deviation
M= male; s= serum; NT=not tested
Since the Kemper (2003) data were not tied to toxicological endpoints and were only used in model development they are not
included in this table.
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TABLE 5-7. Predicted Final Serum Concentration and Time-integrated Serum Concentration (AUC) for
Studies in Mice
Study
Wolfet
al, 2007;
White et
al., 2009
DeWitt et
al., 2008
Lauet
al., 2006
Species/
Strain
Mouse
(F)
CD-I
Mouse
(F)
C57BL/6
N
Mouse
(F)
CD-I
Exposure
Duration
GD1-173
Oral
gavage
GD7-17
Oral
gavage
15 days
Drinking
water
GD1-17
Oral
gavage
Oral
Doses
mg/kg/d
3
5
5
0.94
1.88
3.75
7.5
15
30
1
3
5
10
20
40
Measure
d Serum
Value
Hg/ml
NT
NT
24.8
NTb
NTb
35.3
42.8
50.0
162.6
21.9C
40.5 c
71.9C
116C
181C
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
Jig/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
31700(1740)
38600 (2050)
23300(1160)
7320 (557)
13800 (952)
22400 (1330)
30500 (1520)
40100 (4470)
55900 (12000)
16500 (1010)
33600 (2690)
40900 (2950)
51100(2810)
65400 (6480)
90400 (17500)
Numbers in parentheses indicate standard deviation
NA = not applicable; could not be determined
F=female; GD = gestation day; s=serum; NT=not tested
"Sacrificed on PND 22.
bDeWitt (2008) had 0.94 and 1.88 mg/kg/day dose groups in a second experiment.
The Lau et al. (2006) serum data were provided by the author for animals treated GDs 1-17.
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TABLE 5-8. Predicted Final Serum Concentration and Time-integrated Serum (AUC) in Studies of Monkeys
Study
Thomford,
200 Ib,
Butenhoff
etal.,
2002,
2004b
Species/
Strain
Monkey
(M)
Cyano-
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
Hg/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
Jig/ml
89.1 (12.4)
121 (14)
149(31)
Predicted
Exposure
(AUC) mg/L*h
387000 (50800)
565000(61100)
710000
(144000)
Numbers in parentheses indicate standard deviation
M= male
The AUC for the LOAEL or NOAEL of each data set can be used to determine an
average serum concentration by dividing it by the duration of the study in days with adjustment
for the hours in a day. The following equation is used for the conversion:
Average Serum Concentration = AUC (mg/L*h) x 1 day/24 hours + exposure duration
(days)
For example, in a case where the AUC was 30,000 mg/L*h and the study duration was 90
days, the Average Serum Concentration would be calculated as follows:
Average Serum Concentration = 30,000 mg/L*h H- (90 days x 24 h/day) = 13.89 mg/L
This calculation has the advantage of normalizing the serum concentration across the
exposure durations to generate a uniform metric for internal dose in situations where the dosing
durations varied and serum measurements were taken immediately prior to sacrifice. The
averaged serum concentration is a hybrid of the AUC and the maximum serum concentration.
Compared across studies, PFOA average serum concentration appears to be a stable reflection of
internal dosimetry.
Table 5-9 provides the AUC from the model, the dosing duration from each of the
modeled studies, and the resultant average serum concentration. The internal doses associated
with the effect levels (LOAELs) differ by less than an order of magnitude (20.3 mg/L to 88.6
mg/L) while the AUC values differ by over two orders of magnitude (7320 mg/L*h to 387000
mg/L*h). Given the differences in external doses, the projected serum levels are proportionally
quite similar.
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TABLE 5-9. Average Serum concentrations Derived from the AUC and the duration of Dosing
Study
Dewitt, 2008 liver
Dewitt, 2008
Immunotox
Lau, 2006 liver
Palazzolo, 1993
Liver
Thomford, 200 Ib;
Butenhoff etal.,
2004 Liver
Wolf, 2007 gd 7-
17
Liver
Wolf, 2007 gd 1-
17
Liver
York, 2002
liver
Dosing
duration
days
15
15
17
91
182
11
17
84
NOAEL
(AUC
mg/L*h)
None
1.88
(13800)
None
0.06(7310)
None
None
None
none
NOAEL
(Av serum
mg/L)
None
38.33
None
3.35
None
None
None
none
LOAEL
(AUC
mg/L*h)
0.94
(7320)
3.75
(22400)
1
(16500)
0.64
(69900)
3 (387000)
5
(23300)
o
5
(31700)
1
(93500)
LOAEL
(Av serum
mg/L)
20.33
62.22
40.44
32.00
88.60
88.26
77.70
46.38
Table 5-9 identifies 20.33 mg/L as the lowest concentration associated with an increase in
liver weight in female mice (DeWitt et al, 2008). Average serum values for increased relative
liver weight in male rats (Palazzolo et al., 1993) and female mice (Lau et al. 2006) are no more
than 2-fold higher than the average serum value in DeWitt et al. (2008). Thus, it appears that the
LOAELs are roughly consistent across gender, species, and treatment with respect average serum
concentration. Assuming that mode of action and susceptibility to toxicity do not vary and that
pharmacokinetics alone explains variation, it is reasonable to expect similar concentrations to
cause similar effects in humans.
The predicted serum concentrations can be converted into an oral equivalent dose at
steady state by recognizing that, at steady state, clearance from the body must equal dose to the
body. Clearance can be calculated if the rate of elimination (derived from half-life) and the
volume of distribution are both known.
Measures of half-life in humans have been determined for both workers and the general
population (Section 3.5.2). Olsen et al. (2007) gives the human half-life as 3.8 years for PFOA
in an occupationally exposed U.S. cohort. Bartell et al. (2010) determined a value of 2.3 years
based on the decline in serum levels among members of the general population exposed via
drinking water in the area near the Dupont Works plant in Washington, WV after the drinking
water concentrations decreased. The Office of Water has chosen to use the Bartell et al. (2010)
half-life value because it is appears to be the one most relevant to scenarios where exposures
result from ingestion of contaminated drinking water by members of the general population.
Thompson et al. (2010) gives a volume of distribution of 0.17 L/kg bw (Section 3.5.3).
The volume of distribution is defined as the total amount of PFOA in the body divided by the
blood or serum concentration. The volume of distribution was calibrated using human serum
concentrations and exposure data from NHANES and assumes that most PFOA intake came
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from contaminated drinking water. The value for volume of distribution was calibrated so that
the model prediction of elevated blood levels of PFOA was consistent with the values from
NHANES. Lack of consideration of PFOA contributions from sources other than drinking water
sources is a weakness of the volume of distribution value determined by Thompson et al. (2010),
however this estimate is not radically different from the 0.198 L/kg bw determined for the
monkey in the study by Buttanhoff et al (2004).
The half life (t/2) and volume of distribution (Vd) are utilized to calculate the clearance for
PFOA according to the following equation assuming first order kinetics for clearance (CL)
(Medinsky and Klaassen, 1996):
CL = Vd x (In 2 -H ti/2) = 0.17 L/kgbw/day x (0.693 - 839.5 days) = 0.00014 L/kg bw/day
Where:
Vd = 0.17 L/kg
In 2 = 0.693
ti/2 = 839.5 days (2.3 years x 365 days/year = 839.5 days)
These values combined give a clearance of 1.4 x 10"4 L/kg BW/day.
Scaling the derived average concentrations (in mg/L) for the NOAELs and LOAELs in
Table 5-10 gives predicted oral human equivalent doses (HEDs) in mg/kg/day for each
corresponding serum measurement. The FED values are the predicted human oral exposures
necessary to achieve serum concentrations equivalent to the LOAEL (and NOAEL where
available) in the animal toxicity studies. Note that this scaling assumes linear human kinetics in
contrast to the non-linear phenomena observed at high doses in animals.
Thus, FLED = average serum concentration (in mg/L) x CL
Where:
Average serum is from model output in Table 5-9
CL = 0.00014 L/kg bw/day
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TABLE 5-10. Human Equivalent Doses Derived from the Modeled Animal Average Serum Values
Study
Dewitt, 2008
liver
Dewitt,
Immunotox
Lau, 2006 liver
Palazzolo, 1993
Liver
Thomford,
200 Ib liver
Wolf, 2007 GD
1-17
liver
Wolf, 2007 gd
7-17
Liver
York 2002
male F0 liver
Dosing
duration
days
15
15
17
91
182
17
11
84
NOAEL
mg/kg/d
None
1.88
None
0.06
None
None
None
None
NOAEL
Av serum
mg/L
-
38.33
-
3.35
-
-
RED
mg/kg/d
-
0.0054
-
0.00047
-
-
LOAEL
mg/kg/d
0.94
3.75
1
0.64
o
J
o
J
5
1
LOAEL
(Av serum)
mg/L
20.33
62.22
40.44
32.00
88.60
77.70
88.26
46.38
RED
mg/kg/d
0.0028
0.0087
0.0057
0.0045
0.0124
0.0109
0.0124
0.0065
5.1.1.3 RfD Quantification
There are several acceptable points of departure (PODs) that can be used in the process of
identifying the POD for RfD development:
• NOAEL or LOAEL values
• Lower 95% confidence bounds on the BMD (BMDLs), and
• Human Equivalent Doses (HED).
Studies that have more than one POD for the same NOAEL or LAOEL are summarized
in Table 5-11. There are very few NOAELs that have been identified for PFOA; LOAELs are
the more common benchmark identified through classical toxicological studies. Some of the
studies with NOAELs and/or LOAELs are chronic and others are short term. Modeling of dose
response to identify a BMD and BMDL was successful for only a subset of studies. All
benchmark models targeted a 10% increase in liver weight as a common metric for effect. This
selection was not made based on toxicity but on the desire to find a common denominator against
which to evaluate dose-response across studies and justified by the fact that other adverse effects
accompanied the LOAEL for increased liver weight in some cases.
The subset of studies that were amenable for derivation of HED based on average serum
measurements from the pharmacokinetic model was limited because of the need to have dose and
species-specific serum values for model input. Most entries in Table 5-11 provide at least two
potential PODs for consideration; only one provides five values (Palazzolo et al. 1993).
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TABLE 5-11. RfD Point of Departure Options (mg/kg/day) from the PFOA Animal Studies
Studies
Dewitt, 2008
female mouse
Dewitt, 2008
female mouse
Lau, 2006, GD
1-17, mouse
Loveless etal.,
2008 male rat
Loveless etal.,
2008 mouse
male
Palazzolo et al,
1993 male rat
Wolf, 2007 gd
7-17
Liver
Wolf, 2007 gd
1-17
York 2002
male F0
NOAEL
none
1.88
none
1
0.3
0.06
None
none
none
LOAEL
0.94
3.75
1
10
1
0.64
5
3
1
BMDL10
0.247
0.152
0.0681
0.456
0.274
RED
NOAEL
0.0054
0.00047
RED
LOAEL
0.0028
0.0087
0.0057
0.0045
0.0109
0.0124
0.0065
Notes
Increased absolute and
relative liver weight;
Decreased IgM, increased
IgG, decreased absolute
and relative spleen
weights;
Increased maternal liver
weight
Increased liver weight,
focal liver necrosis
Increased liver weight,
liver hypertrophy with
single cell and focal
necrosis
Increased liver weight,
hepatocellular
hypertrophy
Increased maternal
relative liver weight 23
days after last dose,
reduced male pup body
weight at birth
Increased maternal
absolute and relative liver
weight 23 days after last
dose
Increased liver and kidney
weight
As explained previously, human data have identified significant relationships between
serum levels and specific indicators of adverse health effects but lack the exposure information
for dose-response modeling. For this reason none of the human studies provided an appropriate
POD for RfD derivation. The pharmacokinetically-modeled average serum values from the
animal studies are restricted to the animal species selected for their low dose response to oral
PFOA intakes. Extrapolation to humans adds a layer of uncertainty that must be accommodated
in deriving the RfD.
Each of the POD values represented in Table 5-11 requires a different quantification
approach. Thus, EPA has systematically examined the impact of POD on outcome through three
sets of calculations as follows:
• Derivation from 5 PODs: NOAEL, LOAEL BMDLio, HEDNOAEL, and HEDLOAEL
from the same study (Palazzolo et al., 1993)
• Derivation from the BMDLio values
• Derivation from FLED values derived from modeled average serum values.
The Palazzolo et al. (1993) study is the only one among the studies selected for
consideration as candidates for RfD determination that offers the opportunity to examine the
impact of five PODs on outcome. It is the only study with a NOAEL, LOAEL, BMDLio, and
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HED values. Thus, it is ideally suited for use in examining the impact of quantification approach
on the RfD outcome. The critical effect was a significant increase in liver weight accompanied
by hepatic hypertrophy and increased palmitoyl Co A oxidase activity. As described earlier, the
increase in liver weight was selected as a low dose common biomarker for systemic loss of
homeostasis across studies. In other studies, adverse reproductive (Lau et al., 2006) or
immunological effects in mice (Loveless et al., 2008) co-occurred in combination with the
increased liver weights at the LOAEL justifying selection of the liver as a common metric.
Table 5-12 provides the potential RfDs derived from the five individual Palazzolo et al
(1993) PODs.
TABLE 5-12. The Impact of Quantification Approach on the RfD outcome for the PODs from the Palazzolo
et al (1993) Study
POD
NOAEL
LOAEL
BMDL10
HEDNOAEL
HEDLOAEL
POD dose
mg/kg/day
0.06
0.64
0.456
0.00047
0.0045
UFH
10
10
10
10
10
UFA
219
219
219
o
5
o
3
UFL
-
10
-
-
10
UFS
10
10
10
-
-
UFD
-
-
-
-
-
UFtotal
21400
214000
21400
30
300
Potential
RfD
mg/kg/day
0.000003
0.000003
0.00002
0.00002
0.00002
There is good agreement between the adjusted PODs from the HEDs from the modeled
average serum levels and the BMDLio value. The adjusted PODs derived from the NOAEL and
LOAEL are about an order of magnitude more conservative than those from the BMDLio and
HEDs. The differences in outcome reflect the differences in the quantification approach rather
than the experimental data utilized. The BMDLio value more closely approximates the study
LOAEL rather than the study NOAEL.
Uncertainty Factor (UF) Application.
UFH A ten-fold adjustment is assigned to account for intrahuman variability and applied for all
PODs.
UFA. Determination of the interspecies uncertainty factor for the NOAEL, LOAEL and
BMDLIO requires application of the equation for first order kinetics in order to determine the
pharmacokinetic adjustment associated with differences in half-life between humans and male
rats. The equation that describes first order kinetics is as follows (Medinksy and Klaassen,
1996):
CL = Vd x (In 2 - ti/2)
Where:
Vd = 0.17L/kg
Ln 2 = 0.693
Half-life =11.5 days for rats and 839.5 days for humans
CLrat = 0.17 L/kg x (0.693 - 11.5 days) = 0.01024 L/kg/day
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= 0.17 L/kg - (0.693 -H 839.5 days) = 0.00014 L/kg/day
The ratio of CLratto CLhuman (0.01024 L/kg/day - 0.00014 L/kg/day = 73) is used as the
pharmacokinetic adjustment for differences between species. The total UFA requires an
additional 3-fold factor for species differences in pharmacodynamics (73 x 3 = 219).
UFL A ten-fold UF was applied for the LOAEL and HEDLOAEL to account for using an effect
level as the POD. A UF of 1 was used for the BMDLio and the NOAEL-derived PODs
following Agency policies.
UFS A ten-fold factor was applied to the NOAEL, LOAEL, and BMDLio PODs to account for
less than lifetime exposure. The HEDs reflect steady state serum values from the
pharmacokinetic model and reduce the UFs to 1 for the HED PODs.
UFD In all cases the uncertainty factor for the strength of the database (UFD) is 1. The data base
for oral PFOA exposure studies is essentially complete although mechanistic questions relative
to the MOA have not yet been fully elucidated.
BMDLio PODs. The lowest, species-specific benchmark dose values in Table 5-11 are those
from the Loveless et al. (2008) study in the male mouse and male rat. As was the case with the
Palazzolo et al. (1993) study, the effect modeled was a 10% change in liver weight. In this case,
the LOAELs from the Loveless et al. (2008) publication for both species included co-critical
focal hepatic cellular necrosis as well as decreased spleen weight for the mouse. The BMDLio
for the mice was lower than that for the rats (Table 5-11).
The data for the York et al. (2002) and Palazzolo et al. (1993) are included in Table 5-13
for comparison with the Loveless et al. results in the rat. In York et al. (2002), the change in
liver weight was accompanied by a change in kidney weight. A UFL of 10 was applied to
account for the intraindividual variability.
TABLE 5-13. The Impact of Quantification Approach on the RfD Outcome for PODs from the BMDLs
POD
BMDLio rat Loveless
BMDLi o mouse Loveless
BMDLi o rat Palazzolo
BMDL10 rat York
Dose
mg/kg/day
0.152
0.0681
0.456
0.274
UFH
10
10
10
10
UFA
219
150
219
219
UFL
-
-
-
-
UFS
10
10
10
10
UFD
-
-
-
-
UFtotal
21900
15000
21900
21900
Potential
RfD
mg/kg/day
0.000007
0.000005
0.00002
0.00001
The BMDLio values from the Loveless et al. (2008) 29-day studies are lower than those
from the longer-term Palazzolo et al. (1993) and York et al. (2002) studies but differ by a
maximum factor of 4 which is relatively small when considering the magnitude of the total
uncertainty factors.
Uncertainty Value Application
The UFn, UFL, UDs and UFo values are assigned as described for the Palazzolo et al. (1993) data
in Table 5-11.
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The UFA for the rat component of the Loveless et al (2008) study is the same as that described
above for the Palazzolo et al (1993) study. It is necessary to derive the pharmacokinetic
component of UFA for the mouse and monkey using half-life and volume of distribution to
calculate the clearance ratio as was done for the rat:
Mouse
CLmouse = 0.17 L/kg x (0.693 - 15.6 days) = 0.00697 L/kg/day
= 0.17 L/kg x (0.693 -H 839.5 days) = 0.00014 L/kg/day
The ratio of CLmouseto CLhuman (0.00697 L/kg/day - 0.00014 L/kg/day = 50) is used as the
pharmacokinetic adjustment for differences between species. Thus, the total UFA adds a 3-fold
factor for species differences in pharmacodynamics between mice and humans (50 x 3 = 150).
HED PODs. The HEDs derived from Palazzolo et al. (1993), Dewitt et al. (2008), Lau et al.
(2006), Thomford (2001), and York et al. (2002) were each examined as the potential basis for
the RfD. Only two of these studies, Palazzolo et al. (1993) and Dewitt et al., (2008), contained a
NOAEL from which the HED could be derived. Because the immunological effects which were
the basis of the NOAEL in the Dewitt et al. study occurred at a dose about four-fold higher than
the effects on liver weight, this value is not included in Table 5-14. The Dewitt et al. (2008)
LOAEL HED based on increased liver weight has the lowest quantitative values from among
these studies. LOAEL HEDs from the other studies are about twice the Dewitt et al. (2008)
value. The Lau et al. (2006) LOAEL HED included reduced fetal ossification; accelerated
puberty in male offspring as co-critical effects. The York et al. (2001) study included increased
kidney weight as co-critical at the LOAEL and the Palazzolo et al. (1993) study reported hepatic
hypertrophy and increased palmitoyl Co A oxidase activity.
TABLE 5-14. The Impact of Quantification Approach on the RfD Outcomes for the HEDs from the
Pharmacokinetic Model Average Serum Values
POD
PK-HEDNOAEL Palazzolo
PK-HEDLOAEL Palazzolo
PK-HEDLOAEL Dewitt
PK-HEDLOAEL Lau
PK-HEDLOAEL York
PK-HEDLOAEL Thomford
Value
mg/kg/day
0.00047
0.0045
0.0028
0.0057
0.0065
0.0124
UFH
10
10
10
10
10
10
UFA
3
3
3
3
3
3
UFL
-
10
10
10
10
10
UFS
-
-
-
-
-
-
UFD
-
-
-
-
-
UFtotal
30
300
300
300
300
300
Potential
RfD
mg/kg/day
0.00002
0.00002
0.000009
0.00002
0.00002
0.00004
The outcomes for potential RfD values as determined in Tables 5-11, 5-12, and 5-13
differ by about an order of magnitude (0.000003 mg/kg/day to 0.00004 mg/kg/day). Of the
values derived from the HED values one is 0.000009 mg/kg/day (Dewitt et al., 2008) and 4 are
0.00002 (Palazzolo et al., 1993; Lau et al., 2006; York et al., 2002). These results demonstrate
the ability of the model to normalize the animal data across species, strain, gender, and exposure
duration.
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The Palazzolo et al. (1993) and York et al. (2002) studies were conducted in male
Sprague Dawley rats with durations of 91 days via diet and 84 days via gavage, respectively.
The Lau et al. (2006) study was conducted in pregnant female rats with an 18-day average
exposure via gavage leading to increased liver weights in the dams and developmental effects in
their male offspring. The Dewitt et al. (2008) drinking water study was conducted in female
C57BL/6 mice with a 15-day exposure. The Thumford (200Ib) study was conducted in male
cynomolgus monkeys for 182 days; the PFOA was administered in capsules.
Uncertainty Value Application
The UFH, UFL, UFS and UFD values are assigned as described for the HED LOAEL entry for the
Palazzolo et al. (1993) data in Table 5-12.
The UFA is 3 for each study because the HED was derived using the steady state serum values
from the pharmacokinetic model. The 3-fold factor is applied to account for toxicodynamic
differences between the animals and humans.
RfD Selection Based on the consistency of the response and with recognition of the use of liver
weight having been chosen as a common denominator in the RfD analysis for loss of
homeostatisis and protection against co-occurring adverse effects, the 0.00002 mg/kg/day
outcome is selected as the RfD for PFOA. This value is the outcome for all modeled rat and
mouse serum values except for the Dewitt et al. (2008) 15-day study with an impact on liver
weight but not the co-monitored immunological effects. The liver endpoint in the Lau et al.
(2006) and York et al. (2002) studies were accompanied by developmental effects and effects on
kidney weights, respectively. The modeled serum value from Thumford (2001) based on liver
effects in the monkey, also strongly supports the chosen RfD.
The results from the modeled serum values are supported by the benchmark dose results
wherein the Palazzolo et al. (1993) POD supports a potential RfD of 0.00002 mg/kg/day for
change in liver weight. Other than Lau et al. (2006) all of the studies supporting an RfD of
0.00002 mg/kg/day are studies with exposures to PFOA for > 84 days meeting the duration
requirements for determination of a lifetime exposure value. Confidence in the RfD is medium
because of the uncertainties surrounding mode of action for some of the observed effects and the
role of saturable renal resorption in determining the body burden from chronic exposures.
5.1.2 RfC Determination
Derivation of the RfC requires evaluation of toxicity data for concentration-response and
identification of target organs and noncancer effects. Limited data from human epidemiology
and animal toxicity studies were available with which to evaluate the potential health effects
resulting from continuous inhalation exposure to PFOA. The available data base, summarized
below for human and animal data, does not meet the minimal requirements for derivation of the
RfC as defined by US EPA (1994). Thus, the RfC for PFOA is not recommended or derived.
Human Data. Studies have examined occupational and residential populations at or near 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
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animal toxicological database. While inhalation is an important route of exposure to workers,
drinking water was identified as the primary contributor to exposure in the general population.
In all of the epidemiology studies, wide ranges of serum levels were reported which are likely a
reflection, not only of intra-human toxicokinetic variability, but also of diversity in external
exposure sources and routes of exposure. Thus the data cannot be clearly applied to
quantification of dose-response via inhalation
Animal Data. Inhalation toxicity data in laboratory animals were limited to acute exposure,
single and repeated exposures for pharmacokinetic studies, and a developmental toxicity study in
rats. No subchronic or chronic inhalation toxicity studies in animals were available for
assessment. Generally, adverse effects observed following inhalation exposure to PFOA were
similar to effects following exposure to an irritating dust. For male rats exposed to PFOA as a
dust in air, the 4-hour 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 three
weeks (Hinderliter, 2003). Absorption was indicated in both males and females after single and
repeated exposures with plasma PFOA concentrations proportional to exposure concentration.
The Cmax values were approximately 2-3 times higher in males than in females and persisted for
up to six hours in males compared to just one hour in females. Similarly, the elimination of
PFOA was rapid by females at all exposure levels, and by 12 hours after exposure the plasma
levels had dropped below the analytical limit of quantitation (0.1 jig/ml). In males, the plasma
concentration remained approximately 90% of the peak concentration at all exposure levels at 24
hours after exposure, and steady state was reached following repeated exposures. While these
results clearly show toxicokinetic differences between male and female rats, toxicity data were
not included for use in quantitative risk assessment.
In a developmental toxicity study, pregnant Sprague-Dawley rats were exposed 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.
5.2 Dose-Response for Cancer Effects
As discussed in section 4.4.4, there is some evidence that PFOA exposure may be
associated with an increased risk for cancer from the human epidemiology database and animal
studies. The Dupont (2003b) report on the West Virginia Washington Works Plant found
statistically significant elevations in cancer of the bladder, kidney, and urinary tract organs and a
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nonstatistically significant elevation in leukemia incidence. The findings for an increased
incidence of leukemia from two earlier worker studies were statistically significant, but both had
methodological limitations (Walrath and Burke, 1989; Karns and Fayerweather, 1991). The
mortality data for liver cancer, pancreatic cancer, prostate cancer, and testicular cancer from
human epidemiological studies all showed no association to exposure.
The only chronic bioassays of PFOA were conducted in rats (Butenhoff et al., 2012;
Biegel et al, 2001). The two studies support a positive finding for the ability of PFOA to be
weakly tumorigenic in one or more organs of male but not female rats. There are no
carcinogenicity data from a second animal species. The study by Butenhoff et al. (2012)
examined males and females; the Biegel et al. study only evaluated males. The tumor types
observed were:
• Liver (Butenhoff et al., 2012)
• Leydig Cell (Butenhoff et al., 2012; Biegel et al., 2001)
• Pancreatic Ascinar Cell (Biegel et al., 2001)
The dose response information and tumors incidence data from the Butenhoff et al.
(2012) and Biegel et al. (2001) study are summarized in Table 5-15, below. The data are limited
in that only Butenhoff et al. tested more than one dose and only one tumor-type (Leydig Cell
adenoma) demonstrated a dose-response relationship.
TABLE 5-15. Summary of Tumor Data from Animal Studies
Tissue
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
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
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, mitigate concern for their relevance to humans.
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.
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PPARa activation may also play a role in the other tumor types observed, however for
the Ley dig tumors the PPARa involvement is indirect. The favored hypothesis for the DNA
replication errors responsible for induction of Ley dig tumors are postulated be a consequence of
the following sequence of events:
• Decreased testosterone synthesis
• Increased GnRH and increased levels of leutinizing hormone leading to chronic
stimulation of Ley dig cells by growth-stimulating mediators including IGF-1, TGF-a,
leukotrienes and various free radicals (Clegg et al., 1997; Li et al, 2011).
There are some experimental data that demonstrate systemic effects of PFOA leading to
decreased testosterone and increased estradiol as a result of increased activity of aromatase, the
cellular enzyme responsible for the metabolic conversion of testosterone to estradiol (Biegel et
al., 1995). However, more data to support the relationship of PFOA to intermediate steps in the
proposed MO As are needed.
Current MOA theories for the PACT tumors are linked to the impact of either the
mitogenic effects elevated testosterone levels or intestinal tissue hormones (cholecystokinin
and/or gastrin) in promoting proliferation of acinar cell preneoplastic foci (Osboourn et al., 1997;
Klaunig et al., 2003). PACT tumors are most commonly found in rats but also occur in humans.
Because PFOA is associated with decreased rather than increased levels of testosterone the
mechanistic link between PFOA exposure and PACT is more likely associated with gastric
hormone changes possibly associated with alterations in bile composition. Some of the
membrane transporters that are impacted by PFOA function in transport of bile components from
the liver to the gallbladder and thereby to the intestines. Cholecystokinen and gastrin stimulate
contraction of the gallbladder and release of bile into the intestines. Data to support this
hypothesis are not available for PFOA. Minimal data for an effect on bile composition are
available for the PPARa activating agent Wyeth-14,643 (Osborn et al., 1997).
Under the EPA 2005 cancer guidelines the evidence for the carcinogenicity of PFOA is
considered suggestive because only one species has been evaluated and the tumor responses
occurred primarily in males. Dose-response data are only available for the Leydig Tumors in
one study. In cases where the evidence is suggestive, the Agency generally does not attempt
dose-response quantification. However, if the evidence includes a well-conducted study with
adequate data to support modeling of dose-response, quantitative analyses may be useful in
providing a sense of the magnitude and uncertainty of potential risks, ranking potential hazards,
or setting research priorities.
The increase in hepatocellular tumors did not show a direct relationship to dose in male
rats and was not significantly elevated in either males or females at the high dose when
compared to controls. There was a dose-related significant increase in Leydig cell tumors in
male rats in the Butenhoff et al. (2012) study which was confirmed by the high dose in the single
dose mechanistic study by Biegel et al. (2011). At the high dose (300 ppm in the diet; 14.2
mg/kg/day) tumors were found in 14% of the male rats at the end of 2 years in the Butenhoff et
al. (2012) study and 4% at the low dose (1.3 mg/kg/day). In the Biegel et al (2001) study 11%
were affected at a dose of 300 ppm in the diet (13.6 mg/kg/day). In each case there were no
Leydig cell tumors in the controls. The PACT tumors, only detected in the single dose Biegel et
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al (2001) study, do not support quantification.
EPA modeled the dose-response for the Ley dig cell tumors from Butenhoff et al. (2012)
using the Agency benchmark dose software (BMDS) v 2.3.1. The multistage cancer model was
used for a dichotomous dataset to predict the dose at which a 4% increase in tumor incidence
would occur. The 4% was chosen as the low-end of the observed response range within the
study results. Both the first and second degree polynomials gave identical goodness-of-fit
criteria (p value and AIC). Results are shown in Table 5-16.
TABLE 5-16. Model Results for Leydig Cell Tumors (Buttenhoff et al. (2012)
4% increase
4% increase
AIC = 62.6936
BMD (mg/kg/day)
3.51
3.51
P = 0.2245
BMDL (mg/kg/day)
1.99
1.99
Multistage Cancer Model with 0.95 Confidence Level
Multistage Cancer
Linear extrapolation
BMDL BMD
10 12 14
11:5905/092013
FIGURE 5-2. BMD Model Results for Leydig Cell Tumors (Buttenhoff et al., 2012)
The Cancer Slope Factor (CSF) for PFOS is is derived from the BMDL04 of 1.99
mg/kg/day after converting the animal BMDL to a Human Equivalent Dose (FED) using body
weights the % power. The FLED is calculated as follows:
FLED = Animal BMDL x (animal body weight)174 + (human body weight)174
FffiD = 1.99 mg/kg/day x [(0.523 kg1)174 - (70 kg)174] = 1.99 mg/kg/day x 0.29 = 0.58
mg/kg/day
Body weight for male Sprague Dawley rats (chronic Exposures) U.S. EPA, 1988
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5-27
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Where:
1.99 mg/kg/day = BMDL04 for Ley dig Cell tumors
0.29 = The dosimetric adjustment factor
The CSF is calculated from the BMDL04 HED as follows
CSF = response - BMDL04 HED
CSF = 0.04 -H 0.58 mg/kg/day = 0.07 (mg/kg/day)'1
The CSF should not be used at doses greater than 0.58 mg/kg/day which is the human
equivalent dose corresponding to the point of departure for the 4% incidence of Ley dig Cell
tumors follow lifetime exposure to PFOA. The observed dose-response relationships do not
continue linearly above this level and the fitted dose-response models better characterizes the
dose-response for the higher exposures. The FLED of 0.58 mg/kg/day is nearly 30,000-fold
greater than the RfD (0.58 mg/kg/day - 0.00002 mg/kg/day = 29,000). Thus, the proposed RfD,
based on increased liver weight as a common denominator for loss of homeostasis and protection
of co-occurring effects, will also be protective of Ley dig cell tumors.
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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/6mice. Toxicol. Sci. doi:10.1093/toxsci/kfq030.
Perfluorooctanoic Acid - February 2014 6-22
Draft - Do Not Cite or Quote
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APPENDIX A
Contents: Benchmark Dose Modeling Output Files 10% increased liver weight for:
1) Palazzolo, 1993; Perkins et al., 2004: male rat;
2) Loveless et al., 2008: male rat;
3) York, 2002; Butenhoff et al., 2004a; York et al., 2010: male rat;
4) Loveless etal., 2008: male mouse;
5) Lau et al., 2006: female mouse
Perfluorooctanoic Acid - February 2014 A-l
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1) Palazzolo, 1993; Perkins et al., 2004: male rat
Exponential Model. (Version: 1.7; Date: 12/10/2009)
Thu Aug 18 07:37:14 2011
BMDS Model Run
The form of the response function by Model:
Model 2: Y[dose] = a * expfsign * b * dose}
Model 3: Y[dose] = a * expfsign * (b * dose)Ad}
Model 4: Y[dose] = a * [c-(c-l) * exp{-b * dose}]
Model 5: Y[dose] = a * [c-(c-l) * exp{-(b * dosej^
Note: Y[dose] is the median response for exposure = dose;
sign = +1 for increasing trend in data;
sign = -1 for decreasing trend.
Model 2 is nested within Models 3 and 4.
Model 3 is nested within Model 5.
Model 4 is nested within Model 5.
Variable
Lnalpha
rho
a
b
c
d
Dependent variable = Mean
Independent variable = Dose
Data are assumed to be distributed: normally
Variance Model: exp(lnalpha +rho *ln(Y[dose]))
The variance is to be modeled as Var(i) = exp(lalpha + log(mean(i)) *
Total number of dose groups = 5
Total number of records with missing values = 0
Maximum number of iterations = 250
Relative Function Convergence has been set to: le-008
Parameter Convergence has been set to: le-008
MLE solution provided: Exact
Initial Parameter Values
rho)
Model 2
-11.0843
4.39551
19.3889
0.0525111
Model 3
-11.0843
4.39551
19.3889
0.0525111
1
Model 4
-11.0843
4.39551
17.1285
0.329637
1.64165
Model 5
-11.0843
4.39551
17.1285
0.329637
1.64165
1
Variable
Lnalpha
rho
a
b
c
d
Model 2
-10.3398
4.17399
19.3333
0.055133
Parameter Estimates by Model
Model 3 Model 4
-10.3398
4.17399
19.3333
0.055133
1
-11.3983
4.50168
18.8392
0.281713
1.49669
Model 5
-11.4672
4.523
18.9098
0.393804
1.42426
1.2698
Perfluorooctanoic Acid - February 2014
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A-2
-------
Table of Stats From Input Data
Dose N Obs Mean Obs Std Dev
0
0.06
0.64
1.94
6.5
15
15
15
15
15
19.73
18.03
20.44
22.74
26.78
2.01
2.81
2.87
4.21
5.47
Estimated Values of Interest
Model
2
3
4
5
Dose
0
0.06
0. 64
1.94
6.5
0
0.06
0. 64
1.94
6.5
0
0.06
0. 64
1.94
6.5
0
0.06
0.64
1.94
6.5
Est Mean
19.33
19.4
20.03
21.52
27.67
19.33
19.4
20.03
21.52
27.67
18.84
19
20.38
22.78
26.7
18.91
18.98
20.19
22.99
26. 64
Est Std
2.75
2.769
2.96
3.437
5.809
2.75
2.769
2.96
3.437
5.809
2.482
2.529
2.964
3.806
5.441
2.496
2.516
2.894
3.882
5.416
Scaled Residual
0.5588
-1.913
0.5396
1.379
-0.5906
0.5588
-1.913
0.5396
1.379
-0.5906
1.39
-1.479
0.07453
-0.03971
0.05896
1.273
-1.46
0.3352
-0.2494
0.1028
Other models for which likelihoods are calculated:
Model Al: Yij = Mu(i) + e(ij)
Var{e(ij)} = SigmaA2
Model A2: Yij = Mu(i) + e(ij)
Var{e(ij)} = Sigma(i)^2
Model A3: Yij = Mu(i) + e(ij)
Var{e(ij)} = exp(lalpha + log(mean(i)) * rho)
Model R: Yij = Mu + e(i)
Var{e(ij)} = Sigma/x2
Model
Likelihoods of Interest
Log(likelihood) DF
AIC
Al
A2
A3
R
2
3
-132.6899
-123.7484
-125.6728
-153.0718
-129.1329
-129.1329
6
10
7
2
4
4
277.3799
267.4967
265.3456
310.1437
266.2658
266.2658
Perfluorooctanoic Acid - February 2014
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A-3
-------
-126.8109
-126.6814
263.6218
265.3627
Additive constant for all log-likelihoods = -68.92. This constant added to
the above values gives the log-likelihood including the term that does not
depend on the model parameters.
Explanation of Tests
Test 1: Does response and/or variances differ among Dose levels? (A2 vs. R)
Test 2: Are Variances Homogeneous? (A2 vs. Al)
Test 3: Are variances adeguately modeled? (A2 vs. A3)
Test 4: Does Model 2 fit the data? (A3 vs. 2)
Test 5a: Does Model 3 fit the data? (A3 vs 3)
Test 5b: Is Model 3 better than Model 2? (3 vs. 2)
Test 6a: Does Model 4 fit the data? (A3 vs 4)
Test 6b: Is Model 4 better than Model 2? (4 vs. 2)
Test 7a: Does Model 5 fit the data? (A3 vs 5)
Test 7b: Is Model 5 better than Model 3? (5 vs. 3)
Test 7c: Is Model 5 better than Model 4? (5 vs. 4)
Tests of Interest
Test
Test 1
Test 2
Test 3
Test 4
Test 5a
Test 5b
Test 6a
Test 6b
Test 7a
Test 7b
Test 7c
-2*log(Likelihood Ratio) D. F.
58.65 8
17.88 4
3.849 3
6.92 3
6.92 3
0 0
2.276 2
4.644 1
2.017 1
4.903 2
0.2591 1
p-value
< 0.0001
0.001301
0.2783
0.07449
0.07449
N/A
0.3204
0.03116
0.1555
0.08616
0.6108
The p-value for Test 1 is less than .05. There appears to be a
difference between response and/or variances among the dose levels, it seems
appropriate to model the data.
The p-value for Test 2 is less than .1.
appears to be appropriate.
A non-homogeneous variance model
The p-value for Test 3 is greater than .1. The modeled variance appears
to be appropriate here.
The p-value for Test 4 is less than .1. Model 2 may not adeguately
describe the data; you may want to consider another model.
The p-value for Test 5a is less than .1. Model 3 may not adeguately
describe the data; you may want to consider another model.
Degrees of freedom for Test 5b are less than or egual to 0. The Chi-Sguare
test for fit is not valid.
Perfluorooctanoic Acid - February 2014
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A-4
-------
The p-value for Test 6a is greater than .1. Model 4 seems to adequately
describe the data.
The p-value for Test 6b is less than .05. Model 4 appears to fit the data
better than Model 2.
The p-value for Test 7a is greater than .1. Model 5 seems to adequately
describe the data.
The p-value for Test 7b is greater than .05. Model 5 does not seem to fit
the data better than Model 3.
The p-value for Test 7c is greater than .05. Model 5 does not seem to fit
the data better than Model 4.
Benchmark Dose Computations:
Specified Effect = 0.100000
Risk Type = Relative deviation
Confidence Level = 0.950000
BMD and BMDL by Model
Model BMD BMDL
2
3
4
5
1.72873
1.72873
0.798012
0.902365
1.34495
1.34495
0.456224
0.468724
Exponential Model 4 with 0.95 Confidence Level
c
o
Q.
in
-------
2) Loveless etal., 2008: male rat
Exponential Model. (Version: 1.7; Date: 12/10/2009)
Thu Dec 22 12:43:43 2011
BMDS Model Run
The form of the response function by Model:
Model 2: Y[dose] = a * expfsign * b * dose}
Model 3: Y[dose] = a * expfsign * (b * dose)Ad}
Model 4: Y[dose] = a * [c-(c-l) * exp{-b * dose}]
Model 5: Y[dose] = a * [c-(c-l) * exp{-(b * dosej^
Note: Y[dose] is the median response for exposure = dose;
sign = +1 for increasing trend in data;
sign = -1 for decreasing trend.
Model 2 is nested within Models 3 and 4.
Model 3 is nested within Model 5.
Model 4 is nested within Model 5.
Dependent variable = Mean
Independent variable = Dose
Data are assumed to be distributed: normally
Variance Model: exp(lnalpha +rho *ln(Y[dose]))
The variance is to be modeled as Var(i) = exp(lalpha + log(mean(i)) *
Total number of dose groups = 5
Total number of records with missing values = 0
Maximum number of iterations = 250
Relative Function Convergence has been set to: le-008
Parameter Convergence has been set to: le-008
MLE solution provided: Exact
rho)
Variable
Model 2
Initial Parameter Values
Model 3 Model 4
Model 5
Lnalpha
rho
a
b
c
d
-7.96162
3.3934
15.5663
0.00879874
-7.96162
3.3934
15.5663
0.00879874
-7.96162
3.3934
12.54
0.0515133
1.80024
-7.96162
3.3934
12.54
0.0515133
1.80024
1
Variable
Lnalpha
rho
a
b
c
d
Model 2
-4.07115
2.30379
15.7112
0.00896759
Parameter Estimates by Model
Model 3 Model 4 Model 5
-4.07115
2.30379
15.7112
0.00896759
-10.0648
4.14318
13.0813
0.861545
1.5326
-10.3628
4.24407
13.1947
0.977136
1.51565
1.36232
Perfluorooctanoic Acid - February 2014
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A-6
-------
Table of Stats From Input Data
Dose N Obs Mean Obs Std Dev
0
0.3
1
10
30
10
10
10
10
10
13.2
14.4
17.2
21.5
18.7
1.4
1.6
2.9
2.9
2.9
Estimated Values of Interest
Model
2
3
4
5
Dose
0
0.3
1
10
30
0
0.3
1
10
30
0
0.3
1
10
30
0
0.3
1
10
30
Est Mean
15.
15.
15.
17.
20.
15.
15.
15.
17.
20.
13.
14.
17
20.
20.
13.
14.
17.
71
75
85
19
56
71
75
85
19
56
08
67
.1
05
05
19
36
42
20
20
Est Std
3.118
3.128
3.15
3.457
4.251
3.118
3.128
3.15
3.457
4.251
1.342
1.701
2.339
3.249
3.25
1.34
1. 604
2.416
3.24
3.24
Scaled Residual
-2.547
-1.368
1.352
3.947
-1.385
-2.547
-1.368
1.352
3.947
-1.385
0.2798
-0.4985
0.1288
1.414
-1.312
0.01248
0.07811
-0.2836
1.466
-1.268
Other models for which likelihoods are calculated:
Model Al: Yij = Mu(i) + e(ij)
Var{e(ij)} = SigmaA2
Model A2: Yij = Mu(i) + e(ij)
Var{e(ij)} = Sigma(i)^2
Model A3: Yij = Mu(i) + e(ij)
Var{e(ij)} = exp(lalpha + log(mean(i)) * rho)
Model R: Yij = Mu + e(i)
Var{e(ij)} = Sigma/x2
Likelihoods of Interest
Model
Log(likelihood)
DF
AIC
Al
A2
A3
R
2
3
4
-66.
-62.
-63.
-91.
-86.
-86.
-65.
,95077
,37207
,18657
,38557
,12482
,12482
,31811
6
10
7
2
4
4
5
145.
144.
140.
186.
180.
180.
140.
,9015
,7441
,3731
,7711
,2496
,2496
,6362
Perfluorooctanoic Acid - February 2014
Draft - Do Not Cite or Quote
A-7
-------
-64.98792
141.9758
Additive constant for all log-likelihoods = -45.95. This constant added to
the above values gives the log-likelihood including the term that does not
depend on the model parameters.
Explanation of Tests
Test 1: Does response and/or variances differ among Dose levels? (A2 vs. R)
Test 2: Are Variances Homogeneous? (A2 vs. Al)
Test 3: Are variances adeguately modeled? (A2 vs. A3)
Test 4: Does Model 2 fit the data? (A3 vs. 2)
Test 5a: Does Model 3 fit the data? (A3 vs 3)
Test 5b: Is Model 3 better than Model 2? (3 vs. 2)
Test 6a: Does Model 4 fit the data? (A3 vs 4)
Test 6b: Is Model 4 better than Model 2? (4 vs. 2)
Test 7a: Does Model 5 fit the data? (A3 vs 5)
Test 7b: Is Model 5 better than Model 3? (5 vs. 3)
Test 7c: Is Model 5 better than Model 4? (5 vs. 4)
Tests of Interest
Test
Test 1
Test 2
Test 3
Test 4
Test 5a
Test 5b
Test 6a
Test 6b
Test 7a
Test 7b
Test 7c
-2*log(Likelihood Ratio) D. F.
58.03 8
9.157 4
1.629 3
45.88 3
45.88 3
-1.563e-012 0
4.263 2
41.61 1
3.603 1
42.27 2
0.6604 1
p-value
< 0.0001
0.05728
0.6528
< 0.0001
< 0.0001
N/A
0.1187
< 0.0001
0.05769
< 0.0001
0.4164
The p-value for Test 1 is less than .05. There appears to be a difference
between response and/or variances among the dose levels, it seems appropriate
to model the data.
The p-value for Test 2 is less than .1. A non-homogeneous variance model
appears to be appropriate.
The p-value for Test 3 is greater than .1. The modeled variance appears
to be appropriate here.
The p-value for Test 4 is less than .1. Model 2 may not adeguately
describe the data; you may want to consider another model.
The p-value for Test 5a is less than .1. Model 3 may not adeguately
describe the data; you may want to consider another model.
Degrees of freedom for Test 5b are less than or egual to 0. The Chi-Sguare
test for fit is not valid.
The p-value for Test 6a is greater than .1. Model 4 seems to adeguately
describe the data.
Perfluorooctanoic Acid - February 2014
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A-8
-------
The p-value for Test 6b is less than .05. Model 4 appears to fit the data
better than Model 2.
The p-value for Test 7a is less than .1. Model 5 may not adequately
describe the data; you may want to consider another model.
The p-value for Test 7b is less than .05. Model 5 appears to fit the data
better than Model 3.
The p-value for Test 7c is greater than .05. Model 5 does not seem to fit
the data better than Model 4.
Benchmark Dose Computations:
Specified Effect = 0.100000
Risk Type = Relative deviation
Confidence Level = 0.950000
Model
2
3
4
5
BMD and BMDL by Model
BMD
10.6283
10.6283
0.241376
0.331813
BMDL
6.02231
6.02231
0.151833
0.161379
Perfluorooctanoic Acid - February 2014
Draft - Do Not Cite or Quote
A-9
-------
Exponential Model 4 with 0.95 Confidence Level
c
o
Q.
(/)
0)
o:
c
(0
OJ
24
22
20
18
16
14
12
BMDLBMD
Exponential
T
10
15
dose
20
25
30
13:1212/222011
Perfluorooctanoic Acid - February 2014
Draft - Do Not Cite or Quote
A-10
-------
3) York, 2002; Butenhoff et al., 2004a; York et al., 2010
Exponential Model. (Version: 1.7; Date: 12/10/2009)
Wed Oct 26 10:57:26 2011
BMDS Model Run
The form of the response function by Model:
Model 2: Y[dose] = a * expfsign * b * dose}
Model 3: Y[dose] = a * expfsign * (b * dose)Ad}
Model 4: Y[dose] = a * [c-(c-l) * exp{-b * dose}]
Model 5: Y[dose] = a * [c-(c-l) * exp{-(b * dosej^d}]
Note: Y[dose] is the median response for exposure = dose;
sign = +1 for increasing trend in data;
sign = -1 for decreasing trend.
Model 2 is nested within Models 3 and 4.
Model 3 is nested within Model 5.
Model 4 is nested within Model 5.
Dependent variable = Mean
Independent variable = Dose
Data are assumed to be distributed: normally
Variance Model: exp(lnalpha +rho *ln(Y[dose]))
The variance is to be modeled as Var(i) = exp(lalpha + log(mean(i)) * rho)
Total number of dose groups = 5
Total number of records with missing values = 0
Maximum number of iterations = 250
Relative Function Convergence has been set to: le-008
Parameter Convergence has been set to: le-008
MLE solution provided: Exact
Variable
Model 2
Initial Parameter Values
Model 3 Model 4
Model 5
Inalpha
rho
a
b
c
d
-3.83873
1.89443
24.2168
0.00591038
-3.83873
1.89443
24.2168
0.00591038
-3.83873
1.89443
19.285
0.0671202
1.56261
-3.83873
1.89443
19.285
0.0671202
1.56261
1
Variable
Inalpha
rho
a
b
c
d
Model 2
5.96811
-0.961701
24.5758
0.0049661
Parameter Estimates by Model
Model 3 Model 4
5.96811
-0.961702
24.5758
0.0049661
1
-4.75762
2.17955
20.2791
0.737783
1.39208
Model 5
-4.61574
2.13551
20.2956
0.745578
1.38868
1.12767
Perfluorooctanoic Acid - February 2014
Draft - Do Not Cite or Quote
A-ll
-------
Table of Stats From Input Data
Dose N Obs Mean Obs Std Dev
0 30
1 30
3 30
10 30
30 29
20.3
24.3
27.7
28.7
27.5
2.5
3.2
2.7
3.9
3.7
Estimated Values of Interest
Model
2
3
4
5
Dose
0
1
3
10
30
0
1
3
10
30
0
1
3
10
30
0
1
3
10
30
Est Mean
24.58
24.7
24.94
25.83
28.52
24.58
24.7
24.94
25.83
28.52
20.28
24.43
27.36
28.23
28.23
20.3
24.34
27.52
28.18
28.18
Est Std
4.24
4.23
4.209
4.14
3.947
4.24
4.23
4.209
4.14
3.947
2.462
3.016
3.412
3.53
3.531
2.476
3.005
3.427
3.515
3.515
Scaled Residual
-5.524
-0.5157
3.585
3.801
-1.397
-5.524
-0.5157
3.585
3.801
-1.397
0.04647
-0.2328
0.5445
0.7368
-1.114
0.009804
-0.06777
0.2834
0.8048
-1.048
Other models for which likelihoods are calculated:
Model Al: Yij = Mu(i) + e(ij)
Var{e(ij)} = SigmaA2
Model A2: Yij = Mu(i) + e(ij)
Var{e(ij)} = Sigma(i)^2
Model A3: Yij = Mu(i) + e(ij)
Var{e(ij)} = exp(lalpha + log(mean(i)) * rho)
Model R: Yij = Mu + e(i)
Var{e(ij)} = Sigma/x2
Likelihoods of Interest
Model
Log(likelihood)
DF
AIC
Al
A2
A3
R
2
3
4
5
-247.2345
-242.9088
-244.9146
-296.4613
-286.6477
-286.6477
-245.8857
-245.8276
6
10
7
2
4
4
5
6
506.469
505.8177
503.8291
596.9226
581.2954
581.2954
501.7713
503. 6552
Perfluorooctanoic Acid - February 2014
Draft - Do Not Cite or Quote
A-12
-------
Additive constant for all log-likelihoods = -136.9. This constant added to
the above values gives the log-likelihood including the term that does not
depend on the model parameters.
Explanation of Tests
Test 1: Does response and/or variances differ among Dose levels? (A2 vs. R)
Test 2: Are Variances Homogeneous? (A2 vs. Al)
Test 3: Are variances adeguately modeled? (A2 vs. A3)
Test 4: Does Model 2 fit the data? (A3 vs. 2)
Test 5a: Does Model 3 fit the data? (A3 vs 3)
Test 5b: Is Model 3 better than Model 2? (3 vs. 2)
Test 6a: Does Model 4 fit the data? (A3 vs 4)
Test 6b: Is Model 4 better than Model 2? (4 vs. 2)
Test la: Does Model 5 fit the data? (A3 vs 5)
Test 7b: Is Model 5 better than Model 3? (5 vs. 3)
Test 7c: Is Model 5 better than Model 4? (5 vs. 4)
Tests of Interest
Test
Test 1
Test 2
Test 3
Test 4
Test 5a
Test 5b
Test 6a
Test 6b
Test 7a
Test 7b
Test 7c
-2*log(Likelihood Ratio) D. F.
107.1 8
8.651 4
4.011 3
83.47 3
83.47 3
-2.274e-013 0
1.942 2
81.52 1
1.826 1
81.64 2
0.1162 1
p-value
< 0.0001
0.07043
0.2602
< 0.0001
< 0.0001
N/A
0.3787
< 0.0001
0.1766
< 0.0001
0.7332
The p-value for Test 1 is less than .05. There appears to be a difference
between response and/or variances among the dose levels, it seems appropriate
to model the data.
The p-value for Test 2 is less than .1. A non-homogeneous variance model
appears to be appropriate.
The p-value for Test 3 is greater than .1. The modeled variance appears
to be appropriate here.
The p-value for Test 4 is less than .1. Model 2 may not adeguately
describe the data; you may want to consider another model.
The p-value for Test 5a is less than .1. Model 3 may not adeguately
describe the data; you may want to consider another model.
Degrees of freedom for Test 5b are less than or egual to 0. The Chi-
Sguare test for fit is not valid.
The p-value for Test 6a is greater than .1. Model 4 seems to adeguately
describe the data.
Perfluorooctanoic Acid - February 2014
Draft - Do Not Cite or Quote
A-13
-------
The p-value for Test 6b is less than .05. Model 4 appears to fit the data
better than Model 2.
The p-value for Test 7a is greater than .1. Model 5 seems to adequately
describe the data.
The p-value for Test 7b is less than .05. Model 5 appears to fit the data
better than Model 3.
The p-value for Test 7c is greater than .05. Model 5 does not seem to fit
the data better than Model 4.
Benchmark Dose Computations:
Specified Effect = 0.100000
Risk Type = Relative deviation
Confidence Level = 0.950000
Model
BMD and BMDL by Model
BMD
BMDL
2
3
4
5
19.1922
19.1922
0.399087
0.45764
13.6199
13.6199
0.274007
0.276304
30
28
26
24
22
20
BMDLBMD
Exponential Model 4 with 0.95 Confidence Level
Exponential
10
10:5710/262011
15
dose
20
25
30
Perfluorooctanoic Acid - February 2014
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A-14
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4) Loveless et al., 2008: male mouse
Exponential Model. (Version: 1.7; Date: 12/10/2009)
Wed Aug 24 10:48:22 2011
BMDS Model Run
The form of the response function by Model:
Model 2: Y[dose] = a * expfsign * b * dose}
Model 3: Y[dose] = a * expfsign * (b * dose)Ad}
Model 4: Y[dose] = a * [c-(c-l) * exp{-b * dose}]
Model 5: Y[dose] = a * [c-(c-l) * exp{-(b * dosej^
Note: Y[dose] is the median response for exposure = dose;
sign = +1 for increasing trend in data;
sign = -1 for decreasing trend.
Model 2 is nested within Models 3 and 4.
Model 3 is nested within Model 5.
Model 4 is nested within Model 5.
Dependent variable = Mean
Independent variable = Dose
Data are assumed to be distributed: normally
Variance Model: exp(lnalpha +rho *ln(Y[dose]))
rho is set to 0.
A constant variance model is fit.
Total number of dose groups = 5
Total number of records with missing values = 0
Maximum number of iterations = 250
Relative Function Convergence has been set to: le-008
Parameter Convergence has been set to: le-008
MLE solution provided: Exact
Variable
Inalpha
rho(S)
a
b
c
d
Model 2
-0.698397
0
2.65446
0.0323602
Initial Parameter Values
Model 3 Model 4
-0.698397
0
2.65446
0.0323602
-0.698397
0
1.691
0.0970613
3.76286
Model 5
-0.698397
0
1.691
0.0970613
76286
(S) = Specified
Perfluorooctanoic Acid - February 2014
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A-15
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Variable
Inalpha
rho
a
b
c
d
Model 2
0.636609
0
3.06456
0.0242388
Parameter Estimates by Model
Model 3 Model 4 Model 5
0.636609
0
3.06456
0.0242388
-0.686179
0
1.8131
0.446049
3.30881
-0.686255
0
1.8194
0.450771
3.29563
1.01732
Table of Stats From Input Data
Dose N Obs Mean Obs Std Dev
0 20
1 20
0.3 20
30 20
10 20
1.78
3.27
2.41
5.9
6.06
0.18
0.23
0.26
0.85
1.32
Estimated Values of Interest
Model
2
3
4
5
Dose
0
1
0.3
30
10
0
1
0.3
30
10
0
1
0.3
30
10
0
1
0.3
30
10
Est Mean
3.065
3.14
3.087
6.341
3.905
3.065
3.14
3.087
6.341
3.905
1.813
3.319
2.337
5.999
5.951
1.819
3.318
2.331
5.996
5.955
Est Std
1.375
1.375
1.375
1.375
1.375
1.375
1.375
1.375
1.375
1.375
0.7096
0.7096
0.7096
0.7096
0.7096
0.7095
0.7095
0.7095
0.7095
0.7095
Scaled Residual
-4.179
0.4237
-2.202
-1.435
7.01
-4.179
0.4237
-2.202
-1.435
7.01
-0.2086
-0.3117
0.4575
-0. 6252
0.688
-0.2483
-0.3055
0.4988
-0.6055
0.6605
Other models for which likelihoods are calculated:
Model Al: Yij = Mu(i) + e(ij)
Var{e(ij)} = SigmaA2
Model A2: Yij = Mu(i) + e(ij)
Var{e(ij)} = Sigma(1)^2
Model A3: Yij = Mu(i) + e(ij)
Var{e(ij)} = exp(lalpha + log(mean(i)) * rho)
Model R: Yij = Mu + e(i)
Var{e(ij)} = Sigma/x2
Likelihoods of Interest
Perfluorooctanoic Acid - February 2014
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A-16
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Model
Log(likelihood)
DF
AIC
Al
A2
A3
R
2
3
4
5
-15.08015
40.89337
-15.08015
-114.7709
-81.83045
-81.83045
-15.69106
-15.68723
6
10
6
2
3
3
4
5
42.16031
-61.78674
42.16031
233.5419
169. 6609
169. 6609
39.38211
41.37446
Additive constant for all log-likelihoods = -91.89. This constant added
to the above values gives the log-likelihood including the term that does not
depend on the model parameters.
Explanation of Tests
Test 1: Does response and/or variances differ among Dose levels? (A2 vs. R)
Test 2: Are Variances Homogeneous? (A2 vs. Al)
Test 3: Are variances adeguately modeled? (A2 vs. A3)
Test 4: Does Model 2 fit the data? (A3 vs. 2)
Test 5a: Does Model 3 fit the data? (A3 vs 3)
Test 5b: Is Model 3 better than Model 2? (3 vs. 2)
Test 6a: Does Model 4 fit the data? (A3 vs 4)
Test 6b: Is Model 4 better than Model 2? (4 vs. 2)
Test 7a: Does Model 5 fit the data? (A3 vs 5)
Test 7b: Is Model 5 better than Model 3? (5 vs. 3)
Test 7c: Is Model 5 better than Model 4? (5 vs. 4)
Tests of Interest
Test
Test 1
Test 2
Test 3
Test 4
Test 5a
Test 5b
Test 6a
Test 6b
Test 7a
Test 7b
Test 7c
-2*log(Likelihood Ratio) D. F.
p-value
311.3 8
111.9 4
111.9 4
133.5 3
133.5 3
-7.788e-012 0
1.222 2
132.3 1
1.214 1
132.3 2
0.007649 1
< 0.0001
< 0.0001
< 0.0001
< 0.0001
< 0.0001
N/A
0.5429
< 0.0001
0.2705
< 0.0001
0.9303
The p-value for Test 1 is less than .05. There appears to be a difference
between response and/or variances among the dose levels, it seems appropriate
to model the data.
The p-value for Test 2 is less than .1.
homogeneous variance model.
The p-value for Test 3 is less than .1.
different variance model.
Consider running a non-
You may want to consider a
The p-value for Test 4 is less than .1. Model 2 may not adeguately
describe the data; you may want to consider another model.
Perfluorooctanoic Acid - February 2014
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A-17
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The p-value for Test 5a is less than .1. Model 3 may not adequately
describe the data; you may want to consider another model.
Degrees of freedom for Test 5b are less than or equal to 0. The Chi-Square
test for fit is not valid.
The p-value for Test 6a is greater than .1. Model 4 seems to adequately
describe the data.
The p-value for Test 6b is less than .05. Model 4 appears to fit the data
better than Model 2.
The p-value for Test 7a is greater than .1. Model 5 seems to adequately
describe the data.
The p-value for Test 7b is less than .05. Model 5 appears to fit the data
better than Model 3.
The p-value for Test 7c is greater than .05. Model 5 does not seem to fit
the data better than Model 4.
Benchmark Dose Computations:
Specified Effect = 0.100000
Risk Type = Relative deviation
Confidence Level = 0.950000
Model
2
3
4
5
BMD and BMDL by Model
BMD
3.93213
3.93213
0.099268
0.104178
BMDL
3.38551
3.38551
0.0680743
0.0681103
Perfluorooctanoic Acid - February 2014
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A-18
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c
o
Q.
in
a)
CC
5 -
4 -
3 -
Exponential Model 4 with 0.95 Confidence Level
Exponential
10
10:4808/242011
15
dose
20
25
30
Perfluorooctanoic Acid - February 2014
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A-19
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5) Lau et al., 2006: female mouse
Exponential Model. (Version: 1.7; Date:
12/10/2009)
Tue Oct 25 15:12:47 2011
BMDS Model Run
The form of the response function by Model:
Model 2: Y[dose] = a * expfsign * b * dose}
Model 3: Y[dose] = a * expfsign * (b * dose)Ad}
Model 4: Y[dose] = a * [c-(c-l) * exp{-b * dose}]
Model 5: Y[dose] = a * [c-(c-l) * exp{-(b * dosej^
Note: Y[dose] is the median response for exposure = dose;
sign = +1 for increasing trend in data;
sign = -1 for decreasing trend.
Model 2 is nested within Models 3 and 4.
Model 3 is nested within Model 5.
Model 4 is nested within Model 5.
Dependent variable = Mean
Independent variable = Dose
Data are assumed to be distributed: normally
Variance Model: exp(lnalpha +rho *ln (Y[dose]))
The variance is to be modeled as Var(i) = exp(lalpha + log(mean(i))
Total number of dose groups = 6
Total number of records with missing values = 0
Maximum number of iterations = 250
Relative Function Convergence has been set to: le-008
Parameter Convergence has been set to: le-008
MLE solution provided: Exact
rho)
Variable
Model 2
Initial Parameter Values
Model 3 Model 4 Model 5
Inalpha
rho
a
b
c
d
-5.21753
2.68479
2.91529
0.0531403
-5.21753
2.68479
2.91529
0.0531403
-5.21753
2.68479
2.3085
0.149585
2.46979
-5.21753
2.68479
2.3085
0.149585
2.46979
1
Variable
Inalpha
rho
a
b
c
d
Model 2
-5.42859
3.52242
2.63415
0.0990961
Parameter Estimates by Model
Model 3 Model 4 Model 5
-5.42858
3.52242
2.63415
0.0990961
1
-4.62127
2.25579
2.4311
0.279706
2.27404
-4.62765
2.26038
2.43306
0.286892
2.25849
1.03267
Table of Stats From Input Data
Dose N Obs Mean Obs Std Dev
Perfluorooctanoic Acid - February 2014
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A-20
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0
1
3
5
10
20
40
15
16
20
15
5
2.43
3.17
4.3
4.66
5.43
5.36
0.29
0.27
0.51
0.56
0.74
0.77
Estimated Values of Interest
Model
2
3
4
5
Dose
0
1
3
5
10
20
0
1
3
5
10
20
0
1
3
5
10
20
0
1
3
5
10
20
Est Mean
2.634
2.909
3.546
4.323
7.096
19.12
2.634
2.909
3.546
4.323
7.096
19.12
2.431
3.187
4.19
4.763
5.34
5.517
2.433
3.17
4.195
4.778
5.338
5.488
Est Std
0.3648
0.4343
0.6158
0.873
2.089
11.97
0.3648
0.4343
0.6158
0.873
2.089
11.97
0.2702
0.3666
0.4992
0.5769
0.6562
0.6809
0.2701
0.3643
0.4999
0.5791
0.6564
0.6773
Scaled Residual
-3.54
2.331
4.897
1.724
-3.088
-2.57
-3.54
2.331
4.897
1.724
-3.088
-2.57
-0.02564
-0.1777
0.8807
-0.8022
0.5341
-0.5152
-0.07155
-0.002448
0.8422
-0.91
0.5436
-0.4226
Other models for which likelihoods are calculated:
Model Al: Yij = Mu(i) + e(ij)
Var{e(ij)} = SigmaA2
Model A2: Yij = Mu(i) + e(ij)
Var{e(ij)} = Sigma(i)^2
Model A3: Yij = Mu(i) + e(ij)
Var{e(ij)} = exp(lalpha + log(mean(i)) * rho)
Model R: Yij = Mu + e(i)
Var{e(ij)} = Sigma/x2
Model
Likelihoods of Interest
Log(likelihood) DF
AIC
Al
A2
A3
R
2
3
29.04002
44.97661
43.40484
-80.9645
-15. 64228
-15.64228
7
12
8
2
4
4
-44.08005
-65.95321
-70.80967
165.929
39.28456
39.28456
Perfluorooctanoic Acid - February 2014
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A-21
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42.25345
42.2837
-74.50691
-72.56741
Additive constant for all log-likelihoods = -102. This constant added
to the above values gives the log-likelihood including the term that does not
depend on the model parameters.
Explanation of Tests
Test 1: Does response and/or variances differ among Dose levels? (A2 vs. R)
Test 2: Are Variances Homogeneous? (A2 vs. Al)
Test 3: Are variances adeguately modeled? (A2 vs. A3)
Test 4: Does Model 2 fit the data? (A3 vs. 2)
Test 5a: Does Model 3 fit the data? (A3 vs 3)
Test 5b: Is Model 3 better than Model 2? (3 vs. 2)
Test 6a: Does Model 4 fit the data? (A3 vs 4)
Test 6b: Is Model 4 better than Model 2? (4 vs. 2)
Test 7a: Does Model 5 fit the data? (A3 vs 5)
Test 7b: Is Model 5 better than Model 3? (5 vs. 3)
Test 7c: Is Model 5 better than Model 4? (5 vs. 4)
Tests of Interest
Test
Test 1
Test 2
Test 3
Test 4
Test 5a
Test 5b
Test 6a
Test 6b
Test 7a
Test 7b
Test 7c
-2*log(Likelihood Ratio) D. F.
251.9 10
31.87 5
3.144 4
118.1 4
118.1 4
-9.564e-012 0
2.303 3
115.8 1
2.242 2
115.9 2
0.0605 1
p-value
< 0.0001
< 0.0001
0.5341
< 0.0001
< 0.0001
N/A
0.512
< 0.0001
0.3259
< 0.0001
0.8057
The p-value for Test 1 is less than .05. There appears to be a difference
between response and/or variances among the dose levels, it seems appropriate
to model the data.
The p-value for Test 2 is less than .1.
appears to be appropriate.
A non-homogeneous variance model
The p-value for Test 3 is greater than .1. The modeled variance appears
to be appropriate here.
The p-value for Test 4 is less than .1. Model 2 may not adeguately
describe the data; you may want to consider another model.
The p-value for Test 5a is less than .1. Model 3 may not adeguately
describe the data; you may want to consider another model.
Degrees of freedom for Test 5b are less than or egual to 0. The Chi-Sguare
test for fit is not valid.
Perfluorooctanoic Acid - February 2014
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A-22
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The p-value for Test 6a is greater than .1. Model 4 seems to adequately
describe the data.
The p-value for Test 6b is less than .05. Model 4 appears to fit the data
better than Model 2.
The p-value for Test 7a is greater than .1. Model 5 seems to adequately
describe the data.
The p-value for Test 7b is less than .05. Model 5 appears to fit the data
better than Model 3.
The p-value for Test 7c is greater than .05. Model 5 does not seem to fit
the data better than Model 4.
Benchmark Dose Computations:
Specified Effect = 0.100000
Risk Type = Relative deviation
Confidence Level = 0.950000
Model
2
3
4
5
BMD and BMDL by Model
BMD
0.961795
0.961796
0.292244
0.312258
BMDL
0.832065
0.832065
0.246387
0.246874
Perfluorooctanoic Acid - February 2014
Draft - Do Not Cite or Quote
A-23
-------
Exponential Model 4 with 0.95 Confidence Level
c
o
Q.
(/)
0)
o:
c
(0
OJ
6.5
5.5
2.5
20
15:1210/252011
Perfluorooctanoic Acid - February 2014
Draft - Do Not Cite or Quote
A-24
-------
APPENDIX B
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
Perfluorooctanoic Acid - February 2014 B-l
Draft - Do Not Cite or Quote
-------
Interval
Variable
Conf. Limit
Background
*
Beta(l)
*
Beta(2)
Estimate
0.00409839
0.0116288
0
Std. Err.
95.0% Wald Confidence
Lower Conf. Limit Upper
* - Indicates that this value is not calculated.
Model
Full model
Fitted model
Reduced model
AIC:
Analysis of Deviance Table
Log(likelihood)
-28.6454
-29.3468
-34.0451
62.6936
# Param's Deviance Test d.f.
3
1.40286
10.7995
P-value
0.2362
0.004518
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
= I.-
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
interval for the BMD
% two-sided confidence
Multistage Cancer Slope Factor =
0.0200656
Perfluorooctanoic Acid - February 2014
Draft - Do Not Cite or Quote
B-2
-------
Multistage Cancer Model with 0.95 Confidence Level
•
I
C
o
13
ro
0.3
0.25
0.2
0.15
0.1
0.05
Multistage Cancer
Linear extrapolation
BMDL
BMD
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*doseAl) ]
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
Perfluorooctanoic Acid - February 2014
Draft - Do Not Cite or Quote
B-3
-------
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
Default Initial Parameter Values
Background = 0.0132945
Beta(l) = 0.0097738
Asymptotic Correlation Matrix of Parameter Estimates
Background Beta(l)
Background 1 -0.64
Beta(l) -0.64 1
Parameter Estimates
Interval
Variable
Conf. Limit
Background
*
Beta(l)
Estimate
0.00409839
0.0116288
Std. Err.
95.0% Wald Confidence
Lower Conf. Limit Upper
Indicates that this value is not calculated.
Model
Full model
Fitted model
Reduced model
AIC:
Analysis of Deviance Table
Log(likelihood)
-28.6454
-29.3468
-34.0451
62.6936
# Param's
3
2
1
Deviance Test d.f.
1.40286
10.7995
P-value
0.2362
0.004518
Dose
0.0000
1.3000
14.2000
Est. Prob.
0.0041
0.0190
0.1557
Goodness of Fit
Expected Observed Size
0.205
0.952
7.784
0.000
2.000
7.000
50
50
50
Scaled
Residual
-0.454
1.084
-0.306
= I.-
d.f. = 1
P-value = 0.2245
Benchmark Dose Computation
Perfluorooctanoic Acid - February 2014
Draft - Do Not Cite or Quote
B-4
-------
Specified effect = 0.04
Risk Type = Extra risk
Confidence level = 0.95
BMD = 3.51044
BMDL = 1.99346
BMDU = 8.7003
Taken together, (1.99346, 8.7003 ) is a 90 % two-sided confidence
interval for the BMD
Multistage Cancer Slope Factor = 0.0200657
Perfluorooctanoic Acid - February 2014 B-5
Draft - Do Not Cite or Quote
------- |