REVISED DRAFT HAZARD ASSESSMENT OF PERFLUOROOCTANOIC ACID AND ITS SALTS U. S. Environmental Protection Agency Office of Pollution Prevention and Toxics Risk Assessment Division November 4, 2002 PREFACE This is a preliminary assessment of the potential hazards to human health and the environment associated with exposure to perfluorooctanoic acid ( PFOA) and its salts. The majority of the toxicology information is for ammonium perfluorooctanoic acid ( APFO). This assessment includes a review of the studies that were available as of October, 2002. Table of Contents Executive Summary 1 1.0 Chemical Identity 7 1.1 Physicochemical Properties 7 2.0 Production of PFOA and its Salts 9 2.1 Uses of PFOA and its Salts 11 2.2 Environmental Fate 12 2.2.1 Photolysis 12 2.2.2 Volatility 12 2.2.3 Biodegradation 13 2.2.4 Hydrolysis 13 2.2.5 Bioaccumulation 14 2.2.6 Soil Adsorption 15 2.3 Environmental Exposure 15 2.3.1 Discharge to Air 15 2.3.2 Discharge to Water 15 2.3.3 Discharge to Land 16 2.3.4 Environmental Monitoring 16 2.4 Human Biomonitoring 18 3.0 Human Health Hazards 23 3.1. Metabolism and Pharmacokinetics 23 3.1.1 Half life in Humans 23 3.1.2 Absorption Studies in Animals 24 3.1.3 Distribution Studies in Animals 25 3.1.4 Metabolism Studies in Animals 28 3.1.5 Elimination Studies in Animals 28 3.2 Epidemiology Studies 32 3.2.1 Medical Surveillance Studies from Antwerp and Decatur Plants 32 3.2.2 Medical Surveillance Studies from Cottage Grove Plant 34 3.2.3 Mortality Studies 36 3.2.4 Hormone Study 39 3.2.5 Study on Episodes of Care ( Morbidity) 40 3.3 Acute Toxicity Studies in Animals 42 3.3.1 Oral Studies 42 3.3.2 Inhalation Studies 43 3.3.3 Dermal Studies 43 3.3.4 Eye Irritation Studies 43 3.3.5 Skin Irritation Studies 44 3.4 Mutagenicity Studies 44 3.5 Subchronic Toxicity Studies in Animals 44 3.6 Developmental Toxicity Studies in Animals 60 3.7 Reproductive Toxicity Studies in Animals 64 3.8 Carcinogenicity Studies in Animals 74 3.8.1 Cancer Bioassays 74 3.8.2 Mode of Action Studies 75 <IP_ADDRESS> Liver Tumors 76 <IP_ADDRESS> Leydig Cell Tumors 76 <IP_ADDRESS> Mammary Gland Tumors 77 <IP_ADDRESS> Pancreatic Tumors 77 3.9 Immunotoxicology Studies in Animals 78 4.0 Hazards to the Environment 80 4.1 Introduction 80 4.2 Acute Toxicity to Freshwater Species 81 5.0 References 87 ANNEX I   Robust Summaries 104 1 EXECUTIVE SUMMARY Introduction Perfluorooctanoic acid ( PFOA) and its salts are fully fluorinated organic compounds that can be produced synthetically or through the degradation or metabolism of other fluorochemical products. PFOA is primarily used as a reactive intermediate, while its salts are used as processing aids in the production of fluoropolymers and fluoroelastomers and in other surfactant uses. In recent years, less than 600 metric tons per year of PFOA and its salts have been manufactured in the United States or imported. Most of the toxicology studies have been conducted with the ammonium salt of perfluorooctanoic acid, which is referred to as APFO in this report.

Environmental Fate and Effects PFOA is persistent in the environment. It does not hydrolyze, photolyze or biodegrade under environmental conditions.

Groundwater samples taken near fire training areas that used fire fighting foams containing perfluorinated surfactants had elevated PFOA concentrations many years after the foam use. This demonstrates the following: ( 1) PFOA either existed in or was formed via degradation of the surfactants, ( 2) PFOA or its precursors migrate through the soil, and ( 3) PFOA persists in groundwater.

Several wildlife species have been sampled around the world to determine levels of PFOA. PFOA has rarely been found in fish sampled from the U. S., certain European countries, the North Pacific Ocean and Antarctic locations, or in fish eating bird samples collected from the U. S., including Midway atoll, the Baltic and Mediterranean Seas, and Japanese and Korean coasts. PFOA was found in a few mink livers from Massachusetts, but not found in mink from Louisiana, South Carolina and Illinois. PFOA concentrations in river otter livers from Washington and Oregon States were less than the quantification limit of 36 ng/ g, wet wt. PFOA was not detected at quantifiable concentrations in oysters collected in the Chesapeake Bay and Gulf of Mexico of the U. S. coast.

The concentrations of PFOA in surface water, sediments, clams, and fish collected from locations upstream and downstream of the 3M manufacturing facility at Decatur AL have been determined. Of the three downstream water and sediment sampling locations, the two closest to the facility had PFOA surface water concentrations significantly greater than the two upstream sites; the three downstream locations also had sediment concentrations significantly greater than the upstream sites. The small sample size prevented determination of significance for fish whole body PFOA concentrations. The average PFOA concentration in clams was not significantly different between the upstream and downstream locations.

Based on available laboratory data, APFO does not appear to bioaccumulate in fish. In a study of fathead minnows, the calculated BCF for APFO was 1.8. In a study of carp, the BCF ranged 2 from 3.1 to 9.1.

Several species were tested to assess the acute toxicity of APFO; these included the fathead minnow ( Pimephales promelas), bluegill sunfish ( Lepomis machrochirus), water flea ( Daphnia magna), and a green algae ( Selenastrum capricornutum). Comparisons of the different studies are problematic for several reasons. The studies were conducted with different test substances. Generally the ammonium salt or the tetrabutylammonium salt was tested. Purity of the test material is a major concern and was not sufficiently characterized in these tests. In some tests it appeared that 100% test chemical was used, for others a chemical of lesser purity ( approximately 27 to 85%) was used. Water, a solvent ( isopropanol) or a combination of both was used in other tests, for no obvious stated reason. Finally, only nominal test chemical concentrations were reported; the actual concentrations were not reported.

Twelve tests were conducted with fathead minnows; 96 h LC50 values ( based on mortality) ranged from 70 to 843 mg/ L. It is unclear why this range is so wide. Assuming these studies are valid, and due to the limitations discussed above, these toxicity values indicate low toxicity. The two acute values for bluegill sunfish also indicate low toxicity ( 96 h LC50s of > 420, and 569 mg/ L).

Nine acute tests were conducted with daphnids and 48 h EC50 values ( based on immobilization) ranged from 39 to > 1000 mg/ L. The lower values are indicative of moderate toxicity, but the wide range makes interpretation difficult.

Seven tests were conducted with green algae; 96 h EC50 values ( based on growth rate, cell density, cell counts, and dry weights) ranged from 1.2 to > 666 mg/ L ( the EC50 cell density value of 1,000 mg/ L is excluded from this discussion). The lower value indicates high to moderate toxicity, based on the acute criteria. The lower value would also be indicative of moderate toxicity, based on the chronic moderate criterion ( 0.1< 10 mg/ L). A 14 d EC50 value of 43 mg/ L, based on cell counts, for green algae was also calculated in one study. This is indicative of low chronic toxicity, based on the chronic criterion ( 10 mg/ L). Green algae appeared to be the most sensitive test species in the 44% APFO test sample, daphnids were the next most sensitive, and fathead minnows were the least sensitive.

Human Health Effects and Biomonitoring Little information is available concerning the pharmacokinetics of APFO in humans. An ongoing 5 year, half life study in 9 retired workers has suggested a mean serum PFOA half life of 4.37 years ( range, 1.50   13.49 years). These data provide evidence of the potential to bioaccumulate PFOA in humans.

Animal studies have shown that APFO is well absorbed following oral and inhalation exposure, and to a lesser extent following dermal exposure. In the past, Chemolite workers have been exposed to large dermal doses of PFOA. It appears that dermal exposure may have played a significant role in the absorption of PFOA in these workers. Upon recognition that PFOA could 3 be absorbed dermally, work practices were changed and engineering controls were adopted that reduced dermal exposures.

PFOA distributes primarily to the liver, plasma, and kidney, and to a lesser extent, other tissues of the body including the testis and ovary. It does not partition to the lipid fraction or adipose tissue. PFOA binds to macro molecules in the tissues listed above. PFOA is not metabolized and there is evidence of enterohepatic circulation of the compound. The urine is the major route of excretion of PFOA in the female rat, while the urine and feces are both major routes of excretion in male rats. There are major gender differences in the elimination of PFOA in rats. In female rats, the half life is 24 h in the serum and 60 h in the liver; in male rats, the half life is 105 h in the serum and 210 h in the liver. The rapid excretion of PFOA by female rats is due to active renal tubular secretion ( organic acid transport system); this renal tubular secretion is believed to be hormonally controlled, since castrated male rats treated with estradiol have excretion rates of PFOA similar to those of female rats. Hormonal changes during pregnancy do not appear to change the rate of elimination in rats. This gender difference has not been observed in primates and humans.

There are limited data on PFOA serum levels in workers and the general population. Occupational data from plants in the U. S. and Belgium that manufacture or use PFOA indicate that the most recent mean serum levels in workers range from 0.84 to 6.4 ppm. The highest level reported in a worker in 1997 was 81.3 ppm. In non occupational populations, serum PFOA levels were much lower. In both pooled blood bank samples and in individual samples, mean serum PFOA levels ranged from 3 to 17 ppb. The highest serum PFOA levels were reported in a sample of children from different geographic regions in the U. S. ( range, 1.9   56.1 ppb).

Epidemiological studies on the effects of PFOA in humans have been conducted on workers. Two mortality studies, a morbidity study, and studies examining effects on the liver, pancreas, endocrine system, and lipid metabolism, have been conducted to date. In addition, a crosssectional as well as a longitudinal study of the worker surveillance data have recently been submitted. However, these latest 2 studies focus primarily on PFOS rather than PFOA, even though recent PFOA levels are similar to or higher than PFOS levels in workers at these plants. ( It should be noted that PFOS levels in the sampled general population are much higher than PFOA levels).

A retrospective cohort mortality study demonstrated a weak, although statistically significant association between prostate cancer mortality and employment duration in the chemical facility of a plant that manufactures PFOA. However, in a recent update to this study in which more specific exposure measures were used, a significant association for prostate cancer was not observed. In a morbidity study, workers with the highest PFOA exposures for the longest durations sought care more often for prostate cancer treatment than workers with lower exposures.

Another study reported an increase in estradiol levels in workers with the highest PFOA serum levels; however, none of the other hormone levels analyzed indicated any adverse effects. Some 4 of the same employees who participated in the hormone study also were included in a study of cholecystokinin ( CCK) levels in employees. No positive association was noted between CCK values and PFOA. The other available study examined cholesterol and other serum components in workers. There did not appear to be any significant differences among workers of different exposure levels. At plants where the serum PFOA levels were lower, cross sectional and longitudinal studies found positive significant associations between PFOA and cholesterol and triglyceride levels. In addition, a positive, significant association was reported between PFOA and T3 hormone and a negative association with HDL in the cross sectional study. There are many limitations to these studies, and therefore, all of these results must be interpreted carefully.

In acute toxicity studies in animals, the oral LD50 values for CD rats were > 500 mg/ kg for males and 250 500 mg/ kg for females, and < 1000 mg/ kg for male and female Wistar rats. There was no mortality following inhalation exposure of 18.6 mg/ L for one hour in rats. The dermal LD50 in rabbits was determined to be greater than 2000 mg/ kg. APFO is a primary ocular irritant in rabbits, while the data regarding potential skin irritancy are conflicting.

APFO is not mutagenic. APFO did not induce mutation in either S. typhimurium or E. coli when tested either with or without mammalian activation. APFO did not induce gene mutation when tested with or without metabolic activation in the K 1 line of Chinese hamster ovary ( CHO) cells in culture. APFO did not induce chromosomal aberrations in human lymphocytes when tested with and without metabolic activation up to cytotoxic concentrations. APFO was tested twice for its ability to induce chromosomal aberrations in CHO cells. In the first assay, APFO 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, APFO induced significant increases in chromosomal aberrations and in polyploidy. APFO was negative in a cell transformation assay in C3H 10T ½ mouse embryo fibroblasts and in the mouse micronucleus assay.

Subchronic studies in rats and mice with 28 and 90 days of exposure have demonstrated that the liver is the primary target organ. In rat studies, males are far more sensitive than females. Dietary exposure to APFO for 90 days resulted in significant increases in liver weight and hepatocellular hypertrophy in female rats at 1000 ppm ( 76.5 mg/ kg/ day) and in male rats at doses as low as 100 ppm ( 5 mg/ kg/ day). Analyses of serum and liver levels of PFOA showed a marked gender difference that accounts for the difference in sensitivity. Chronic dietary exposure of rats to 300 ppm APFO ( 14.2 and 16.1 mg/ kg/ day for males and females, respectively) for 2 years resulted in increased liver and kidney weights, hematological effects and liver lesions in males and females. In addition, testicular masses were observed in males at 300 ppm and ovarian tubular hyperplasia was observed in females after exposure to 30 ppm ( 1.6 mg/ kg/ day), the lowest dose tested.

In a 90 day study with rhesus monkeys, exposure to doses of 30 mg/ kg/ day or higher resulted in death, lipid depletion in the adrenals, hypocellularity of the bone marrow, and moderate atrophy of the lymphoid follicles in the spleen and lymph nodes. Unlike rodent studies, analyses of the 5 serum and liver levels did not reveal a gender difference in monkeys, but the sample size was very small ( N= 2). In a 6 month study of male cynomolgus monkeys, dosing of animals in the 30 mg/ kg/ day dose group was stopped from days 11   21 because of toxicity. When dosing was resumed on day 22, animals received 20 mg/ kg/ day and this group was designated the 30/ 20 mg/ kg/ day group. This treatment was also not tolerated and treatment was stopped for 3/ 6 monkeys. Mortality was observed in one monkey at 3 mg/ kg/ day and at 30/ 20 mg/ kg/ day. There were no consistent effects on hormone levels. Increased absolute and relative liver weights were noted at 3, 10 and 30/ 20 mg/ kg/ day. While there was no evidence of peroxisome proliferation, there was evidence of mitochondrial proliferation suggesting a different mode of action than observed in rats. The serum levels were highly variable and should be treated with caution. On day 9 of treatment, the serum levels were 126 ± 36.1 µ g/ mL in the 3 mg/ kg/ day group and 1597 ± 2392 µ g/ mL in the 30/ 20 mg/ kg/ day group, and during weeks 26/ 27, the serum levels were 52.5 ± 9.14 µ g/ mL in the 3 mg/ kg/ day group and 51.5 ± 77.6 µ g/ mL in the 30/ 20 mg/ kg/ day group. The LOAEL for this study was 3 mg/ kg/ day and a NOAEL was not established.

PFOA is immunotoxic in mice. Feeding C57Bl/ 6 mice a diet containing 0.02% PFOA resulted in adverse effects to both the thymus and spleen. In addition, this feeding regimen resulted in suppression of the specific humoral immune response to horse red blood cells, and suppression of splenic lymphocyte proliferation in response to LPS and ConA. The suppressed mice recovered their ability to generate a humoral immune response when they were fed a diet devoid of PFOA. Studies using transgenic mice showed that the peroxisome proliferator activated receptor alpha was involved in causing the adverse effects to the immune system.

Prenatal developmental toxicity studies in rats resulted in death and reduced body weight in dams exposed to oral doses of 100 mg/ kg/ day or by inhalation to 25 mg/ m3 APFO. There was no evidence of developmental toxicity after oral exposure to doses as high as 150 mg/ kg/ day, while inhalation exposure to 25 mg/ m3 resulted in reduced fetal body weights. In a rabbit oral developmental toxicity study there was a significant increase in skeletal variations after exposure to 50 mg/ kg/ day APFO. There was no evidence of maternal toxicity at 50 mg/ kg/ day, the highest dose tested.

In a two generation reproductive toxicity study in rats exposed to 0, 1, 3, 10, and 30 mg/ kg/ day APFO, significant increases in absolute and relative liver and kidney weights were observed in F0 males at 1 mg/ kg/ day, while significant reductions in absolute and relative kidney weights were observed in F0 females at 30 mg/ kg/ day. Reproductive indices were not affected in the F0 animals. Serum levels of the 10 and 30 mg/ kg/ day groups were measured for F0 males after mating and F0 females at weaning of the F1 pups. In F0 males, the serum levels were 51.1+ 9.30 and 45.3+ 12.6 ug/ l, respectively for the 10 and 30 mg/ kg/ day groups, and in F0 females, the serum levels were 0.37+ 0.0805 and 1.02+ 0.425 ug/ l, respectively for the 10 and 30 mg/ kg/ day groups. In F1 females, there was a significant increase in post weaning mortality, a significant decrease in mean body weight, and a significant delay in sexual maturation at 30 mg/ kg/ day. In F1 males, significant decreases in body weights and body weight gains, and significant changes in absolute liver and spleen weights and in the ratios of liver, kidney, and spleen weights to 6 brain weights were observed in all treated groups. The increase in post weaning mortality and the delay in sexual maturation was also noted in F1 males at 30 mg/ kg/ day. Reproductive indices were not affected in the F1 animals. The LOAEL for the F1 females was 30 mg/ kg/ day, and the NOAEL was 10 mg/ kg/ day; the LOAEL for F1 males was 1 mg/ kg/ day and a NOAEL was not determined. The difference in sensitivity is presumed to be related to the gender difference in elimination of APFO. No treatment related effects were observed in the F2 generation. However, it should be noted that the F2 pups were sacrificed at weaning, and thus it was not possible to ascertain the potential post weaning effects that were noted in the F1 generation.

Carcinogenicity studies in Sprague Dawley ( CD) rats show that APFO is weakly carcinogenic, inducing Leydig cell adenomas in the male rats and mammary fibroadenomas in the females following dietary exposure to 300 ppm for 2 years ( equivalent to 14.2 mg/ kg/ day in males and 16.1 mg/ kg/ day in females). APFO has also been reported to be carcinogenic toward the liver and pancreas of male CD rats at 300 ppm.

The mechanism( s) of APFO tumorigenesis is not clearly understood. Available data appear to indicate that the induction of tumors by APFO is due to a non genotoxic mechanism, involving activation of receptors and perturbations of the endocrine system. There is sufficient evidence to suggest that APFO is a PPAR  agonist and that the liver carcinogenicity/ toxicity of APFO is mediated by binding to PPAR  in the liver. Recently, IARC ( 1995) concluded that the liver tumors induced in rodents by PPAR  agonists are unlikely to be operative in humans based on our current understanding of the animal mode of action. The Agency is currently examining the scientific knowledge associated with PPAR  agonist   induced liver tumors in rodents and the relevance to humans. Available data suggest that the induction of Leydig cell tumors ( LCT) and mammary gland neoplasms by APFO may be due to hormonal imbalance resulting from activation of the PPAR  and induction of the cytochrome P450 enzyme, aromatase. Preliminary data suggest that the pancreatic acinar cell tumors are related to an increase in serum level of the growth factor, cholecystokinin. 7 1.0 Chemical Identity Chemical Name: Perfluorooctanoic Acid Molecular formula: C8 H F15 O2 Structural formula: F CF2 CF2 CF2 CF2 CF2 CF2 CF2 C(= O) X,

The free acid and some common derivatives have the following CAS numbers: The perfluorooctanoate anion does not have a specific CAS number.

Free Acid ( X = OM+; M = H) [ 335 67 1]

Ammonium Salt ( X = OM+; M = NH4) [ 3825 26 1] Sodium Salt ( X = OM+; M = Na) [ 335 95 5] Potassium Salt ( X = OM+; M = K) [ 2395 00 8] Silver Salt ( X = OM+; M = Ag) [ 335 93 3]

Acid Fluoride ( X = F) [ 335 66 0]

Methyl Ester ( X = CH3) [ 376 27 2] Ethyl Ester ( X = CH2 CH3) [ 3108 24 5]

Synonyms: 1 Octanoic acid, 2,2,3,3,4,4,5,5,6,6,7,7,8,8,8 pentadecafluoro PFOA 1.1 Physicochemical Properties For this report, perfluorooctanoic acid is consistently referred to as PFOA. Most of the toxicology studies have been conducted with the ammonium salt of perfluorooctanoic acid, which will be referred to as APFO in this report. PFOA is a completely fluorinated organic acid. The typical structure has a linear chain of eight carbon atoms. The physical chemical properties noted below are for the free acid, unless otherwise stated. The data for the free acid, pentadecafluorooctanoic acid [ 335 67 1], is the most complete. The reported vapor pressure of 10 mm Hg appears high for a low melting solid when compared to other low melting solids ( chloroacetic acid: solid; MP = 61 to 63 ° C; BP = 189 ° C; VP = 0.1 kPa ( 0.75 mm Hg) @ 20 ° C; NIOSH), but is consistent with other perfluorinated compounds with similar boiling points ( perfluorobutanoic acid BP = 120 ° C, VP 10 mm Hg @ 20 ° C; Beilstein IV 2 p. 811). Another explanation may be that the 10 mm vapor pressure was measured at an elevated temperature ( but the temperature inadvertently omitted), as perfluorooctanoic acid is typically handled as a liquid at 65 ° C ( 3M data sheet for FC 26).

The free acid is expected to completely dissociate in water, leaving the anionic carboxylate in the water and the perfluoroalkyl chain on the surface. In aqueous solutions, individual molecules of PFOA anion loosely associate on the water surface and partition between the air / water 8 interface. Several reports note that PFOA salts self associate at the surface, but with agitation they disperse and micelles form at higher concentrations. ( Simister, 1992; Calfours, 1985; Edwards, 1997). Water solubility has been reported for PFOA, but it is unclear whether these values are for a microdispersion of micelles, rather than true solubility. Due to these same surface active properties of PFOA, and the test protocol for the OECD shake flask method, PFOA is anticipated to form multiple layers in octanol/ water, much like those observed for PFOS. Therefore, an n octanol/ water partition coefficient cannot be determined.

The available physicochemical properties for the PFOA free acid are:

MW: 414 ( Beilstein, 1975) MP: 45   5 ° C ( Beilstein, 1975) BP: 189   192 ° C / 736 mm Hg ( Beilstein, 1975) VP: 10 mm Hg @ 25 ° C ( approx.) ( Exfluor MSDS) Sol.  Water: 3.4 g/ L ( telomeric [ MP = 34 ° C ref. 0.01  0.02 mol/ L ~ 4  8 g/ L) ( MSDS from Merck, Fischer, and Chinameilan Internet sites) pKa: 2.5 ( USEPA AR226 0473) pH ( 1g/ L): 2.6 ( MSDS Merck)

The PFOA derivative of greatest concern and most wide spread use is the ammonium salt ( APFO), [ 3825 26 1]). This substance was the testing substance for the toxicity testing performed and has the P Chem data provided in the Table below.

The water solubility of APFO has been inconsistently reported. One 3M study reported the water solubility of APFO to be > 10%. It was noted in an earlier study that at concentrations of 20 g/ L, the solution " gelled" ( 3M, 1979). These numbers seem surprising low for a salt , in light of Apollo Scientific selling a 31% aq. solution of APFO. One author reported the APFO partition coefficient log Pow = 5. Another author reported an estimated APFO log Pow =  0.9. This value might not be accurate due to the estimation method used ( Hansch and Leo 1979). Again, the anticipated formation of an emulsified layer between the octanol and water surface interface would make determination of log Kow impossible.

Determination of the vapor pressure of APFO is complicated. APFO had recently reported a vapor pressure of 7 x 10 5 mm Hg at 20E C, which seems too low for a material that sublimates as the ammonium salt. ( 3M Environmental Laboratory, 1993). The ammonium salt begins to sublimate at 130 ° C ( USEPA AR 226 473). As the temperature increases from when APFO begins to sublimate, 20% of the sample weight is lost by 169 ° C. Other salts ( Cs, K, Ag, Pb, Li) do not demonstrate similar weight loss until 237 ° C or higher. ( Lines, 1984). Decomposition of different salts produces perfluoroheptene ( loss of metal fluoride and carbon dioxide). This occurs at 320 ° C for the sodium salt and at 250 290 ° C for the silver salt ( Beilstein 1975).

The physicochemical properties of PFOA and its common derivatives are summarized in Table 1. 9 Table 1. Reported Physicochemical Properties Compound CAS REG # MP BP VP Sol. H2O Log P* Rf C(= O) F 335 64 8 131 ° C Rf CO2H 335 67 1 55 ° C 189 ° C 10 mm Hg 3.4 g/ L Rf CO2 NH4+ 3825 26 1 130 ° C ( sub) sublimes 1 x 10E 5 mm Hg 20 g/ L gels Rf C(= O) OMe 376 27 2 159 ° C pH ( 1 g free acid / L Water) = 1.5   2.5 Free acid pKa is approximately 0.6 Sodium or Silver salts of PFOA decompose above 250 ° C to generate perfluoroolefins.   Surfactants traditionally emulsify octanol and water 2.0 Production of PFOA and its Salts PFOA has been commercially manufactured by two major alternative processes: 1) the Simons Electro Chemical Fluorination ( ECF) process or 2) the telomerization process. The 3M Company was reported to be largest manufacturer and importer of PFOA and its salts in the United States in 1999. They predominantly used the ECF process to produce a PFOA precursor which is ultimately converted to PFOA. Data for the FC 26 ( 97% C8) suggests this purified material was about 80% linear.

In the ECF process to make PFOA, an electric current is passed through a solution of anhydrous hydrogen fluoride and an organic feedstock, typically an octanoic acid derivative. The ECF process replaces the carbon hydrogen bonds on molecules of the organic feedstock with carbonfluorine bonds, in an identical manner used to make PFOS. Perfluorination occurs when all the carbon hydrogen bonds are replaced with carbon fluorine bonds. The ECF process yields between 30 45 percent straight chain ( normal) perfluorooctanoyl fluoride ( PFOF), along with a variable mixture of byproducts and impurities. The product from the ECF process is not a pure chemical, but instead a mixture of isomers and homologs including higher and lower straightchain homologs; branched chain perfluoroalkyl fluorides of various chain lengths; straight chain, branched, and cyclic perfluroalkanes and ethers; and other byproducts ( 3M Company, 2000a). After separation and recovery of desired material from the byproducts and impurities of the crude reaction mixture, the initially formed acid fluoride is base hydrolyzed in batch reactors and acidified to ultimately yield PFOA. The PFOA salts are synthesized by base neutralization of the acid to the salt in a separate reactor ( 3M Company, 2000b).

3M has characterized its manufacture of PFOA and its derivatives in 1997 at less than 500,000 kg per year in the US, and its importation at less than 100,000 kg ( 3M Company, 2000a). ( These figures may overstate the total production volume of PFOA since the vast majority of PFOA is consumed in the manufacture of the ammonium or other metallic salts.) Industry sources have characterized 3M as the dominant global producer of PFOA related chemicals, manufacturing approximately 85 percent or more of total worldwide volumes of the ammonium salt of PFOA 10 ( FMG, 2001). USEPA has not located information that would contradict this claim. More precise production volumes of PFOA and the ammonium and sodium salts have been reported to USEPA by 3M, but have been claimed as TSCA confidential business information, preventing disclosure in this report. Since 1985, USEPA has received a total of approximately 25 notifications for PFOA related chemicals that were not previously on the TSCA Chemical Inventory. Most of these notifications were from companies other than 3M. In most cases, the notifications qualified for the Low Volume Exemption for new chemicals with a production volume less than 10 metric tons per year.

Current production volume information for manufacturers other than 3M has not been provided by industry, nor is it available in USEPA's Chemical Update System ( which contains information on non polymeric organic chemicals manufactured in the United States or imported in volumes above 4,525 kg). Furthermore, there is no information on the total cumulative production volumes of PFOA since initial commercialization. In terms of on going production, in comments to the draft hazard assessment of PFOA 3M has stated that their May 16, 2000 announcement of the phase out of the production of perfluorooctanyl chemistry and related products includes PFOA and its salts. This commits 3M to a complete phase out of PFOA and PFOA related chemicals identical to the phase out of PFOS and PFOS related chemicals by the end of 2002. In addition, the Fluoropolymer Manufacturers Group ( 3M largest customer for PFOA products, specifically the ammonium salt, APFO) has informed the EPA that it will be manufacturing 300,000 kg/ yr to meet the industry demand for APFO that 3M has not met ( FMG 2002 AR226 1094). The FMG manufacturing process is based on telomerization.

The telomerization process, according to 3M's Bultman and Pike, " is the reaction of a telogen ( such as pentafluoroethyl iodide...) with a polymerizable ethylenic compound ( such as tetrafluoroethylene) to form ` telomers'. In this process the resultant telomer iodides are then reacted with ethylene, via free radical addition. This forms a mixture of iodides that can be reacted subsequently to form a variety of useful materials....". Dupont uses an analogous telomerization process. Either process yields predominantly pure straight chain acids with an even number of carbon atoms. Commercial products manufactured through the telomerization process are generally mixtures of perfluorinated compounds with ranges of even carbon numbers ( Renner, 2001). Distillation can be used to obtain pure components ( ECT, 1994).

Aside from the United States, OECD Member countries that reportedly have production capacity include France, Germany, Italy, and Japan. There may also be some production in non OECD countries such as China. Following are companies that may manufacture PFOA and its salts ( 3M Company, 2000b; Directory of World Chemical Producers, 1998; Dynax, 2000; Renner, 2001; SEMI, 2001):

OECD ° 3M Company ( United States) ° DuPont ( United States) ° Exfluor Research Corporation ( United States) 11 ° PCR Inc. ( United States) ° Ciba Specialty Chemicals ( Germany) ° Clariant ( Germany) ° Dyneon ( Germany) ° Hoechst Aktiengesellschaft ( Germany) ° EniChem Synthesis S. p. A. ( Italy) ° Miteni S. p. A. ( Italy) ° Asahi Glass ( Japan) ° Daikin ( Japan) ° Dainippon ( Japan) ° Tohkem Products Corporation ( Japan)

Non OECD ° Chenguang Research Institute of the Chemical Industry ( China) ° Shanhai 3F New Materials Co., Ltd. ( China)

2.1 Uses of PFOA and its Salts PFOA is used mainly as a chemical intermediate, and its salts are used in emulsifier and surfactant applications.

According to 3M, the vast majority of PFOA is consumed to make the ammonium or sodium salts. 3M also uses PFOA as a reactive intermediate in the industrial synthesis of a fluoroacrylic ester. The fluoroacrylic ester is used in an industrial coating application ( 3M Company, 2000a).

The salts of PFOA have additional uses, mostly in surfactant and emulsifier applications. These include the following:

Processing aid in the industrial synthesis of fluoropolymers and fluoroelastomers such as polytetrafluoroethylene and polyvinylidene fluoride with a variety of industrial and consumer uses ( 3M Company, 2000a; DuPont, 2000; Daikin, 2001).

Post polymerization processing aids in the stabilization of suspensions of fluoropolymers and fluoroelastomers prior to further industrial processing ( 3M Company, 2000a).

Processing aid for factory applied fluoropolymer coatings on fabrics, metal surfaces, and fabricated or molded parts ( 3M Company, 2000a).

Extraction agent in ion pair reversed phased liquid chromatography ( Petritis, 1999).

Based on the physicochemical properties of the salts of PFOA, they may also have other related surfactant or emulsifier uses as a photographic chemical or in the manufacture of electronic components such as semiconductors. These same properties may lead industry to explore PFOA 12 as a replacement chemical for PFOS in other applications in which PFOA is not currently used.

2.2 Environmental Fate 2.2.1 Photolysis Direct photolysis of APFO was examined in two separate studies ( Todd, 1979; Hatfield, 2001) and photodegradation was not observed in either study. In the Todd ( 1979) study, a solution of 50 mg/ l APFO in 2.8 liters of distilled water was exposed to simulated sunlight at 22 ± 2 º C. Spectral energy was characterized from 290 600 nm with a max output at ~ 360 nm. Direct photolysis of the test substance was not detected. However, the author noted that sample purity was not properly characterized which may have contributed to experimental error.

In the Hatfield ( 2001) study, both direct and indirect photolysis were examined utilizing techniques based on EPA and OECD guidance documents. To determine the potential for direct photolysis, APFO was dissolved in pH 7 buffered water and exposed to simulated sunlight ( Scrano, 1999; Nubbe, 1995). For indirect photolysis, APFO was dissolved in 3 separate matrices and exposed to simulated sunlight for periods of time from 69.5 to 164 hours. These exposures tested how each matrix would affect the photodegradation of APFO. One matrix was a pH 7 buffered aqueous solution containing H2O2 as a well characterized source of OH radicals ( Ogata, 1983; Lunak, 1992). This tested the propensity of APFO to undergo indirect photolysis. The second matrix contained Fe2O3 in water that has been shown to generate hydroxyl radicals via a Fenton type reaction in the presence of natural and artificial sunlight ( Kachanova, 1973; Behar, 1966). The third matrix contained a standard solution of humic material. Neither direct nor indirect photolysis of APFO was observed based on loss of starting material. Predicted degradation products were not detected above their limits of quantitation. There was no conclusive evidence of direct or indirect photolysis whose rates of degradation are highly dependent on the experimental conditions. Using the iron oxide ( Fe2O3) photoinitiator matrix model, the APFO half life was estimated to be greater than 349 days.

2.2.2 Volatility Impinger studies were performed to examine the volatility of APFO and PFOS. Solutions of APFO or PFOS containing ammonium acetate in water/ 1 propanol ( 50: 50) or phase transfer agents, e. g., n alkyldimethylbenzylammonium chloride ( 3M Environmental Laboratory, 1993) were blown with 280 liters of air at a flow rate of 1 L/ min. ( 3M Environmental Laboratory, 1993). The results indicate there is some loss of APFO and PFOS, but most of the solutions retained over 80% or more of the fluorochemicals. The average retention was 92% for both APFO and PFOS. This indicates that there is loss from the solutions. However, some of the solutions, particularly the n alkyldimethylbenzylammonium chloride solution, appear to retain all the fluorochemicals. These results were reviewed by Dr. Edwin Tucker of the Chemistry Dept. at the University of Oklahoma ( 3M Environmental Laboratory, 1993). He concluded that it is very unlikely that these fluorochemicals were removed by bubbling air through water due to 13 their very low vapor pressures. He suggested that a more plausible mechanism for loss from the solution phase is concentration of the surfactants in foam and loss from the bubbled solutions as foam or micro droplets.

In the second part of the experiment, air was passed over the fluorochemicals and bubbled through a train of impingers containing the ammonium acetate solution. It was expected that if any fluorochemicals were present in the air they would be transferred and retained by the ammonium acetate solution. However, no fluorochemicals were present in either the first or second impinger. The report concludes that the vapor pressure of both compounds is less than 10 E 07 mm Hg.

According to these experiments, APFO and PFOS ( potassium salt) have very low volatility and vapor pressure. Quantitative conclusions regarding rates of volatilization from water or Henry's Law constant are not possible. However, APFO and PFOS are capable of transport out of water. Also, the loss of the fluorochemicals may have been as the free acids, not the salt forms. APFO sublimes at 130 C ( see Physicochemical Properties Section 1.1). There is no information on the validity of the test method for determining volatility of the test substance. The study also lacks characterization of the purity of the test substance.

2.2.3 Biodegradation Using an acclimated sludge inoculum, the biodegradation of APFO was investigated using a shake culture study modeled after the Soap and Detergent Association's presumptive test for degradation ( Reiner, 1978). Both thin layer and liquid chromatography did not detect the presence of any metabolic products over the course of 2 1/ 2 months indicating that PFOA does not readily undergo biodegradation. In a related study, 2.645 mg/ L APFO was not measurably degraded in activated sludge inoculum ( Pace Analytical, 2001). Test flasks were prepared using a mineral salts medium, 1 mL methanol, and 50 mL settled sludge. Analysis was conducted with a HPLC/ MSD system. Several other studies conducted between 1977 1987 also did not observe APFO biodegradation using what probably were standard COD and BOD methods, however, the methods used in these studies were either insufficiently described ( i. e. no description of experimental protocols) or there were indications of a high degree of experimental error. The results were, therefore, deemed unreliable by the submitter ( 3M Company, 1977; 3M Company, 1980; 3M Company, 1985b; Pace Analytical, 1997).

2.2.4 Hydrolysis The 3M Environmental Laboratory ( 2001a) performed a study of the hydrolysis of PFOA. The study procedures were based on EPA's OPPTS Guideline Document 835.2110 ( EPA 1998); although the procedures do not fulfill all the requirements of the guideline, they were more than adequate for these studies. Results were based on the observed concentrations of PFOA in buffered aqueous solutions as a function of time. The chosen analytical technique was high performance liquid chromatography with mass spectrometry detection ( HPLC/ MS). 14 During the study, samples were prepared and examined at six different pH levels from 1.5 to 11.0 over a period of 109 days. Experiments were performed at 50 ° C and the results extrapolated to 25 ° C. Data from two of the pH levels ( 3.0 and 11) failed to meet the data quality objective and were rejected. Also rejected were the data obtained for pH 1.5 because ion pairing led to artificially low concentrations for all the incubation periods. The results for the remaining pH levels ( 5.0, 7.0, and 9.0) indicated no clear dependence of the degradation rate of PFOA on pH. From the data pooled over the three pH levels, it was estimated that the hydrolytic half life of PFOA at 25 ° C is greater than 92 years, with the most likely value of 235 years. From the mean value and precision of PFOA concentrations, it was estimated the hydrolytic half life of PFOA to be greater than 97 years.

2.2.5 Bioaccumulation Three studies have been conducted to determine the potential for bioaccumulation of APFO. Howell et al. ( 1995) exposed Fathead minnows to 25 mg/ l APFO for 13 days. After 13 days exposure, the fish were then removed from APFO contaminated water and analyzed for depuration over 15 days. After 192 and 312 hours exposure to APFO contaminated water, the average concentration of APFO in fish tissue was 44.7 and 46.7 µ g/ g wet weight ( ww), respectively. At this point, APFO appeared to reach steady state. Twenty four hours after being transferred to clean water, the concentration of APFO decreased to 19.9 µ g/ g ww and by 96 hours post exposure, the concentration had decreased to approximately 8 µ g/ g ww and remained relatively constant until test termination at 360 hours. The calculated BCF for APFO was 1.8. It should be noted that questions have been raised about this study regarding the analytical techniques, high test chemical concentration, and short test duration.

In a bioaccumulation study of carp ( Cyprinus carpio), 28 carp were exposed to two concentrations of PFOA, 5 and 50 : g / L respectively for 28 days in a flow through system ( Kurume Laboratory, 2001). The chemical was prepared for exposure by using a dispersant made up of hydrogenated castor oil and mixed in acetone. The purity of PFOA was 100%. Water quality was monitored daily throughout the duration. The bioaccumulation potential of PFOA was found to be low. The carp exposed to 5 : g/ L resulted in a BCF of 3.1 while the carp exposed to 50 : g / L showed a BCF of < 5.1   9.1.

Vraspir ( 1979) conducted a study to determine if bluegill sunfish bioaccumulate fluorochemicals from the 3M Decatur plant. Two lots of 30 fish were used. One lot was exposed to Decatur plant effluent for 21 days and the other to river water only for 23 days. Exposed fish, both living and dead, as well as the control fish were homogenized and analyzed for fluorochemicals by GC, TLC, and GC/ MS. There were no detectable amounts of APFO in the ethyl acetate or toluene extracts of the tissues. No fluorochemicals were detected in the river water exposed fish. However, interpretation of this study is problematic for several reasons. Effluent concentrations of subject fluorochemicals were not characterized and the specific protocol for exposure of the fish was not found. There was also no information on analysis of the Tennessee River water or effluent used in the study. Additionally, it was not known if there was any opportunity for the depuration of the fish prior to sacrifice. No explanation was attempted as to what was the cause 15 of the twelve dead fish in the effluent exposed group. The study also did not differentiate between the bioaccumulation of the test compound and the sorption onto the surface of the fish.

2.2.6 Soil Adsorption The adsorption desorption of APFO was studied in 25 ml solutions of 14C labeled APFO in distilled water with 5 g Brill sandy loam soil for 24 hours at a temperature of 16 19 º C. The study reported a Kd of 0.21 and a Koc of 14 indicating that PFOA has high mobility in Brill sandy loam soil ( Welsh 1978). The Koc value, however, is questionable due to the lack of accurate information on the purity of the 14C labeled test substance ( Boyd 1993a, b).

Moody and Field ( 1999) conducted sampling and analysis of samples taken from groundwater 1 to 3 meters below the soil surface in close proximity to two fire training areas with a history of aqueous film forming foam use. Perfluorooctanoate was detected at maximum concentrations ranging from 116 to 6750 ug/ L at the two sites many years after its use at those sites had been discontinued. These results suggest that APFO may have the potential to migrate through soils to relatively shallow groundwater where it persists.

2.3 Environmental Exposure 2.3.1 Discharge to Air For 1997, 3M estimated 1950 lbs. of PFOA compound ( PFOA and related salts) stack releases at its Cottage Grove, MN location and another 4500 lbs. from Cottage Grove incinerated offsite ( 3M, 2000). In 1998, 70% of the fluoride containing wastes at 3M's Decatur location were incinerated off site; incineration is now the primary disposal method for these materials ( 3M, 2000). For 1999, DuPont estimated stack releases of 24,000 lbs. APFO at its Washington Works, WV location, plus another 16,000 lbs. from Washington Works incinerated offsite ( DuPont, 2000).

The above data may not account for all of the sources of air releases of PFOA. Other potential sources addressed in the literature include thermolysis of fluoropolymers ( Ellis et al, 2001); the low temperature of sublimation (~ 135C) that would yield PFOA from drying, aka sintering, of fluoropolymer made with PFOA as a processing aid; the thermally refractive character of the CF bond ( Napoli, et al, 1984; Taylor, et al, 1990; Tsang, et al, 1998) that would require very high temperature incineration as a means of destruction of this molecule.

2.3.2 Discharge to Water By analogy to PFOS, PFOA discharged to water may remain there, become adsorbed to particulate matter and sediment, and/ or be assimilated by organisms. For 1999, 3M estimated PFOA compound water releases of < 30,000 lbs. at its Decatur AL location, and < 15,000 lbs. at its Cottage Grove MN location ( 3M Company, 2000a, b). For 1999, DuPont estimated the following APFO water releases per location: Washington Works WV, 55,000 lbs; Parlin NJ, 300 16 lbs.; Spruance VA, 150 lbs.; Chambers Works NJ, 9500 lbs. ( DuPont, 2000).

DuPont measured an APFO concentration at its site Washington Works, WV site of 0.552 ug/ l from a 1999 drinking water sample obtained from GE Plastics immediately downstream on the Ohio River. Modeled 1996 APFO compound releases indicated an average annual PFOA concentration of 0.423 ug/ l, with APFO concentrations likely to exceed 1 ug C 8/ l about 50% of the time during the year, and likely to exceed 10 ug APFO/ L about 2.2% of the time during the year.

2.3.3 Discharge to Land 3M reported that land treatment of sludge from wastewater treatment at their Decatur, AL location ended in mid 1998; less than 500 lbs. were disposed to land at that site in 1997. Sludge from the Decatur site is now transported to an offsite landfill; sludge from 3M's Cottage Grove, MN facility is sent to an industrial landfill ( 3M Company, 2000a, b). DuPont ( 2000) estimated 3,900 lbs. of APFO sludge landfilled on site in 1999 at their Chambers Works, NJ facility. DuPont estimated 2,600 lbs. APFO transferred offsite to a hazardous waste landfill from their Washington Works, WV facility.

Prior operations resulted in ground and surface water concentrations of APFO monitored at three landfills operated by DuPont's Washington Works WV facility. Average surface water concentrations for two landfills were 1392 ug/ L and 18.5 ug/ L, respectively. A third landfill had a maximum concentration of 33 ug/ L in the permitted outfall. Average groundwater concentrations for two landfills were 2537 ug/ L and 8.83 ug/ L, respectively. A third landfill had a maximum groundwater concentration of 15 ug/ L ( DuPont, 2000).

DuPont also reported the following APFO concentrations, measured January 2000, in three drinking water wells of the Lubeck Public Service District, downstream of DuPont's Washington Works WV site: 0.8 ug/ L, 0.44 ug/ L, and 0.313 ug/ L ( DuPont, 2000). As of August 2000, the Lubeck Public Service District ( LPSD) reported APFO concentrations of 0.2 ppb in drinking water at DuPont's Washington Works facility, and 0.2, 0.5, and 0.1 ppb in the three LPSD wells ( LPSD, 2000).

2.3.4 Environmental Monitoring 3M's Multi City Study reported on PFOA concentrations from water, sludge, sediment, POTW effluent, and landfill leachate samples taken in six cities ( 3M, 2001a). Four of the cities ( Decatur, AL; Mobile, AL; Columbus, GA; Pensacola, FL) were " supply" cities that have manufacturing or industrial use of fluorochemicals; two of the cities ( Cleveland, TN; Port St. Lucie, FL) were " control" cities that do not have significant fluorochemical activities. Across all cities, POTW effluent concentrations ranged from 0.040 to 2.42 ppb. The POTW sludge ( dry wt.) range was non detect to 244 ppb; the drinking water range was non detect to 0.029 ppb; the landfill leachate range was non detect to 48.1 ppb; the surface water range was non detect to 0.083; the sediment range was non detect to 1.75 ppb ( dry wt.); and the quiet water range was 17 non detect to 0.097 ppb. The " control" cities samples generally inhabited the lower end of the above ranges, except for the POTW effluent and sludge findings for Cleveland, which were intermediate in their ranges.

The Multi City Study also included a market basket sampling of PFOA residue in a total of over 200 samples taken from green beans, apples, pork muscle, cow milk, chicken muscle, chicken eggs, bread, hot dogs, catfish, and ground beef ( 3M, 2001a). Measurable quantities of PFOA, ranging up to 2.35 ng/ g, were found in two ground beef samples ( control cities), two bread samples ( control and supply cities), two apple samples ( supply cities), and one green bean sample ( supply city).

Giesy reported that PFOA was rarely found in fish and fish eating water birds. Fish were sampled from the U. S., certain European countries, the North Pacific Ocean, and Antarctic locations ( Giesy, 2001a). Fish eating bird samples were collected from the U. S., including Midway atoll, the Baltic and Mediterranean Seas, Japanese and Korean coasts ( Giesy, 2001b)

Giesy reported on PFOA in mink and river otter livers from the U. S. ( Giesy, 2001c). PFOA was found in a few mink livers from Massachusetts at a concentration range of < 18 to 108 ng/ g, dry wt., but not found in mink from Louisiana, South Carolina and Illinois. PFOA concentrations in river otter livers from Washington and Oregon States were less than the quantification limit of 36 ng/ g, wet wt.

Giesy reported that PFOA was not detected at quantifiable concentrations in oysters collected in the Chesapeake Bay and Gulf of Mexico of the U. S. coast ( Giesy, 2001d).

Giesy reported on the concentrations of PFOA in surface water, sediments, clams, and fish collected from locations upstream and downstream of the 3M facility at Decatur, AL ( Giesy, 2001e). Of the three downstream water and sediment sampling locations, the two closest to the 3M facility had PFOA surface water concentrations significantly greater than the two upstream sites ( means of 1900ug/ L and 1024 ug/ L, vs. 0.008 ( est.) and 0.028 ug/ L); the three downstream locations also had sediment concentrations significantly greater than the upstream sites ( wet wt. means 1855 ug/ kg, 892 ug/ kg, 238 ug/ kg vs. 0.08( est.) and 0.09( est.)). The average fish whole body PFOA concentration for the upstream location was 11.7 ug/ kg ( wet wt.), while that for the downstream location was 106.4 ug/ kg; the small sample size prevented determination of significance. The average PFOA concentration in clams at the upstream location was 4.38 ug/ kg; that for the downstream location was 8.42 ug/ kg. These differences were not significant.

Hansen ( 2002) reported concentrations of PFOA measured from surface water samples taken from the Tennessee River up and downstream of the outfall from the fluorochemical manufacturing facility at Decatur AL ( the 3M facility mentioned above). There were 20 sampling sites above and 20 sites below the outfall location, spaced at approximately 2 mile intervals. None of the samples taken from upstream of the facility had measurable concentrations of PFOA. The downstream concentrations of PFOA were observed to increase at a point approximately six miles below the outfall; the average PFOA concentration from that 18 point downstream was 394 + 128 ng/ L. The report states that the consistency of the PFOA concentrations within these two regions suggests the absence of either major environmental sinks or additional sources of PFOA in the areas sampled.

Moody reported the concentrations of PFOA in surface water from stream locations sampled upstream and downstream of a spill of fire fighting foam that contained perfluorinated surfactants, including PFOA ( Moody et al., 2002). Upstream surface water samples taken over the three week period post spill had PFOA levels of 0.008 0.033 ug/ L; corresponding downstream levels ranged from 0.035 10.6 ug/ L. Statistical significance was not evaluated for these measurements.

2.4 Human Biomonitoring Table 2 provides serum PFOA levels in both occupational cohorts and in the general population. The highest levels reported to date in the general population are similar to some of the lowest levels in workers exposed to PFOA occupationally. The data are currently limited to those discussed below.

3M has been offering voluntary medical surveillance to workers at plants that produce or use perfluorinated compounds since 1976. Serum PFOA levels have been measured and reported since 1993. Prior to this time, only total organic fluorine was measured. The results of biomonitoring for PFOA have been reported for 3 plants: Cottage Grove, MN; Decatur, AL and Antwerp, Belgium. Surveillance years include 1993, 1995, 1997, 1998, and 2000, although not all of the plants offered surveillance in all of these years. The 1998 data reported for the Decatur plant consist of a random sample of employees; however, volunteers participated in all of the other sampling periods for all of the plants.

Mean serum PFOA levels have increased slightly at both the Cottage Grove and Decatur plants since 1993. Workers at the Cottage Grove plant, where PFOA exposures are highest, have the highest PFOA serum levels. The latest sample was in 1997 ( Olsen, et al., 1998b). The mean serum PFOA level was 6.4 ppm ( range = 0.1   81.3 ppm). Only 74 employees participated in the 1997 surveillance. The eligible voluntary participation rates ranged from approximately 50% in 1997 to 70% in 1993.

At the Decatur plant, 263 of 500 employees participated in 2000 ( Olsen, et al., 2001a). The mean serum PFOA level was 1.78 ppm. It was higher in males ( n = 215) than females ( n = 48), 1.90 and 1.23 ppm, respectively. In addition, male production employees had higher mean serum levels ( 2.34 ppm). Five employees had serum levels greater than 5 ppm, the Biological Limit Value established by the 3M Exposure Guideline Committee. Cell operators had the largest increase in serum PFOA between 1998 and 2000. The highest level was in a chemical operator on the Scotchgard team ( 12.70 ppm). The mean level for the rest of the members of the team was 5.06 ppm ( range 5  9 ppm). Other job categories did not exhibit such a large increase. 3M reports that this is due to increased PFOA production at the Decatur plant beginning in 1999. 19 Serum PFOA levels for the Antwerp plant are lower than at Decatur or Cottage Grove, and have decreased slightly since 1995 ( Olsen, et al., 2001b). Participation in medical surveillance at the Antwerp plant was the highest it had ever been in 2000 ( 258 volunteers out of 340 workers). The mean serum PFOA level was 0.84 ppm, and the highest serum level reported was 7.04 ppm. As in the Decatur plant, males ( n = 209) had higher mean serum PFOA levels ( 1.03 ppm) than females ( n = 49, 0.07 ppm). Three employees had levels greater than 5 ppm.

3M's Specialty Materials Manufacturing Division Laboratories, where employees perform fluorochemical research ( Building 236), conducted voluntary biomonitoring of 45 employees in 2000 ( Olsen, et al., 2001c). The mean PFOA serum level was 0.106 ppm ( range 0.008   0.668 ppm).

Serum PFOA levels in corporate staff and managers at a 3M plant in St. Paul, MN, where occupational exposure to PFOA should not have occurred, were reported ( 3M Report, 1999). Four of 31 employees had serum PFOA levels greater than the detection limit of 10 ppb. The mean for these employees was 12.5 ppb.

Data on PFOA levels in the general population are very limited. They are very recent so that trends over time cannot be established. The mean serum PFOA levels are lower in the general population than in workers exposed to PFOA.

Pooled blood samples from U. S. blood banks indicate mean PFOA levels of 3 to 17 ppb ( 3M Company, Feb. 5, 1999; 3M Company, May 26, 1999). The highest pooled sample reported was 22 ppb. Samples were collected in 1998 and 1999. However, it cannot be assumed that these levels are generalizable to the U. S. population for several reasons: 1) blood donors are a unique group that does not necessarily reflect the U. S. population as a whole, 2) many of the blood banks originally contacted for possible inclusion in the study declined to participate, 3) only a small number of samples have actually been analyzed for PFOA, and 4) no other data such as age, sex, or other demographic information are available on the donors.

Individual blood samples from 3 different age populations were recently analyzed for PFOA and other fluorochemicals using high pressure liquid chromatography/ electrospray tandem mass spectrometry ( HPLC/ ESMSMS) ( Olsen, et al., 2002a, 2002b, 2002c). The studies' participants included adult blood donors, an elderly population participating in a prospective study in Seattle, WA, and children from 23 states participating in a clinical trial. Overall, the PFOA geometric means were similar across all 3 populations ( 4.6 ppb, 4.2 ppb, and 4.9 ppb, respectively). The geometric means and 95% tolerance limits ( the proportion of the population expected to be found) and their upper bounds were comparable across all 3 studies. However, the upper ranges for the children and adults were much higher than for the elderly population. It is not clear whether this is the result of geographic differences in PFOA levels or some other factor. It should be noted that PFOS and PFOA were highly correlated in all three studies ( r = .63, r = .70, and r = .75) and that PFOA did not meet the criteria for a log normal distribution based on the Shapiro Wilk test in any of the studies. The authors suggest that it may be due to the greater 20 proportion of subjects with values < LLOQ; however, only geometric means were reported. The details of each study are provided below.

Blood samples from 645 U. S. adult blood donors ( 332 males, 313 females), ages 20 69, were obtained from six American Red Cross blood banks located in: Los Angeles, CA; Minneapolis/ St. Paul, MN; Charlotte, NC; Boston, MA; Portland, OR, and Hagerstown, MD ( Olsen, et al., 2002a). Each blood bank was requested to provide approximately 10 samples per 10 year age intervals ( 20 29, 30 39, etc.) for each sex. The only demographic factors known for each donor were age, gender, and location.

The geometric mean serum PFOA level was 4.6 ppb. The range was < lower limit of quantitation ( 1.9) to 52.3 ppb. Males had significantly higher ( p < .05) geometric mean PFOA levels than females ( 37.8 ppb vs. 32.1 ppb). Age was not an important predictor of adult serum fluorochemical concentrations. When stratified by geographic location, the highest geometric mean for PFOA was in the samples from Charlotte, NC ( 6.3 ppb, range: 2.1   29.0) and the lowest from Portland ( 3.6 ppb, range: 2.1   16.7). The highest individual value was reported in Hagerstown ( 52.3 ppb).

Serum PFOA levels were reported for 238 ( 118 males and 120 females) elderly volunteers in Seattle participating in a study designed to examine cognitive function in adults aged 65 96 ( Olsen, et al., 2002b). Age, gender and number of years' residence in Seattle were the only data available on the participants. Most of the participants were under the age of 85 and had lived in the Seattle area for over 50 years.

The geometric mean of PFOA for all samples was 4.2 ppb ( 95% CI, 3.9   4.5). The range was 1.4   16.7 ppb. There was no significant ( p < .05) difference in geometric means for males and females. In simple linear regression analyses, age was negatively ( p < .05) associated with PFOA in elderly men and women. In bootstrap analyses, the mean of the 95% tolerance limit for PFOA was 9.7 ppb with an upper 95% confidence limit of 11.3 ppb. PFOS and PFOA were highly correlated ( r = .75) in this study.

A sample of 598 children, ages 2 12 years old, participating in a study of group A streptococcal infections, was analyzed for serum PFOA levels ( Olsen, et al., 2002c). The samples were collected in 1994 1995 from children residing in 23 states and the District of Columbia. PFOA did not meet the criteria for a log normal distribution based on the Shapiro Wilk test. The authors suggest that it may be due to the greater proportion of subjects with values < LLOQ for PFOA. However, only geometric means were reported. The geometric mean of PFOA for all of the participants was 4.9 ppb ( 95% CI, 4.7   5.1). The range was 1.9 to 56.1 ppb. Male children had significantly ( p<. 01) higher geometric mean serum PFOA levels than females: 5.2 ppb and 4.7 ppb, respectively. In simple linear regression analyses, age was significantly ( p < .05) negatively associated with PFOA in both males and females. When stratified by age, the geometric mean of PFOA was highest at age 4 ( 5.7 ppb) and lowest at age 12 ( 3.5 ppb). Although the data were not reported, a graphical presentation of log PFOA levels for each state by gender looked similar across the states; however, it is difficult to interpret these data without 21 the data and given the limited sample size for each gender/ location subgroup. In bootstrap analyses, the mean of the 95% tolerance limit for PFOA was 10.1 ppb with an upper 95% confidence limit of 11.0 ppb. PFOS and PFOA were highly correlated ( r = .70) in this study. PFOA and PFHS ( perfluorohexanesulfonate) were also correlated ( r = .48).

The above 3 studies indicate similar geometric means and ranges of PFOA among sampled adults, children, and an elderly population. However, an unexpected finding was the level of PFHS and M570 ( N methyl perfluorooctanesulfonamidoacetate) in children. These serum levels were much higher in the sampled children than in the sampled adults or elderly. It is not clear why this occurred, but it is probably due to a different exposure pattern in children.

In another study, the PFOA concentration was analyzed in human sera and liver samples ( Olsen et al., 2001d). Thirty one donor samples were obtained from 16 males and 15 females over an 18 month period from the International Institute for the Advancement of Medicine ( IIAM). The average age of the male donors was 50 years ( SD 15.6, range 5 69) and the average age of the female donors was 45 years ( SD 18.5, range 13 74). The causes of death were intracranial hemorrhage ( n = 16 or 52%), motor vehicle accident ( n = 7 or 23%), head trauma ( n = 4 or 13%), brain tumor ( n = 2 or 6%), drug overdose ( n = 1 or 3%) and respiratory arrest ( n = 1 or 3%). Both serum and liver tissue were obtained from 23 donors; 7 donors contributed liver tissue only and 1 donor contributed serum only. Serum samples were obtained from 5 ml of blood; liver samples consisted of 10 g of tissue. Samples were frozen at IIAM and shipped frozen to 3M for analysis. Samples were extracted using an ion pairing extraction procedure and were quantitatively assayed using HPLC ESMSMS and evaluated versus an unextracted curve. Extensive matrix spike studies were performed to evaluate the precision and accuracy of the extraction procedure. Serum values for PFOA ranged from < LOQ (< 3.0)   7.0 ng/ mL. Assuming the midpoint value between zero and LOQ serum value for samples < LOQ, the mean serum PFOA level was 3.1 ng/ mL with a geometric mean of 2.5 ng/ mL. No liver to serum ratios were provided because more than 90% of the individual liver samples were < LOQ. 22 Table 2. Serum PFOA Levels in Human Populations * Geometric mean and 95% confidence intervals were not included in the reports. ** PFOA detected in about 1/ 3 of the pooled samples but quantifiable in only 2 *** only 4 employees were above the detection limit of 10 ppb Occupational Exposures   Serum Levels ( ppm)

Plant Arithmetic Mean Range Geometric Mean 95% CI Cottage Grove Plant 1997 ( n = 74) 1995 ( n = 80) 1993 ( n = 111) 6.4 6.8 5.0 0.1   81.3 0.0   114.1 0.0   80.0 * * * * * * Decatur Plant 2000 ( n = 263) 1998 ( n = 126) 1997 ( n = 84) 1995 ( n = 90) 1.78 1.54 1.57 1.46 0.04   12.70 0.02   6.76 not reported not reported 1.13 0.90 * * 0.99   1.30 0.72   1.12 * * Antwerp Plant 2000 ( n = 258) 1995 ( n = 93) 0.84 1.13 0.01   7.04 0.00   13.2 0.33 * 0.27   0.40 * Building 236 2000 ( n = 45) 0.106 0.008   0.668 0.053 0.037   0.076 General Population Exposures   Serum Levels ( ppb) Source Arithmetic Mean Range Geometric Mean 95% CI Pooled samples Commercial sources of blood, 1999 ( n = 35 lots) 3 1  13 * *

Blood Banks ( n = 18), 1998 ~ 340 680 donors 17** 12  22 * * Individual samples American Red Cross blood banks, 2000 ( n = 645) 5.6 1.9  52.3 4.6 4.3   4.8 Elderly ( ages 65 96), 2000 ( n = 238) not reported 1.4   16.7 4.2 3.9   4.5 Children ( ages 2 12), 1995 ( n= 598) 5.6 1.9  56.1 4.9 4.7   5.1 3M Corporate managers/ staff St. Paul, MN, 1998 ( n = 31) 12.5*** not reported * * 23 3.0 Human Health Hazards 3.1. Metabolism and Pharmacokinetics 3.1.1 Half life in Humans There are very limited data on the half life of PFOA. With the exception of a 1980 study in which total organic fluorine in blood serum was measured in one worker, no other data were available until June 2000 ( Ubel et al., 1980). A half life study on 27 retirees from the Decatur and Cottage Grove 3M plants was undertaken, in which serum samples were drawn every 6 months over a 5 year period. Two interim reports describing the results thus far have been submitted ( Burris et al., 2000; Burris et al., 2002). The first interim report suggested a median serum half life of PFOA of 344 days, with a range of 109 to 1308 days. The two highest halflife calculations were for the 2 female retirees who participated in this study ( 654 and 1308 days).

There were several limitations to this first analysis including: 1) the limited data available and the range of serum PFOA levels measured; 2) serum was analyzed after each collection period with only one measurement per time period on different days using slightly different analytical techniques; and 3) the reference material purity was not determined until after the first 3 samples had been analyzed. An effort was made to minimize experimental error, including systematic and random error in the analytical method, involving 9 of the original 27 subjects. Serum samples were collected from each of the subjects over 4 time periods spanning 180 days, measured in triplicate with all time points from each subject analyzed in the same analytical run. This would allow for statistical evaluation of the precision of the measurement and assure that all systematic error inherent in the assay equally affected each sample used for half life determination.

Of the 9 retirees included in this analysis, there were 7 males and 2 females, all from the Decatur plant. The average age of the retirees was 61 years, the mean number of years worked at Decatur was 27.7 years, and the average number of months retired was 18.9. Average BMI of this group was 27.9. The mean PFOA value at study initiation was 0.72 ppm ( range 0.06   1.84 ppm, SD = 0.64).

The mean serum half life for PFOA was 4.37 years ( range 1.50   13.49 years, SD = 3.53). Only 1 employee had a half life value that exceeded 4.3 years. The 2 females had values of 3.1 and 3.9 years. Age, BMI, number of years worked or years since retirement were not significant predictors of serum half lives in multivariable regression analyses.

This analysis has attempted to reduce experimental error in the determination of a half life for PFOA. However, several issues should be noted. First, the effect of continued nonoccupational low level exposure on the half life is unknown. Second, because subjects' blood contained concentrations of fluorochemicals that varied by a factor of 30, the data cannot be pooled or averaged unless the serum concentration decay curve shows first order kinetics. 24 Third, it is not known if there are interactions between PFOA and other fluorochemicals in the body. Fourth, this estimate is much higher than that reported in lab animals. Fifth, systematic error of the analytical method could be as high as +/ 20% and still satisfy the data quality criteria.

3.1.2 Absorption Studies in Animals APFO is well absorbed following oral and inhalation exposure, and to a lesser extent following dermal exposure. The liver and serum levels have been measured in several subchronic and reproductive toxicity studies. These results are presented in the relevant studies summarized in subsequent sections. Other studies are summarized here.

In female rats, an average of 749 ug or 37% of the fluorine in the administered dose was recovered in the urine within 4.5 hr after PFOA dose ( by stomach intubation 2 ml of an aqueous solution containing 2 mg PFOA) ( Ophaug and Singer, 1980). The quantity of nonionic fluorine recovered in the urine increased to 61% of the dose at 8 hr, 76% at 24 hr, and 89% at 96 hr.

After a single oral dose of 14C PFOA ( mean dose, 11.0 mg/ kg) in solution to groups of three male rats, at least 93% of the total carbon 14 was absorbed at 24 hours ( Gibson and Johnson, 1979). The half life for elimination of total carbon 14 from plasma was 4.8 days.

Following APFO head only inhalation exposure in male rats ( 6 hr/ day, 5 days/ wk for 2 wk to 0, 1, 8 or 84 mg/ m3) concentrations of organofluoride in the blood showed a dose relationship with initial levels of 108 ppm in rats treated at 84 mg/ m3 ( Kennedy et al., 1986). Immediately after the tenth exposure period, the mean organofluoride blood levels were 13 ppm, 47 ppm, and 108 ppm in the 1, 8, and 84 mg/ m3 dose groups.

Subchronic dermal APFO treatment in rats and rabbits ( 10 applications, 5 doses, 2 rest days, 5 doses) with either 0, 20, 200, or 2000 mg/ kg resulted in elevated blood organofluorine levels which increased in a dose related manner ( Kennedy, 1985).

O'Malley and Ebbins ( 1981) conducted a range finding study which indicates significant dermal absorption of PFOA in male and female rabbits. PFOA ( 100 mg/ kg, 1000 mg/ kg, and 2000 mg/ kg in saline slurry) was applied to approximately 40% of the shaved trunk of the animals, which were then fitted with a plastic collar, and the trunk was wrapped with impervious plastic sheeting. The exposure period was 24 hr, 5 days/ week over 14 days. Mortality was 100% ( 4/ 4) in the 2000 mg/ kg group, 75% ( 3/ 4) in the 1000 mg/ kg group and 0% ( 0/ 4) in the 100 mg/ kg group.

A tetrabutyl ammonium salt of perfluorooctanoate in the form of treated fabric and as a liquid formulation was applied dermally to rabbits ( Johnson, 1995b). Liver samples were analyzed at 28 days post dose for total organic fluorine. The results from treated animals were the same as control values. All total organic values were below the practical quantitation limit. Serum levels were also below the practical quantitation limits of the analysis for samples 25 collected at day 1 and 2 after administration of the mixture or the treated fabric. From the pharmacokinetic study ( Johnson, 1995a), it would be unlikely that any extent of absorption could have been detected in this study.

3.1.3 Distribution Studies in Animals PFOA distributes primarily to the liver, plasma, and kidney, and to a lesser extent, other tissues of the body. It does not partition to the lipid fraction or adipose tissue, but does bind to macromolecules in the tissues. There is evidence of enterohepatic circulation of the compound.

Serum and liver concentrations of PFOA were determined in rhesus monkeys in a 90 day oral toxicity study ( Griffith and Long, 1980). In monkeys at the 3 mg/ kg/ day dose, mean serum PFOA was 50 ppm in males and 58 ppm in females. At the same dose, males had 3 ppm and females 7 ppm in liver samples. At 10 mg/ kg/ day doses, male monkeys had a mean serum PFOA of 58 ppm and females 75 ppm. Liver levels were 9 and 10 ppm for males and females, respectively, measured as organic fluoride.

Ophaug and Singer ( 1980) measured ionic fluoride and total fluorine in the serum of female rats following the administration of PFOA by stomach intubation ( 2 ml of an aqueous solution containing 2 mg PFOA). Serum from rats 4.5 hr after the administration of PFOA had a nonionic fluorine level 13.6 ppm and virtually all of this was bound to components in the serum and not ultrafilterable. Despite the large increase in nonionic fluorine in the serum, the ionic fluoride level remained very low ( 0.03 ppm). Prior to intubation of PFOA, the ionic and nonionic fluorine levels in serum were 0.032 and 0.07 ppm, respectively. The nonionic fluorine level in the serum decreased to 11.2 ppm at 8 hr, 0.35 ppm at 24 hr, and 0.08 ppm at 96 hr. The authors conclude that PFOA is rapidly absorbed from the gastrointestinal tract and rapidly cleared from the serum.

Twenty four hours after oral administration of APFO ( 2 mg APFO in 2 ml aqueous solution by stomach intubation), female rats had a mean serum nonionic fluorine level of 0.35 ppm, while male rats had a mean serum nonionic fluorine level of 44.0 ppm ( Hanhijarvi et al., 1982). APFO was bound to a similar extent in the plasma of male and female rats ( 97.5% bound).

In male and female rats administered 14C PFOA in propylene glycol/ water ( 9.4 umol/ kg, i. p.), the concentration of 14C PFOA derived radioactivity in the blood was higher and eliminated more slowly in males ( t1/ 2= 9 days, males vs 4 hr, females, Vanden Heuvel et al., 1991). In the male rats, the liver had the highest PFOA concentration ( 21% of dose at 2 hr, 2% of dose at 28 days) followed by the plasma and kidney. Far lower PFOA concentrations were found in the heart, testis, fat, and gastrocnemius muscle. In females at 2 hr post dose, the highest concentrations of PFOA were found in the plasma followed by the kidney, liver and ovaries in that order. The average t1/ 2 for elimination of PFOA from the liver in male rats was 11 days compared to an average of 9 days for extrahepatic tissues. In females, the average t1/ 2 for tissue elimination was approximately 3 hr. 26 Vanden Heuvel et al. ( 1991) investigated the disposition of PFOA in perfused male rat liver. Approximately 11% of the cumulative dose of 14C PFOA infused ( 0.08 umol/ min x 48 min, 3.84 umol total) was extracted by the liver during a first pass. In addition, the cumulative percent of PFOA extracted by the liver at 2 min ( 33%) was substantially greater than that seen after 48 min ( 11%) indicating that first pass hepatic uptake of PFOA may be saturable.

Ylinen et al. ( 1990) studied the difference between male and female Wistar rats in the distribution and accumulation of PFOA after a single and subchronic administration. The single dose of PFOA ( 50 mg/ kg in propylene glycol water mixture, 1: 1, vol. 0.25 ml/ 100g) was administered intraperitoneally to 10 week old rats ( 20 male, 20 female). Subchronic administration of PFOA consisted of 3, 10, and 30 mg/ kg/ day by gavage ( in 0.9% NaCl, 0.5 ml/ 100g) to newly weaned rats ( 18 male, 18 female). After the single dose, samples were collected for PFOA determination 12, 24 168 ( at 24 hr intervals), 244 and 336 hours after the administration, and in the subchronic test on the 28th day. The serum was collected by cardiac puncture; after decapitation the brain and at necropsy samples from the liver, kidney, lung, spleen, ovary, testis, and adipose tissue were collected and frozen. The biological half life of PFOA in the serum and tissues was determined from the linear relationship between time and PFOA concentration in the semilogarithmic plot. In the single dose study, concentration of PFOA in the serum and tissues was higher in males than females at all time periods. Twelve hours after the administration of PFOA about 10% of the dose was found in the serum of females, whereas about 40% was in the serum of males. After 14 days about 3.5% of the dose remained in the serum. In females, PFOA concentration in the serum, liver, and kidney occurred in a discontinuous fashion, indicating distinct phases. The half life in the serum was 24 and 105 h in the females and males, respectively. In the females, a half life of 60 hr was estimated in the liver during the first week. In the males, the half life in liver was 210 hr. Although PFOA was retained by the liver, it was not found in the lipid fraction. In the kidney, the half life was 145 hr and 130 hr in females and males, respectively. In the spleen, the half life was 73 hr and 170 hr in females and males, respectively. PFOA was also found in brain tissue. PFOA was not detectable in adipose tissue. In the subchronic study, samples taken on the 28th day indicated significantly higher PFOA concentrations in the serum and tissues of males versus females in all three dose levels. After subchronic, as well as single dose administration, PFOA was mainly distributed in the serum of rats. High concentrations of PFOA were also found in the liver, kidney, and lung of males and females. At the high dose level ( 30 mg/ kg/ day), females and males exhibited, respectively, serum concentrations of 13.92 and 51.65 ug/ ml, liver concentrations of 6.64 and 49.77 ug/ g, kidney concentrations of 12.54 and 39.81 ug/ g, spleen concentrations of 1.59 and 4.10 ug/ g, lung concentrations of 0.75 and 23.71 ug/ g, and brain concentrations of 0.044 and 0.710 ug/ g. The ovary contained 1.16 ug/ g and the testis contained 7.22 ug/ g. A significant positive correlation existed between the administered dose and the concentration of PFOA in the liver, kidney, spleen, and lung of females. On the contrary, no significant correlation between the administered dose and the concentration of PFOA was observed in the males, as 10 mg/ kg/ day produced higher PFOA concentrations in the serum and organs than 30 mg/ kg/ day. However, in males, the concentration in the spleen, testis, and brain correlated positively with the concentration in the serum. 27 Vanden Heuvel et al. ( 1992) demonstrated that PFOA covalently binds to proteins in the liver, plasma, and testes of rats. Carbon 14 labeled PFOA was administered to six week old male Harlan Sprague Dawley rats in propylene glycol/ water ( 1: 1, v/ v; 1 ml/ kg) at a dose of 9.4 umol/ kg, i. p. No time dependent changes in either absolute or relative concentrations of covalently bound PFOA derived 14C were found at 2 h, 1 and 4 days post treatment. Covalently bound PFOA was represented by 0.1 to 0.3% of the tissue 14C content. The absolute concentration of covalently bound PFOA was significantly higher in the plasma than in the liver. The testes had the highest relative concentration of PFOA derived radioactivity covalently bound. In tests, covalent binding of 14C PFOA to a constant concentration of albumin ( 8 uM) increased in a linear fashion with increasing PFOA concentration. The covalent binding of PFOA to hemoglobin was diminished by the addition of cysteine but not methionine, suggesting that protein sulfhydryl groups may be involved.

Hanhijarvi et al. ( 1987) compared the disposition of PFOA between male and female Wistar rats during subchronic administration. PFOA was administered by gavage to 48 newly weaned animals at 0, 3, 10, and 30 mg/ kg ( in 0.9% NaCl, 0.5ml/ 100g) for 28 consecutive days. Urine was collected on the 7th and 28th day of the study ( discussed below). At the end of the study, blood was collected via cardiac puncture. At each dose level, the mean PFOA concentrations in the plasma of the male rats were significantly higher than those of the female rats. The mean plasma PFOA concentrations for the male rats were 48.6+ 26.5 ug/ ml ( dosed at 3 mg/ kg), 83.1+ 24.7 ug/ ml ( 10 mg/ kg), and 53.4+ 11.2 ug/ ml ( 30 mg/ kg). The corresponding figures for female rats were 2.43+ 5.96 ug/ ml, 11.3+ 8.59 ug/ ml, and 9.06+ 8.80 ug/ ml in the same order. The PFOA concentrations in the plasma of the male animals suggested that the binding sites of PFOA may become saturated at the chronic daily dose level of 30 mg/ kg. Although the plasma PFOA concentrations were significantly higher in the male rats, no significant histopathological differences between the sexes were observed at necropsy.

The disposition of PFOA was studied in male Wistar rats after castration and estradiol administration as well as in intact males and females ( Ylinen et al., 1989). The male rats ( N= 20) were castrated at the age of 28 days and after 5 weeks were used in the tests. Half of the operated and 10 intact males were administered estradiol valerate subcutaneously 500 ug/ kg every second day during 14 days before the test. Blood samples were collected by cardiac puncture. At the end of the test ( 96 hr), the concentration of PFOA in the serum of intact males was considerably higher ( 17 40 times) than in the serum of other groups. There was no statistically significant difference in the serum concentrations between the other groups. PFOA was similarly bound to the proteins in the serum of males and females.

Johnson et al. ( 1984) investigated the effect of feeding cholestyramine to rats on the fecal elimination of APFO. Since APFO exists as an anion at physiologic pH, it would be expected to complex with cholestyramine. Ten male Charles River CD rats ( 12 weeks old, 300 342 g) were administered ammonium 14C perfluorooctanoate ( 2.1 mg/ ml) dissolved in 0.9% NaCl as a single intravenous dose ( 2 ml/ rat, average APFO dose 13 mg/ kg). Five rats were given 4% cholestyramine in feed. Urine and feces samples were collected at intervals for 14 days, at which time the animals were sacrificed and liver samples were collected. At 14 days post dose, 28 the mean percentage of PFOA dose eliminated in the feces of cholestyramine treated rats ( 43.2+ 5.5) was 9.8 fold the mean percentage of dose eliminated in feces by untreated rats ( 4.4+ 1.0). Excretion in urine was 41% for treated rats and 67% for untreated rats. Carbon 14 present in the liver represented 12.1+ 2.1 ug eq/ g and 22.3+ 6.2 ug eq/ g in treated and untreated rats, respectively ( 4% and 8% of dose, respectively). In plasma, the levels were 5.1+ 1.7 ug eq/ ml and 14.7+ 6.8 ug eq/ ml in treated and untreated rats, respectively. In red blood cells, the levels were 1.8+ 0.7 ug eq/ ml and 4.2+ 2.4 ug eq/ ml in treated and untreated rats, respectively. The high concentration of 14C APFO in liver at 2 weeks after dosing and the fact that cholestyramine treatment enhances fecal elimination of carbon 14 nearly 10 fold suggests that there is enterohepatic circulation of PFOA.

The disposition of the tetrabutyl ammonium salt of perfluorooctanoic acid in female rabbits has been reported ( Johnson, 1995a). Individual rabbits were given intravenous doses at 0, 4, 16, and 24 mg/ kg and appeared normal throughout the study ( the animal treated at the 40 mg/ kg dose level died within 5 minutes of dosing). Serum samples were analyzed for total organic fluorine at 2, 4, 6, 8, 12, 24, and 48 hours post dose. At 2 hrs, serum organic fluorine levels in the 0, 4, 16, and 24 mg/ kg dosed rabbits were 1.25 ppm, 4.09 ppm, 14.9 ppm, and 41.0 ppm, respectively. There was a rapid decrease in serum level of total organic fluorine with time, non detectable at 48 hr. The biological half life was on the order of 4 hours. The total organic fluorine in whole liver at 48 hr post dose for control animals, 4 mg/ kg, 16 mg/ kg, and 24 mg/ kg intravenous doses were 20 ug, 43 ug, 66 ug, and 54 ug.

3.1.4 Metabolism Studies in Animals Vanden Heuvel et al. ( 1991) investigated the metabolism of PFOA in rats administered 14CPFOA ( 9.4 umol/ kg, i. p.). Pooled daily urine samples ( 0 4 days post treatment) and bile extracts analyzed by HPLC contained a single radioactive peak eluting identically to the parent compound. Tissues were taken from rats treated 4, 14, and 28 days previously with 14C PFOA to determine the presence of PFOA containing lipid conjugates. Only the parent compound was present in rat tissues; no PFOA containing hybrid lipids were detected. Fluoride concentrations in plasma and urine before and after PFOA treatment were unchanged, indicating that PFOA does not undergo defluorination.

Ophaug and Singer ( 1980) also found no change in ionic fluoride level in the serum or urine following oral administration of PFOA to female 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 rats.

3.1.5 Elimination Studies in Animals There are major gender differences in the elimination of PFOA in rats. The biological half life of PFOA in male rats is many times greater than that in female rats and this difference is primarily due to low renal clearance in male rats. The rapid excretion of PFOA by female rats is due to active renal tubular secretion ( organic acid transport system); this renal tubular secretion 29 is believed to be hormonally controlled since castrated male rats treated with estradiol have excretion rates of PFOA similar to those of female rats. Hormonal changes during pregnancy do not appear to change the rate of elimination in rats. This gender difference has not been observed in primates and humans. The studies demonstrating this are described below.

The urine is the major route of excretion of PFOA in the female rat, while the urine and feces are both major routes of excretion of PFOA in male rats ( Vanden Heuval, 1991). Male and female rats were administered 14C PFOA in propylene glycol/ water ( 9.4 umol/ kg, i. p.). Female rats eliminated PFOA derived radioactivity rapidly in the urine with 91% of the dose being excreted in the first 24 hr, while male rats excreted only 6% of the dose in that time period. Negligible radioactivity was recovered in the feces of female rats. In male rats during the 28 day collection period the cumulative excretion of PFOA derived 14C in urine and feces was 36.4% and 35.1%, respectively. The female rat retained less than 10% of the administered dose after 24 hr, while the male rats retained 30% of the administered dose after 28 days. The whole body elimination half life in females was less than one day, and in males it was 15 days. In renal ligated rats injected i. p. with 14C PFOA, approximately 0.3% of the PFOA derived radioactivity was excreted in the bile after 6 hr ( Vanden Heuvel et al., 1991). No sex related difference in the biliary excretion of PFOA was observed when the kidneys were ligated.

Johnson and Gibson ( 1980) observed a sex difference in extent and rate of excretion of total carbon 14 between male and female rats after a single iv dose ( mean dose: female, 16.7 mg/ kg; male 13.1 mg/ kg) of 14C PFOA. Female rats excreted essentially all of the dose via urine in 24 hours while at the same time period male rats excreted only 20 percent of the dose; male rats excreted 83% via urine and 5.4% via feces by 36 days post dose. No radioactivity was detected in tissues of female rats at 17 days post dose; male rats had 2.8% of the dose in liver and 1.1% in plasma at 36 days post dose with lower levels (< 0.5% of the dose) in other organs.

Ophaug and Singer ( 1980) investigated the metabolic fate of PFOA in female Holtzman rats. Animals weighing approximately 250 g were administered by stomach intubation 2 ml of an aqueous solution containing 2 mg PFOA. The animals were then placed in metabolism cages and provided rat chow and tap water for 4.5, 8, 24, or 52.5 hr. In addition, four rats were placed in metabolism cages and fed a low fluoride (< 0.5 ppm) diet and distilled water for a period of 96 hr. At the end of the experimental period the urine, feces and serum were collected. Within 4.5 hr after PFOA dose, an average of 749 ug or 37% of the fluorine in the administered dose was recovered in the urine. The quantity of nonionic fluorine recovered in the urine increased to 61% of the dose at 8 hr, 76% at 24 hr, and 89% at 96 hr. Urinary excretion of ionic fluoride in the PFOA dosed animals was not significantly different than that of the control animals. Fecal excretion of nonionic fluorine was 4.5% of the administered dose at 52.5 hr and 14.3% at 96 hr. The urine from undosed animals contained no detectable nonionic fluorine.

The urinary excretion of APFO in rats was investigated by Hanhijarvi et al. ( 1982). Four male and six female Holtsman rats were administered 2 mg APFO in 2 ml aqueous solution by stomach intubation. Seven female rats were administered 2 ml distilled water as controls. The animals were then placed in metabolism cages with rat chow and tap water. Urine was collected 30 until animals were sacrificed at 24 h by cardiac puncture. Serum was collected. Ionic fluoride and total fluorine content of serum and urine was determined, and nonionic fluorine was calculated as the difference. For clearance studies of APFO and inulin, the rats were anesthetized with Inactin. The femoral artery was cannulated for continuous infusion of 5% mannitol in isotonic saline and the femoral artery was cannulated for drawing blood samples. The urinary bladder was also cannulated for serial collections of urine. Intravenous priming doses of 5.2 5.6 mg [ 1 14C] ammonium perfluorooctanoate ( sp act 0.5 uCi/ mg) and 8.8 ug tritiated inulin ( methoxy 3H, sp act 114 uCi/ mg) were given to each animal. The radiolabled inulin and APFO in 5% mannitol in isotonic saline was then infused at a rate of 0.21 ml/ min. An additional 0.42 0.63 mg/ hr 14C APFO and 9.6 ug/ hr tritiated inulin was infused during the experiments. When the urine and serum collections for the clearance study were complete, probenecid was administered ( 65 68 mg/ kg, ip) and additional clearance tests were performed. In the cumulative excretion study, rats were dosed iv with a mixture of radiolabeled APFO ( 10 20%) and unlabeled APFO ( 80 90%). Five percent mannitol in isotonic saline was infused at a rate of 0.081 ml/ min and urine specimens were collected over 30 min intervals. The effect of probenecid was assessed by administering 65 68 mg/ kg ip at least 30 min prior to the administration of APFO. Twenty four hours after oral administration of APFO, female rats had excreted 76+ 2.7% of the dose in the urine and had a mean serum nonionic fluorine level of 0.35+ 0.11 ppm, while male rats had excreted only 9.2+ 3.5% of the dose and had a mean serum nonionic fluorine level of 44.0+ 1.7 ppm. APFO was bound to a similar extent in the plasma of male and female rats ( 97.5+ 0.25% bound). The clearance studies demonstrated major differences between the sexes in rats. The APFO clearance in female rats was several times greater than the inulin clearance. Administration of probenecid, which strongly inhibits the renal active secretion of organic acids, reduced APFO/ inulin clearance ratio in females from 14.5 to 0.46. APFO clearance was reduced from 5.8 to 0.11 ml/ min/ 100g. Net APFO excretion was reduced from 4.6 to 0.13 ug/ min/ 100g. In male rats, however, the APFO/ inulin clearance ratio and the net excretion of APFO were virtually unaffected by probenecid. In the males, APFO clearance was 0.17 ml/ min/ 100g, APFO/ inulin clearance ratio was 0.22, and net APFO excretion was 0.17 ug/ min/ mg. In the cumulative excretion studies, female rats excreted 76% of the APFO dose, while males excreted only 7.8% of the dose over a 7 hr period. Probenecid administration modified the cumulative excretion curve for males only slightly. However, in females probenecid markedly reduced APFO elimination to 11.8%. The authors concluded that the female rat possesses an active secretory mechanism which rapidly eliminates APFO from the body. This secretory mechanism is lacking or is relatively inactive in male rats and accounts for the greater toxicity of APFO in male rats.

Hanhijarvi et al. ( 1987) compared the urinary elimination of PFOA between male and female Wistar rats during subchronic administration. APFO was administered by gavage to 48 newlyweaned animals at 0, 3, 10, and 30 mg/ kg ( in 0.9% NaCl, 0.5ml/ 100g) for 28 consecutive days. Urine was collected on the 7th and 28th day of the study. At the end of the study, blood was collected via cardiac puncture. At necropsy, tissue specimens for histopathologic examination were collected from the controls and from the group receiving 30 mg/ kg/ day PFOA. On the seventh day of the study period, the female rats in the lowest dose group ( 3 mg/ kg/ day) exhibited significantly greater urinary PFOA excretion than the males ( 3.12+ 0.30 vs 1.50+ 0.57 31 mg/ 24hr/ kg). Unlike the female rats, on the 7th day of the study all three groups of male rats excreted significantly less PFOA than their daily dose of PFOA, which suggested that the males had not reached a steady state by seven days. On the 28th day, the males excreted an amount of PFOA equal to their daily dose.

Hanhijarvi et al. ( 1988) investigated the excretion kinetics of PFOA in the beagle dog. Six laboratory bred beagle dogs ( 3 male, 3 female) were anesthetized with methoxyflurane and catheters were placed in both ureters after laparototomy and cystotomy. The animals were given an intravenous dose of 30 mg/ kg of PFOA followed by continuous infusion with 5% mannitol solution at 1.7 ml/ min. Urine was collected at 10 minute intervals for 60 min. A 5 ml blood sample was collected in the middle of each urine sampling period. Probenicid ( 30 mg/ kg i. v.) was then administered, and urine and blood samples were again collected as before. Renal clearance of PFOA was calculated for the before and after probenecid injection periods. Four additional dogs ( 2male, 2 female) were given 30 mg/ kg PFOA intravenously. These dogs were kept in metabolism cages, and blood samples were collected intermittently for 30 days. Renal clearance rate was approximately 0.03 ml/ min/ kg. Probenecid significantly reduced the PFOA clearance in both sexes, indicating an active secretion mechanism for PFOA. The plasma halflife of PFOA was longer in the male dogs ( 473 h and 541 h) than in the female dogs ( 202 h and 305 h).

The urinary excretion of PFOA was studied in male Wistar rats after castration and estradiol administration as well as in intact males and females ( Ylinen et al., 1989). The male rats ( N= 20) were castrated at the age of 28 days and after 5 weeks were used in the tests. Half of the operated and 10 intact males were administered estradiol valerate subcutaneously 500 ug/ kg every second day during 14 days before the test. Urine was collected in metabolism cages during 96 hr after a single intraperitoneal PFOA dose ( 50 mg/ kg). Blood samples were collected by cardiac puncture. Castration and administration of estradiol to the male rats had a significant stimulatory effect on the urinary excretion of PFOA. During the first 24 hours, female rats excreted 72+ 5% ( N= 6) of the dose, whereas the intact males excreted only 9+ 4% ( N= 6). After the estradiol treatment, both the intact and castrated males excreted PFOA in amounts similar to females ( 61+ 19% and 68+ 14%, respectively). The castrated males without estradiol treatment excreted PFOA in urine faster than the intact males ( 50+ 13%), but less than the females and the estrogen treated males. At the end of the test ( 96 hr), the concentration of PFOA in the serum of intact males was considerably higher ( 17 40 times) than in the serum of other groups. There was no statistically significant difference in the serum concentrations between the other groups. PFOA was similarly bound by the proteins in the serum of males and females.

Vanden Heuvel et al. ( 1992a) investigated whether androgens or estrogens are involved in the marked sex differences in the urinary excretion of PFOA. Castration of males greatly increased (> 1 fold) the elimination of 14C PFOA ( 9.4 umol/ kg, i. p.) in urine, demonstrating that a factor produced by the testis is responsible for the slow elimination of PFOA in male rats. Castration plus 17B estradiol had no further effect on PFOA elimination whereas castration plus testosterone replacement at the physiological level reduced PFOA elimination to the same level as rats with intact testis. Thus, in male rats, testosterone exerts an inhibitory effect on renal 32 excretion of PFOA. In female rats, neither ovariectomy or ovariectomy plus testosterone affected the urinary excretion of PFOA, demonstrating that the inhibitory effect of testosterone on PFOA renal excretion is a male specific response. Probenecid, which inhibits the renal transport system, decreased the high rate of PFOA renal excretion in castrated males but had no effect on male rats with intact testis.

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 ammonium 14C PFOA ( Gibson and Johnson, 1983). At 8 or 9 days after conception, four pregnant rats and 2 nonpregnant female rats were dosed ( mean dose, 15 mg/ kg) and individual urine samples were collected at 12, 24, 36, and 48 hours post dose and analyzed for 14C content. Essentially all of the 14C was eliminated via urine within 24 hours for both groups of rats.

Feeding of cholestyramine to rats enhanced the fecal elimination of APFO ( Johnson et al. ( 1984). Male rats were administered APFO ( 2.1 mg/ ml) dissolved in 0.9% NaCl as a single intravenous dose ( 2 ml/ rat, average APFO dose 13 mg/ kg). At 14 days post dose, the mean percentage of APFO dose eliminated in the feces of cholestyramine treated rats ( 43.2+ 5.5) was 9.8 fold the mean percentage of dose eliminated in feces by untreated rats ( 4.4+ 1.0). Excretion in urine was 41% for treated rats and 67% for untreated rats.

Kudo et al. ( 2002) demonstrated in male and female rats that renal clearance ( CLR) of PFOA and the renal mRNA levels of specific organic anion transporters are markedly affected by sex hormones. The biological half life of PFOA in male rats was found to be 70 times longer than in female rats and this difference is due primarily to low CLR in male rats. 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, which was reduced by estradiol treatment. Treatments of female rats with testosterone reduced the CLR of PFOA. Treatment with probenecid, a known inhibitor of organic anion transporters, markedly reduced the CLR of PFOA in male rats, castrated male rats, and female rats. 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. The level of OAT2 mRNA in male rats was only 13% that in female rats. Castration or estradiol treatment increased the level of OAT2 mRNA whereas treatment of castrated male rats with testosterone reduced it. Ovariectomy of female rats significantly increased the level of OAT3 mRNA. Multiple regression analysis of the data suggested that organic anion transporter 2 ( OAT2) and OAT3 are responsible for urinary elimination of PFOA in the rat.

3.2 Epidemiology Studies 3.2.1 Medical Surveillance Studies from the Antwerp and Decatur Plants A cross sectional analysis of the data from the 2000 medical surveillance program at the Decatur 33 and Antwerp plants was undertaken to determine if there were any associations between PFOA and hematology, clinical chemistries, and hormonal parameters of volunteer employees ( Olsen, et al., 2001e). The data were analyzed for all employees from both plant locations. Mean PFOA serum levels were 1.03 ppm for all male employees at the Antwerp plant and 1.90 ppm for all male employees at the Decatur plant. Male production employees at the Decatur plant had significantly higher ( p < .05) mean serum levels ( 2.34 ppm) than those at the Antwerp plant ( 1.28 ppm). Non production employees at both plants had mean levels below 1 ppm. PFOA serum levels were higher than the PFOS serum values at both plants, especially the Decatur plant where serum levels are higher overall. In addition, values for total organic fluorine were even higher than the PFOA levels.

Multivariable regression analyses were conducted to adjust for possible confounders that may affect the results of the clinical chemistry tests. The following variables were included: production job ( yes or no), plant, age, body mass index ( BMI), cigarettes/ day, drinks/ day and years worked at the plant. A positive significant association was reported between PFOA and cholesterol ( p = .05) and PFOA and triglycerides ( p = .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. When both PFOA and PFOS were included in the analyses, neither reached statistical significance in the cholesterol model, while PFOA remained significant ( p = .02) in the triglycerides model. HDL was negatively associated with PFOA ( p = .04) and remained significant ( p = .04) when both PFOA and PFOS were included in the model. A positive association ( p = .01) between T3 and PFOA was also observed and remained statistically significant ( p = .05) when PFOS was included in the model. BMI, cigarettes/ day, alcohol/ day were also significant in the model. None of the other clinical chemistry, thyroid or hematology measures were significantly associated with PFOA in the regression model.

A longitudinal analysis of the above data and previous medical surveillance results was performed to determine whether occupational exposure to fluorochemicals over time is related to changes in clinical chemistry and lipid results in employees of the Antwerp and Decatur facilities ( Olsen, et al., 2001f). The clinical chemistries included: cholesterol, HDL, triglycerides, alkaline phosphatase, gamma glutamyl transferase ( GGT), aspartate aminotransferase ( AST), alanine aminotransferase ( ALT), total and direct bilirubin. Medical surveillance data from 1995, 1997, and 2000 were analyzed using multivariable regression. 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 34 serum cholesterol ( p = .0008) and triglycerides ( p = .0002) over time. When analyzed by plant and also by subcohort, these associations were limited to the Antwerp employees ( p = .005) and, in particular, the 21 Antwerp employees who participated in all 3 surveillance years ( p = .001). However, the association between PFOA and triglycerides was also statistically significant ( p = .02) for the subgroup in which employees participated in biomonitoring in 1994/ 95 and 2000. There was not a significant association between PFOA and triglycerides among Decatur workers. There were no significant associations between PFOA and changes over time in HDL, alkaline phosphatase, GGT, AST, ALT, total bilirubin, and direct bilirubin.

There are several limitations to the 2000 cross sectional and longitudinal studies including: 1) serum PFOA levels were significantly higher at the Decatur plant than at the Antwerp plant, 2) all participants were volunteers, 3) there were several consistent differences in clinical chemistry profiles and demographics between employees of the Decatur and Antwerp plants ( Antwerp employees as compared to Decatur employees had lower PFOA serum levels, were younger, had lower BMIs, worked fewer years, had higher alcohol consumption, higher mean HDL and bilirubin values, lower mean triglyceride, alkaline phosphatase, GGT, AST, and ALT values, and mean thyroid hormone values tended to be higher), 4) PFOS and other perfluorinated chemicals are also present in these plants, 5) in the cross sectional study, plant populations cannot be compared because they were placed into quartiles based on PFOS serum distributions only which were different for each subgroup and not applicable to PFOA, 6) only one measurement at a certain point in time was collected for each clinical chemistry test, and 7) PFOA serum levels overall have been increasing over time in these employees. In addition, in the longitudinal study only a small number of employees participated in all 3 sampling periods ( 24%), different labs and analytical techniques for PFOA were used each year, and female employees could not be analyzed because of the small number of participants.

3.2.2 Medical Surveillance Studies from the Cottage Grove Plant A voluntary medical surveillance program was offered to employees of the Cottage Grove, Minnesota plant in 1993, 1995, and 1997 ( n = 111, 80 and 74 employees, respectively) ( Olsen, et al., 1998b, Olsen et al., 2000). The clinical chemistry parameters ( cholesterol, hepatic enzymes, and lipoprotein levels) used in the longitudinal and cross sectional studies of the Antwerp and Decatur plants were also used in this study. In addition, in 1997 only, cholecystokinin 33 ( CCK) was also measured at the Cottage Grove plant. CCK levels were observed because certain research has suggested that pancreas acinar cell adenomas seen in rats exposed to PFOA may be the result of increased CCK levels ( Obourn, et al., 1997).

Only male employees involved in PFOA production were included in the study. Sixty eight employees were common to the 1993 and 1995 sampling periods, 21 were common between 1995 and 1997, and 17 participated in all three surveillance years. Mean serum PFOA levels and ranges are provided in Table 2 of the Biomonitoring Section of this report. It should be noted that Cottage Grove has the highest serum PFOA levels of the 3 plants studied.

Employees' serum PFOA levels were stratified into 3 categories (< 1, 1 < 10, and   10 ppm), 35 chosen to provide a greater number of employees in the   10 ppm category. As employees' mean serum PFOA levels increased, no statistically significant abnormal liver function tests, hypolipidemia, or cholestasis were observed in any of the sampling years. Multivariable regression analyses controlling for potential confounders ( age, alcohol consumption, BMI, and cigarettes smoked) yielded similar results. The authors also reported that renal function, blood glucose, and hematology measures were not associated with serum PFOA levels; however, these data were not provided in the paper.

The mean CCK value reported for the 1997 sample was 28.5 pg/ ml ( range 8.8  86.7 pg/ ml). The means in the 2 serum categories < 10 ppm were at least 50% higher than in the   10 ppm category. A statistically significant ( p = .03) negative association between mean CCK levels and the 3 PFOA serum categories was observed. A scatter plot of the natural log of CCK and PFOA shows that all but 2 CCK values are within the assay's reference range of 0  80 pg/ ml. Both of these employees ( CCK values of 80.5 and 86.7 pg/ ml) had serum PFOA levels less than 10 ppm ( 0.6 and 5.6 ppm, respectively). 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.

The cross sectional design is a limitation of this study. Only 17 subjects were common to all 3 sampling years. In addition, the medical surveillance program is a voluntary one. The participation rate of eligible production employees decreased from approximately 70% in 1993 to 50% in 1997. Also, the laboratory reference range changed substantially for ALT in 1997. Finally, different analytical methods were used to measure serum PFOA. Serum PFOA was determined by electrospray high performance liquid chromatography/ mass spectrometry in 1997, but by thermospray in 1993 and 1995.

An earlier medical surveillance study on workers who were employed in the 1980' s was conducted at the Cottage Grove plant; however, total serum fluorine was measured instead of PFOA ( Gilliland and Mandel, 1996). Based on animal studies that reported that animals exposed to PFOA develop hepatomegaly and alterations in lipid metabolism, a cross sectional, occupational study was performed to determine if similar effects are present in workers exposed to PFOA. In a PFOA production facility, 115 workers were studied to determine whether serum PFOA affected their cholesterol, lipoproteins, and hepatic enzymes. Forty eight workers who were exposed to PFOA from 1985 1989 were included in the study ( 96% participation rate). Sixty five employees who either volunteered or were asked to participate, were included in the unexposed group. These employees were assumed to have little or no PFOA exposure based on their job description. However, when serum levels were analyzed, it was noted that this group of workers had PFOA levels much greater than the general population. Therefore, instead of job categories, total serum fluorine was used to classify workers into exposure groups.

Total serum fluorine was used as a surrogate measure for PFOA. Serum PFOA was not measured, due to the cost of analyzing the samples. Blood samples were analyzed for total serum fluorine, serum glutamyl oxaloacetic transaminase ( SGOT or AST), serum glutamyl pyruvic transaminase ( SGPT or ALT), gamma glutamyl transferase ( GGT), cholesterol, low 36 density lipoproteins ( LDL), and high density lipoproteins ( HDL). All of the participants were placed into five categories of total serum fluorine levels: < 1 ppm, 1 3 ppm, > 3  10 ppm, > 10  15 ppm, and > 15 ppm. The range of the serum fluorine values was 0 to 26 ppm ( mean 3.3 ppm). Approximately half of the workers fell into the > 1  3 ppm category, while 23 had serum levels < 1 ppm and 11 had levels > 10 ppm.

There were no significant differences between exposure categories when analyzed using univariate analyses for cholesterol, LDL, and HDL. In the multivariate analysis, there was not a significant association between total serum fluorine and cholesterol or LDL after adjusting for alcohol consumption, age, BMI, and cigarette smoking. There were no statistically significant differences among the exposure categories of total serum fluorine for AST, ALT and GGT. However, increases in AST and ALT occurred with increasing total serum fluorine levels in obese workers ( BMI = 35 kg/ m2). This result was not observed when PFOA was measured directly in serum of workers in 1993, 1995, or 1997 surveillance data of employees of the Cottage Grove plant ( Olsen, et al., 2000).

Since PFOA was not measured directly and there is no exposure information provided on the employees ( eg. length of employment/ exposure), the results of the study provide limited information. The authors state that no adverse clinical outcomes related to PFOA exposure have been observed in these employees; however, it is not clear that there has been follow up of former employees. In addition, the range of results reported for the liver enzymes were fairly wide for many of the exposure categories, indicating variability in the results. Given that only one sample was taken from each employee, this is not surprising. It would be much more helpful to have several samples taken over time to ensure their reliability. It also would have been interesting to compare the results of the workers who were known to be exposed to PFOA to those who were originally thought not to be exposed to see if there were any differences among the employees in these groups. There were more of the " unexposed" employees ( n = 65) participating in the study than those who worked in PFOA production ( n = 48).

3.2.3 Mortality Studies A retrospective cohort mortality study was performed on employees at the Cottage Grove, MN plant which produces APFO ( Gilliland and Mandel, 1993). At this plant, APFO production was limited to the Chemical Division. The cohort consisted of workers who had been employed at the plant for at least 6 months between January 1947 and December 1983. Death certificates of all of the workers were obtained to determine cause of death. There was almost complete follow up ( 99.5%) of all of the study participants. The exposure status of the workers was categorized based on their job histories. If they had been employed for at least 1 month in the Chemical Division, they were considered exposed. All others were considered to be not exposed to PFOA. The number of months employed in the Chemical Division provided the cumulative exposure measurements. Of the 3537 ( 2788 men and 749 women) employees who participated in this study, 398 ( 348 men and 50 women) were deceased. Eleven of the 50 women and 148 of the 348 men worked in the Chemical Division, and therefore, were considered exposed to PFOA. 37 Standardized Mortality Ratios ( SMRs), adjusted for age, sex, and race were calculated and compared to U. S. and Minnesota white death rates for men. For women, only state rates were available. The SMRs for males were stratified for 3 latency periods ( 10, 15, and 20 years) and 3 periods of duration of employment ( 5, 10, and 20 years).

For all female employees, the SMRs for all causes and for all cancers were less than 1. The only elevated ( although not significant) SMR was for lymphopoietic cancer, and was based on only 3 deaths. When exposure status was considered, SMRs for all causes of death and for all cancers were significantly lower than expected, based on the U. S. rates, for both the Chemical Division workers and the other employees of the plant.

In all male workers at the plant, the SMRs were close to 1 for most of the causes of death when compared to both the U. S. and the Minnesota death rates. When latency and duration of employment were considered, there were no elevated SMRs. When employee deaths in the Chemical Division were compared to Minnesota death rates, the SMR for prostate cancer for workers in the Chemical Division was 2.03 ( 95% CI .55  4.59). This was based on 4 deaths ( 1.97 expected). There was also a statistically significant association with length of employment in the Chemical Division and prostate cancer mortality. Based on the results of proportional hazard models, the relative risk for a 1 year increase in employment in the Chemical Division was 1.13 ( 95% CI 1.01 to 1.27). It rose to 3.3 ( 95% CI 1.02  10.6) for workers employed in the Chemical Division for 10 years when compared to the other employees in the plant. The SMR for workers not employed in the Chemical Division was less than expected for prostate cancer (. 58).

An update of this study was conducted to include the death experience of employees through 1997 ( Alexander, 2001a). The cohort consisted of 3992 workers. The eligibility requirement was increased to 1 year of employment at the Cottage Grove plant, and the exposure categories were changed to be more specific. Workers were placed into 3 exposure groups based on job history information: definite PFOA exposure ( n = 492, jobs where cell generation, drying, shipping and packaging of PFOA occurred throughout the history of the plant); probable PFOA exposure ( n = 1685, other chemical division jobs where exposure to PFOA was possible but with lower or transient exposures); and not exposed to fluorochemicals ( n = 1815, primarily nonchemical division jobs).

In this new cohort, 607 deaths were identified: 46 of these deaths were in the PFOA exposure group, 267 in the probable exposure group, and 294 in the non exposed group. When all employees were compared to the state mortality rates, SMRs were less than 1 or only slightly higher for all of the causes of death analyzed. None of the SMRs were statistically significant at p = .05. The highest SMR reported was for bladder cancer ( SMR = 1.31, 95% CI = 0.42   3.05). Five deaths were observed ( 3.83 expected).

A few SMRs were elevated for employees in the definite PFOA exposure group: 2 deaths from cancer of the large intestine ( SMR = 1.67, 95% CI = 0.02   6.02), 1 from pancreatic cancer 38 ( SMR = 1.34, 95% CI = 0.03   7.42), and 1 from prostate cancer ( SMR = 1.30, 95% CI = 0.03   7.20). In addition, employees in the definite PFOA exposure group were 2.5 times more likely to die from cerebrovascular disease ( 5 deaths observed, 1.94 expected; 95% CI = 0.84   6.03).

In the probable exposure group, 3 SMRs should be noted: cancer of the testis and other male genital organs ( SMR = 2.75, 95% CI = 0.07   15.3); pancreatic cancer ( SMR = 1.24, 95% CI = 0.45   2.70); and malignant melanoma of the skin ( SMR = 1.42, 95% CI = 0.17   5.11). Only 1, 6, and 2 cases were observed, respectively. The SMR for prostate cancer in this group was 0.86 ( 95% CI = 0.28   2.02) ( n = 5).

There were no notable excesses in SMRs in the non exposed group, except for cancer of the bladder and other urinary organs. Four cases were observed and only 1.89 were expected ( 95% CI = 0.58   5.40).

It is difficult to interpret the results of the prostate cancer deaths between the first study and the update because the exposure categories were modified in the update. Only 1 death was reported in the definite exposure group and 5 were observed in the probable exposure group. All of these deaths would have been placed in the chemical plant employees exposure group in the first study. The number of years that these employees worked at the plant and/ or were exposed to PFOA was not reported. This is important because even 1 prostate cancer death in the definite PFOA exposure group resulted in an elevated SMR for the group. Therefore, if any of the employees' exposures were misclassified, the results of the analysis could be altered significantly.

The excess mortality in cerebrovascular disease noted in employees in the definite exposure group was further analyzed based on number of years of employment at the plant. Three of the 5 deaths occurred in workers who were employed in jobs with definite PFOA exposure for more than 5 years but less than 10 years ( SMR = 15.03, 95% CI = 3.02   43.91). The other 2 occurred in employees with less than 1 year of definite exposure. The SMR was 6.9 ( 95% CI = 1.39   20.24) for employees with greater than 5 years of definite PFOA exposure. In order to confirm that the results regarding cerebrovascular disease were not an artifact of death certificate coding, regional mortality rates were used for the reference population. The results did not change. When these deaths were further analyzed by cumulative exposure ( time weighted according to exposure category), workers with 27 years of exposure in probable PFOA exposed jobs or those with 9 years of definite PFOA exposure were 3.3 times more likely to die of cerebrovascular disease than the general population. A dose response relationship was not observed with years of exposure.

It is difficult to compare the results of the first and second mortality studies at the Cottage Grove plant since the exposure categories were modified. Although the potential for exposure misclassification was certainly more likely in the first study, it may still have occurred in the update as well. It is difficult to judge the reliability of the exposure categories that were defined without measured exposures. Although serum PFOA measurements were considered in the exposure matrix developed for the update, they were not directly used. In the second study, the 39 chemical plant employees were sub divided into PFOA exposed groups, and the film plant employees essentially remained in the " non exposed" group. This was an effort to more accurately classify exposures; however, these new categories do not take into account duration of exposure or length of employment. Another limitation to this study is that 17 death certificates were not located for deceased employees and therefore were not included in the study. The inclusion or exclusion of these deaths could change the analyses for the causes of death that had a small number of cases. Follow up of worker mortality at Cottage Grove ( and Decatur) needs to continue. Although there were more than 200 additional deaths included in this analysis, it is a small number and the cohort is still relatively young. Given the results of studies on fluorochemicals in both animals and humans, further analysis is warranted.

3.2.4 Hormone Study Endocrine effects have been associated with PFOA exposure in animals; therefore, medical surveillance data, including hormone testing, from employees of the Cottage Grove, Minnesota plant were analyzed ( Olsen, et al., 1998a). PFOA serum levels were obtained for volunteer workers in 1993 ( n = 111) and 1995 ( n = 80). Sixty eight employees were common to both sampling periods. In 1993, the range of PFOA was 0 80 ppm ( although 80 ppm was the limit of detection that year, so it could have been higher) and 0 115 ppm in 1995 using thermospray mass spectrophotometry assay. Eleven hormones were assayed from the serum samples. They were: cortisol, dehydroepiandrosterone sulfate ( DHEAS), estradiol, FSH, 17 gammahydroxyprogesterone ( 17 HP), free testosterone, total testosterone, LH, prolactin, thyroidstimulating hormone ( TSH) and sex hormone binding globulin ( SHBG). 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.

PFOA was not highly correlated with any of the hormones or with the following covariates: age, alcohol consumption, BMI, or cigarettes. Most of the employees had PFOA serum levels less than 10 ppm. In 1993, only 12 employees had serum levels > 10 ppm, and 15 in 1995. However, these levels ranged from approximately 10 ppm to over 114 ppm. There were only 4 employees in the > 30 ppm PFOA group in 1993 and only 5 in 1995. Therefore, it is likely that there was not enough power to detect differences in either of the highest categories. The mean age of the employees in the highest exposure category was the lowest in both 1993 and 1995 ( 33.3 years and 38.2 years, respectively). Although not significantly different from the other categories, BMI was slightly higher in the highest PFOA category.

Estradiol was highly correlated with BMI ( r = .41, p < .001 in 1993, and r = .30, p < .01 in 1995). In 1995, all 5 employees with PFOA levels > 30 ppm had BMIs > 28, although this effect was not observed in 1993. Estradiol levels in the > 30 ppm group in both years were 10% higher than the other PFOA groups; however, the difference was not statistically significant. 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 40 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.

Free testosterone was highly correlated with age in both 1993 and 1995. The authors did not report a negative association between PFOA serum levels and testosterone. There were no statistically significant trends noted for PFOA and either bound or free testosterone. However, 17 HP, a precursor of testosterone, was highest in the > 30 ppm PFOA group in both 1993 and 1995. In 1995, PFOA was significantly associated with 17 HP in regression models adjusted for possible confounders. However, the authors state that this association was based on the results of one employee ( data were not provided in the report). There were no significant associations between PFOA and cortisol, DHEAS, FSH, LH, and SHBG.